http://informahealthcare.com/txc ISSN: 1040-8444 (print), 1547-6898 (electronic) Crit Rev Toxicol, 2014; 44(6): 467–498 © 2014 Informa Healthcare USA, Inc. DOI: 10.3109/10408444.2013.875983

REVIEW ARTICLE

Reproductive and developmental effects of phthalate diesters in males

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Vanessa R. Kay1, Michael S. Bloom2, and Warren G. Foster1 1Department of Obstetrics and Gynecology, McMaster University, Hamilton, ON, Canada, and 2Departments of Environmental Health Sciences &

Epidemiology and Biostatistics, University at Albany, School of Public Health, Rensselaer, NY, USA

Abstract

Keywords

Phthalate diesters are a diverse group of chemicals used to make plastics flexible and are found in personal care products, medical equipment, and medication capsules. Ubiquitous in the environment, human exposure to phthalates is unavoidable; however, the clinical relevance of low concentrations in human tissues remains uncertain. The epidemiological literature was inadequate for prior reviews to conclusively evaluate the effects of phthalates on male reproductive tract development and function, but recent studies have expanded the literature. Therefore, we conducted a systematic review of the literature focused on the effects of phthalate exposure on the developing male reproductive tract, puberty, semen quality, fertility, and reproductive hormones. We conclude that although the epidemiological evidence for an association between phthalate exposure and most adverse outcomes in the reproductive system, at concentrations to which general human populations are exposed, is minimal to weak, the evidence for effects on semen quality is moderate. Results of animal studies reveal that, although DEHP was the most potent, different phthalates have similar effects and can adversely affect development of the male reproductive tract with semen quality being the most sensitive outcome. We also note that developmental exposure in humans was within an order of magnitude of the adverse effects documented in several animal studies. While the mechanisms underlying phthalate toxicity remain unclear, the animal literature suggests that mice are less sensitive than rats and potentially more relevant to estimating effects in humans. Potential for chemical interactions and effects across generations highlights the need for continued study.

cryptorchidism, developmental, hypospadias, phthalates, puberty, reproductive, semen, testosterone

Table of Contents Abstract ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Introduction ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Approach to systematic review ... ... ... ... ... ... ... ... ... ... Epidemiological literature ... ... ... ... ... ... ... ... ... ... ... ... Anogenital distance ... ... ... ... ... ... ... ... ... ... ... ... ... ... Cryptorchidism ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Hypospadias ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Puberty ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Semen quality ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Sperm concentration, count, and morphology populations ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Sperm concentration, count, and morphology populations ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Sperm DNA damage ... ... ... ... ... ... ... ... ... ... ... ... ... Fertility and fecundity ... ... ... ... ... ... ... ... ... ... ... ... ... Reproductive hormones ... ... ... ... ... ... ... ... ... ... ... ... Studies in human tissue ... ... ... ... ... ... ... ... ... ... ... ... Animal studies ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Physical development and in utero exposure. ... ... ... ... Sexual maturation... ... ... ... ... ... ... ... ... ... ... ... ... ... ... Semen quality ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ...

... ... ... 467 ... ... ... 467 ... ... ... 468 ... ... ... 470 ... ... ... 470 ... ... ... 472 ... ... ... 473 ... ... ... 474 ... ... ... 475 in general ... ... ... 475 in clinical ... ... ... 475 ... ... ... 476 ... ... ... 477 ... ... ... 478 ... ... ... 480 ... ... ... 481 ... ... ... 481 ... ... ... 485 ... ... ... 485

Address for correspondence: Dr. Warren G. Foster, Department of Obstetrics and Gynecology, HSC-3N52D, McMaster University, 1200 Main Street West, Hamilton, ON L8S 4K1, Canada. Tel: (905) 9529140; Ext: 22822. Fax: (905) 524-2911. E-mail: [email protected]

History Received 5 March 2013 Revised 10 December 2013 Accepted 11 December 2013 Published online 2 June 2014

Testicular morphology ... ... ... Reproductive hormones ... ... Mixtures ... ... ... ... ... ... ... ... ... Perspective ... ... ... ... ... ... ... ... Modes of action ... ... ... ... ... Limitations of animal studies Critical windows of exposure Transgenerational effects ... ... Conclusions ... ... ... ... ... ... ... ... Acknowledgments ... ... ... ... ... Declaration of interest ... ... ... ... References ... ... ... ... ... ... ... ...

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486 487 488 488 488 490 490 490 491 491 491 491

Introduction Reports of a decline in semen quality together with changes in the prevalence of cryptorchidism and hypospadias and rising rates of testicular cancer in young men has led to the suggestion that male reproductive health is under siege (Toppari et al. 2002). The Testicular Dysgenesis Syndrome (TDS) has been advanced as a potential unifying hypothesis (Toppari et al. 2010). According to the TDS, exposure to environmental toxicants during a critical window of development in utero, birth size, and genetic factors can induce dysgenesis of the testis (Toppari et al. 2010). Increased exposure to environmental

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contaminants is the most likely explanation for the increased incidence of male reproductive tract malformations, decreased semen quality, and an increased incidence of testis cancer (Toppari et al. 2002). Environmental contaminants that can interfere with hormone signaling in particular are thought to be central to the development of the TDS. Phthalates are a family of chemical contaminants that have been shown to decrease testosterone synthesis, thereby antagonizing androgen signaling (Skakkebaek et al. 2001) and producing a “phthalate syndrome” in rodent models analogous in many respects to the proposed human TDS (Foster et al. 2006, Gray et al. 2006). Therefore, phthalates have been suggested to play a role in TDS and in the apparent decline in male reproductive health (Toppari et al. 2010). As such, it is imperative to examine the associations between phthalate exposure and male reproductive health in humans. Phthalate diesters, commonly known as phthalates, are used as plasticizers to make plastics flexible or soft (Wittassek et al. 2011). They are a diverse group of compounds derived from phthalic acid through the formation of esters and named based on the length of the alkyl chains. It has been estimated that 11 billion pounds of phthalates are produced globally each year (Lowell 2011). Previously, human contact with di-2-ethylhexyl phthalate (DEHP) was most prevalent, given its frequent use in polyvinyl chloride (PVC). More recently, di-isodecyl phthalate (DiDP) and di-isononyl phthalate (DiNP) have replaced DEHP in PVC products (Wittassek et al. 2011) with the result that contact with these compounds is increasing. Low-molecular weight phthalates, including dibutyl phthalate (DBP), diethyl phthalate (DEP), and dimethyl phthalate (DMP), have shorter alkyl chains and are commonly used in cosmetics and medicinal coatings (Wittassek et al. 2011). Unfortunately, as phthalates are not covalently bound to the plastic, they leach out of the products and enter the environment, becoming ubiquitous contaminants and leading to widespread human exposure (Blount et al. 2000, Calafat and McKee 2006, Koch et al. 2009, Wittassek et al. 2011). Routes of human exposure to phthalates include consumption of contaminated food, use of personal care products and some medications, and environmental exposures including dust in the air and water (Wittassek et al. 2011). After absorption, phthalate diesters are hydrolyzed to biologically active monoesters which can undergo further oxidation before excretion (Barr et al. 2003, Kato et al. 2004). These metabolites can be used to estimate exposure levels when measured in urine (Barr et al. 2003). Importantly, quantifying secondary oxidative metabolites, such as mono2-ethyl-hydroxy-hexyl phthalate (MEHHP) and mono-2ethyl-oxo-hexyl phthalate (MEOHP) from DEHP, can add accuracy to estimates of phthalate exposure compared to solely evaluating the initial metabolite, in this case mono-2 -ethylhexyl phthalate (MEHP) (Barr et al. 2003). Exclusion of these metabolites underestimates exposure (Koch and Calafat 2009). Greater accuracy from the use of the secondary, oxidative metabolites is common with high molecular weight phthalates, such as the aforementioned DEHP, which have longer alkyl chains and for which a preponderance of secondary metabolites are detected in urine relative to primary metabolites (Barr et al. 2003).

Crit Rev Toxicol, 2014; 44(6): 467–498

Estimated average human intakes are 2.32–12 μg/kg/day DEP, 0.26–0.88 μg/kg/day butylbenzyl phthalate (BBP), 0.84–5.22 μg/kg/day DBP, 0.12–1.4 μg/kg/day di-isobutyl phthalate (DiBP), 0.71–4.6 μg/kg/day DEHP, and 0.29 μg/kg/ day DiNP in German and US populations (Koch and Calafat 2009). Importantly, one study using a representative sample of the 1999–2000 US population (US National Health and Nutrition Examination Survey, NHANES) found that concentrations of mono-benzyl phthalate (MBzP), mono-butyl phthalate (MBP) and MEHP, metabolites of BBP, DBP and DEHP respectively, were higher in children than in adults, the former whom are thought to be more susceptible to reproductive effects given ongoing development (Silva et al. 2004). A paucity of epidemiological studies led prior reviews to the conclusion that insufficient human evidence existed to definitively assess causal associations between exposure to certain phthalates and reproductive health outcomes (Kavlock et al. 2002a, 2002b, 2002c, 2002d, 2002e, 2006). However, a large number of epidemiology studies have recently been brought forward in the literature. Therefore we undertook a systematic review of the literature to re-evaluate the strength and consistency of the epidemiologic evidence for causal associations between the human phthalate exposure and male reproductive outcomes. We further reviewed the experimental animal literature for evidence of dose–response characteristics, biological plausibility, reversibility of effects, and potential mechanism(s) of action relevant to human health.

Approach to systematic review Medline was used to search for epidemiological and animal studies relevant to the effect of phthalates on a number of male reproductive health issues, and the topics included development, puberty, and semen quality and searches for each topic were completed separately. The search terms used and the number of articles retrieved were phthalic acid OR phthalates AND testicular diseases OR gonadal dysgenesis OR hypospadias OR testes OR genitalia (n  418); phthalic acid OR phthalates AND semen OR spermatozoa OR fertility (n  169); phthalic acid OR phthalates AND prostatic neoplasms OR testicular neoplasms OR prostate cancer OR testicular cancer (n  128); and phthalic acid OR phthalates AND puberty OR puberty, delayed OR puberty, precocious (n  55). In total, 770 articles published in English, from the search in 2011 were initially retrieved and reviewed. Articles were retained if the outcomes were relevant to male reproductive health and phthalates were measured or tested in mammals. Duplicate (n  112), review (n  27), clearly unrelated (n  262) or inaccessible (n  17) articles, and articles beyond the scope of the review (n  168) were excluded based on the title, leaving a total of 184 papers. Articles that examined phthalates in combination with other compounds were beyond the scope of this review and were thus excluded. Citations of included papers were reviewed in order to identify the other potentially relevant articles, and relevant papers were added when published. The results of the included epidemiological studies were summarized (Table 1) and a weight-of-evidence approach was applied, considering exposure, epidemiological literature, biological plausibility, and dose–response characterization.

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Reproductive toxicity of phthalates in males 469

Table 1. Summary of the epidemiological studies examining potential effects of phthalate exposure on the male reproductive system. Outcome AGD/AGI

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Hypospadias

Cryptorchidism

Puberty

Gynecomastia

Semen Quality

Associations No association with maternal urinary concentrations of MMP, MEP, MBP, MBzP, or MEHP Maternal urinary concentrations of MEHP associated with decreased AGI (r   0.189, p  0.047), no association with MMP, MEP, MBP, MBzP, MEHHP, or MEOHP Maternal urinary concentrations of MBP, MBzP, MEP, MiBP, and combined exposure to phthalates associated with decreased AGI (MBP: OR  10.2 [2.5–42.2], MBzP OR  3.8, MEP OR  4.7, MiBP OR  9.1 with p  0.05 for all), no association with MMP, MEHP, MEHHP, MEOHP, or MCPP Maternal urinary concentrations of MEP, MBP, MEHP, MEHHP, and MEOHP were associated with decreased AGD , with p  0.05 for all No association with maternal (OR  4.2 [0.4–41.3]) or paternal (OR  1.16 [0.93–1.46]) occupational exposures. No significant association with maternal exposure (OR  1.19 [0.82–1.74]) Occupational exposure associated with increased risk (OR  3.12 [1.04–11.46]) No association with occupational exposure (OR  0.90 [0.74–1.10]) No association with maternal exposure (OR  0.3 [0.0–3.1]) No association with metabolites of DMP, DEP, DBP, BBP, DEHP, or DiNP measured in breast milk Increased risk associated with maternal MEHP (OR  2.38, p  0.048) Suggested increases for MEHHP (OR  2.67, p  0.054) and MEOHP (OR  2.55, p  0.064) No associations with MEP, MBP, MiBP, MMP, MBzP, or MCPP. Occupational exposure associated with increased risk (OR  8.3, p  0.038) No association with metabolites of DEP, DBP, DiBP, BBP, DEHP, or DiNP measured in urine No association with a history of exposure to DEHP through extracorporeal membrane oxygenation as an infant DEHP and MEHP concentrations in serum associated with increased risk (DEHP OR  2.77 [1.48–5.21] and MEHP OR  24.76 [3.5–172.6]) No association with metabolites of DEP, DBP, DiBP, BBP, DEHP, or DiNP measured in urine MBP associated with decreased motility (OR  3.0 [1.2–7.6]) and non-significantly with concentration and morphology (OR  3.3 [0.9–12.6] and OR  2.2 [0.8–6.1]) MBzP with decreased concentration (OR  5.5 [1.3–23.9]), and MMP non-significantly associated with morphology (urinary metabolites) MEP associated with more DNA damage (comet extent 3.61 μm [0.74–6.47] and taildistributed moment 1.17 μm [ 0.05–2.38]), no associations with MMP, MBP, MBzP, and MEHP (urinary metabolites) MEP associated with increased motility (straight-line velocity and curvilinear velocity), MEHP, MBP, and MBzP non-significantly associated with decreased motility, no association with MMP (urinary metabolites) MBP associated with decreased motility (OR  1.8, p  0.04) and concentration (OR  3.3 p  0.04), MBzP non-significantly associated with concentration (OR  1.9, p  0.13), non-significant trend with %MEHP and motility, no associations with MMP, MEP, MEHHP, or MEOHP (urinary metabolites) MEP (p  0.02), MBP, MBzP, and MEHP (p  0.0006) associated with increased DNA damage and MMP inversely with DNA damage (urinary metabolites) %MEHP with DNA damage (Tail% 3.06% [1.33–4.79] p  0.0006), %MEHP nonsignificantly linked with motility (urinary metabolites) No associations with DEHP or %MEHP (urinary metabolites) DEHP exposure measured in the air associated with increased DNA damage (DNA denaturation, p  0.015, DNA fragmentation p  0.010) and decreased motility (β  0.227, p  0.044) MEP associated with decreased motility (8.8% fewer motile sperm [0.8–17], no associations with MBP, MBzP, or MEHP (urinary metabolites) MEHHP (p  0.003), MEHP (p  0.001), and MiNP (p  0.033) associated with decreased motility, MBP associated with decreased linear velocity (p  0.007), curvilinear velocity (p  0.009), and with DNA fragmentation (p  0.047), MBzP, MBP, MEHP, and MEP associated with sperm aneuploidy (p  0.05) MBP associated with decreased concentration (OR  12.0, p  0.05), MMP non-significantly linked with decreased concentration, no associations with MEP, MBzP, MEHP, or MEHHP (urinary metabolites) No association between occupational exposure to DEHP and partner’s time-to-pregnancy Semen DBP associated with decreased concentration of sperm Semen DEHP, DBP, and DEP concentrations associated with decreased sperm concentration (r   0.25,  0.20,  0.19 respectively, p  0.05), semen DEHP and DBP with decreased motility (r   0.18 for both p  0.05), semen DEHP with increased abnormal morphology (r  0.18, p  0.05) Semen DEHP and DBP concentrations associated with decreased motility and sperm concentration (p  0.01)

References Huang et al. (2009) Suzuki et al. (2012) Swan et al. (2005)

Swan et al. (2008) Chevrier et al. (2012) Nassar et al. (2010) Ormond et al. (2009) Vrijheid et al. (2003) Chevrier et al. (2012) Main et al. (2006) Swan et al. (2008)

Wagner-Mahler et al. (2011) Mieritz et al. (2012) Rais-Bahrami et al. (2004) Durmaz et al. (2010) Mieritz et al. (2012) Duty et al. (2003a)

Duty et al. (2003b)

Duty et al. (2004) Hauser et al. (2006)

Hauser et al. (2007) Hauser (2008) Herr et al. (2009) Huang et al. (2011) Jonsson et al. (2005) Jurewicz et al. (2013)

Liu et al. (2012)

Modigh et al. (2002) Murature et al. (1987) Pant et al. (2008)

Pant et al. (2011) (Continued)

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Crit Rev Toxicol, 2014; 44(6): 467–498

Table 1. (Continued) Outcome

Associations

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Semen phthalate concentrations associated with decreased normal morphology (r   0.796, p  0.001) and increased percentage of single-stranded DNA (r  0.855, p  0.001)

Reproductive Hormones

MBP and MBzP associated with increased semen concentration (p  0.05 and p  0.01 respectively). No associations with MMP, MEP, MEHP, MEHHP, or MEOHP (urinary metabolites) MEP and MBP higher in men who were in infertile couples (p  0.001 and p  0.008); MEHP  MEHHP and MBzP non-significantly higher (p  0.072, p  0.053) (urinary metabolites) MEP associated with decreased concentration (OR  6.5 [1.0–43.6]) and increased abnormal morphology (OR  3.4 [0.9–13.8]), MCPP associated with increased abnormal morphology (OR  7.6 [1.7–33.3]), MEHP non-significantly associated with decreased concentration (urinary metabolites) No association for semen DEP (0.47 ng/mL), DEHP, (0.28 ng/mL), or DBP (0.16 ng/mL) with concentration or morphology MBzP associated with decreased FSH (0.9  decrease [0.84–0.96] p  0.003), MBP nonsignificantly associated with increased inhibin B (p  0.07), MEHP non-significantly associated with decreased testosterone (p  0.10) (urinary metabolites) No consistent associations of DBP or DEP with hormone concentrations MEP associated with decreased LH (mean difference 0.7 IU [0.1–1.2]), no associations for MBP, MBzP, and MEHP with LH, T, E2, FSH, inhibin B, and SHBG (urinary metabolites) No association for E2. MEHP associated with decreased T (p  0.038). No associations with MEHHP, MBzP, MBP, MEP, and MiNP (urinary metabolites) 10-fold increase in serum DEHP and DBP concentrations associated with 20% and 26% increases in prolactin respectively, semen DBP associated with decreased T (r   0.21, p  0.02) and T:E2 ratio (r   0.18, p  0.05), semen DEHP associated with increased E2 (r  0.24, p  0.03) and decreased T:E2 ratio (r   0.19, p  0.001) No associations of MMP, MEP, MBP, MEHP, MEHHP, or MEOHP with hormones in male umbilical cord blood (maternal urinary metabolites) MEP and MBP associated with increased SHBG (r  0.323, p  0.002 and r  0.272, p  0.01), MMP, MEP, and MBP associated with increased LH:FT ratio (r  0.210–0.323, p  0.05), MiNP associated with increased LH (r  0.243, p  0.019), MBP associated with decreased FT (r   0.220, p  0.033) (measured in breast milk) MEHP associated with decreased testosterone ( 3.7% [ 6.8 to  0.5] p  0.04) and E2 ( 6.8% [ 11.2 to  2.4] p  0.05), DEHP metabolites associated with decreased FAI (p  0.05), %MEHP associated with increased T:E2 ratio (14% [4–25] p  0.007) (urinary metabolites) No association for exposure to metabolites of DEHP or DiNP, MBP, or total phthalates with testosterone after age adjustment (urinary metabolites) DEHP exposure associated with FT (r   0.14 [ 0.24 to  0.05] p  0.01) and decreased FAI, MEHP associated with increased SHBG (r  0.14 [0.05–0.24] p  0.01) (urinary metabolites) Metabolites of DEHP associated with decreased FT, decreased FAI, decreased E2, and increased SHBG MBP and MEHP associated with decreased FT (r   0.24, p  0.006 and r   0.24, p  0.005), exposure non-significantly associated with decreased LH and FSH (urinary metabolites)

References Rozati et al. (2002) Toshima et al. (2012) Tranfo et al. (2012)

Wirth et al. (2008)

Zhang et al. (2006) Duty et al. (2005)

Janjua et al. (2007) Jonsson et al. (2005) Jurewicz et al. (2013) Li et al. (2011)

Lin et al. (2011) Main et al. (2006)

Meeker et al. (2009)

Mieritz et al. (2012) Mendiola et al. (2011a)

Mendiola et al. (2012) Pan et al. (2006)

AGD, anogenital distance; AGI, anogenital index; BBP, butylbenzyl phthalate; DBP, dibutyl phthalate; DEHP, diethylhexyl phthalate; DEP, diethyl phthalate; DiBP, di-isobutyl phthalate; DiNP, di-isononyl phthalate; DMP, dimethyl phthalate; DNA, deoxyribonucleic acid; E2, estradiol; FAI, free androgen index; FSH, follicle stimulating hormone; FT, free testosterone; LH, luteinizing hormone; MBP, mono-butyl phthalate; MBzP, mono-benzyl phthalate; MCPP, mono(3-carboxypropyl) phthalate; MEHHP, mono-2-ethyl-hydroxy-hexyl phthalate; MEHP, mono ethyl-hexyl phthalate; %MEHP, percent of DEHP metabolites consisting of MEHP; MEOHP, mono-2-ethyl-oxo-hexyl phthalate; MEP, mono-ethyl phthalate; MiBP, mono-isobutyl phthalate; MiNP, mono-isononyl phthalate; MMP, mono-methyl phthalate; MOP, mono-octyl phthalate; OR, odds ratio; SHBG, sex-hormone binding globulin; T, testosterone

Epidemiological literature Anogenital distance Anogenital distance (AGD) is a hormone-sensitive marker of development in rodent models, and presumably reflects the extent of gestational androgen exposure in humans (SalazarMartinez et al. 2004). It has been used as a non-invasive biomarker of androgen disruption in utero by several groups of investigators. Because AGD increases with body mass (Gallavan et al. 1999), it is common in animal studies to correct for body weight which generates a second surrogate outcome called the anogenital index (AGI). Yet, other investigators

have suggested residual confounding by body weight when the AGI is applied to infants of varying ages, necessitating alternate strategies for use in observational studies (Swan 2008). In a seminal cohort study of 85 US boys, strong and statistically significant (p  0.05) associations were detected between phthalate monoesters and smaller than expected AGI for age (Swan et al. 2005). Nine phthalate monoesters were determined in maternal urine specimens collected during prenatal visits, and age-appropriate AGIs for categorizing the outcome were generated using the sample data (Swan et al. 2005). Comparing the highest to the lowest quartile of the sample distribution, increased odds ratios (OR [95% confi-

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dence interval]) for a short AGI ( 6.1 mm) were detected for: MBP (OR  10.2 [2.5–42.2]), MBzP (OR  3.8 [1.03–13.9]), monoethyl phthalate (MEP, OR  4.7 [1.2–17.4]), and monoisobutyl phthalate (MiBP, OR  9.1 [2.3–35.7]) (Swan et al. 2005). Maternal ethnicity, smoking, and time of day, season, and gestational age at specimen collection were considered as potential confounders. Median urine levels of MBP (13.5 ng/ mL), MBzP (8.3 ng/mL), MEP (128.4 ng/mL), and MiBP (2.5 ng/mL) were similar to or lower than reported for US women by biomonitoring studies between 1999 and 2003 (CDC 2013). No associations were detected for DEHP metabolites, including MEHP (median  3.3 ng/mL), MEHHP (11.4 ng/mL), and MEOHP (11.1 ng/mL). An expanded follow-up report of 106 boys (including n  85 from the original study) revealed that log-transformed maternal urinary concentrations of MEP (β   2.934, p  0.005), MBP (β   3.255, p  0.049), MEHP (β   3.503, p  0.017), MEHHP (β   4.977, p  0.002), and MEOHP (β   5.126, p  0.001) were associated with decreased AGD when adjusted for age and weight percentile (Swan 2008). AGD adjusted for population weight percentile was considered in the follow-up study to address residual confounding by weight unaddressed by AGI. MEHP was also correlated with decreased penile width with correction for age and weight percentile (β   0.782, p  0.005), although MEHHP and MEOHP were not (Swan 2008). There were no significant associations detected for MBzP, MiBP, mono methyl phthalate (MMP), or mono(3-carboxypropyl) phthalate (MCPP) with AGD. A significant (p  0.017) but weak association was reported for higher concentrations of specific gravity-corrected, logtransformed urinary MEHP and decreased AGI (β   0.226 [ 0.410 to  0.042]), in a recent cohort study of 111 Japanese boys (Suzuki et al. 2012). The association was adjusted for maternal smoking and age, gestational week, birth order, and urine phytoestrogen levels (Suzuki et al. 2012). Median levels for seven phthalate monoesters determined in urine collected from pregnant women during the 9th–40th week of gestation were similar to or lower than those reported for the earlier US study, including: MEHP (4.68 ng/mL), MEHHP (9.97 ng/ mL), MEOHP (10.0 ng/mL), MMP (7.13 ng/mL), MEP (10.7 ng/mL), MBP (50.8 ng/mL), and MBzP (4.73 ng/mL) (Suzuki et al. 2012). No associations were detected for analytes other than MEHP, including the secondary DEHP metabolites MEHHP and MEOHP (Suzuki et al. 2012). In contrast to the above studies, no association between phthalate levels in amniotic fluid or urine and AGD was found in 33 male newborns delivered to Taiwanese women referred for amniocentesis (Huang et al. 2009). Levels of five phthalate monoesters were determined in maternal urine specimens collected at 35–42 weeks gestation. Median levels for MBP (79.6 ng/mL) and MEHP (26.3 ng/mL) were substantially higher than reported for the US (Swan et al. 2005) and Japanese studies (Suzuki et al. 2012), whereas levels of MEP (19.1 ng/ mL), MBzP (2.5 ng/mL), and MMP (6.8 ng/mL) were similar or lower than in those studies. Amniotic fluid levels for MBP (81.3 ng/mL) and MEHP (22.1 ng/mL) were similar to urine levels, whereas MEP, MBzP, and MMP were not detected. Unlike the US and Japanese studies, women considered in the Taiwanese study were comprised only of high-risk pregnancies, identified by a positive serum screen or advanced mater-

Reproductive toxicity of phthalates in males 471

nal age (Huang et al. 2009). These women may have differed from other pregnant women by exposure level and susceptibility to toxic insult, and so the study results are likely to have limited generalizability to low-risk pregnancies. Though only four human studies have been reported to date, there is limited consistency in terms of individual phthalate monoesters for which associations with AGD have been detected. For example, the strong associations detected for MEP, MBP, MEHP, MEHHP and MEOHP in the expanded US study (Swan 2008) were not reported in the initial publication (Swan et al. 2005), with the exceptions of MBP and MEP, although the latter study sample included the former. Nor were associations reported for the Taiwanese study (Huang et al. 2009). Although MEHP was also identified as a predictor in the Japanese study, MEP, MBP, MEHHP and MEOHP were not (Suzuki et al. 2012). Whereas the US study reported “typical” levels of exposure for MEP and MEHP, with similar or lower levels in the Japanese study, levels in the Taiwanese study were substantially higher. Yet, the sample in the latter study comprised only high-risk pregnancies, unlike the former. Differences between studies could also be due to different methods of measuring AGD. One study measured to the posterior of the scrotum (Huang et al. 2009), and one measured to the anterior base of the penis (Swan 2008), while the other two measured AGD using both methods, reporting weaker associations with phthalate exposure after measuring to the posterior of the genitalia (Suzuki et al. 2012, Swan et al. 2005). Of note, when AGD was measured to the posterior of the scrotum, an association with decreased length was only found with MEP (β   0.429 [ 0.722 to  0.137]; Swan et al. 2005). Though questions have been raised about the biological relevance of changes in AGD (McEwen and Renner 2006, Weiss 2006), AGD and AGI have been linked with other markers of diminished reproductive health in males (Eisenberg et al. 2011, 2012a, 2012b, Hsieh et al. 2008, Mendiola et al. 2011b, Thankamony et al. 2009). Most of these studies measured AGD to the posterior of the scrotum and reported correlations with semen quality (Eisenberg et al. 2011, 2012b, Mendiola et al. 2011b). Importantly, when AGD was measured both to the base of the scrotum and to the top of the penis, an association with semen quality was only found with the first method (Mendiola et al. 2011b), whereas mean differences between boys with normal genitals and with hypospadias were of greater magnitude when using AGD measured to the base of the scrotum (26.5%; P  0.01) versus (vs.) the bottom of the penis (22.5%; P  0.001; Hsieh et al. 2008). Taken together, these studies indicate that AGD is a surrogate marker of biologically important outcomes; however, study results may vary with the definition of AGD in any given investigation and the variability in methods limits inter-study comparisons. We note that phthalates are rapidly metabolized and excreted from the human body without accumulation (Anderson et al. 2001). Variability in exposure sufficient to undermine the detection of epidemiologic associations may thus occur over short time-intervals, associated with deviations in diet, use of personal care products, and other exposure related behaviors. A study of serial urine specimens collected from 28 pregnant women over a six-week interval demonstrated moderate-tohigh within-woman variability in the levels of nine phthalate

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monoesters, including MEHP, MEHHP, MEOHP, MEP, MBP, MiBP, and MBzP (Adibi et al. 2008). More recently, a study of 113 women providing three specimens across pregnancy also reported substantial variability for urine MEP, MBP, MBzP, and DEHP metabolites measured (Braun et al. 2012). Substantial variation within-person has also been attributed to the time of daily specimen collection (Silva et al. 2004). In the studies reviewed, urine collections were not necessarily timed to the window of human sexual differentiation, during the 7th–15th week of gestation (Sharpe 2006). Use of a single urine phthalate determination likely introduced exposure measurement misclassification into the studies reviewed, with bias likely to have been toward the null hypothesis as we do not anticipate the extent of the error would have been related to AGD. Despite the likelihood for exposure measurement misclassification, several strong and statistically significant associations were reported in the US study (Swan 2008), and detected results consistently indicate inverse associations across studies, with preservation of temporality ensured by prospective cohort designs. Furthermore, the reported results are unlikely to reflect exogenous phthalate contamination from environmental or iatrogenic sources, as such spurious exposure would be contributed from the parent diester, rather than the monoester metabolites measured in urine (Hauser and Calafat 2005a). The investigators accounted for confounding by covariates relevant to the study populations and the statistical analyses were appropriate. Yet, the confidence intervals for reported effects were wide, and numerous independent statistical tests were conducted in each study, raising the possibility for type-1 error inflation, and thus the potential detection of spurious results due to chance. Given the paucity of human studies to date, inconsistent results by congeners, and variation in the assessment of study outcome, the epidemiologic evidence for an association between phthalates and AGD is limited. Therefore, we propose that the data indicate only some evidence of an association between phthalate exposure and AGD or AGI. Cryptorchidism Testicular descent occurs in two phases (Hutson 1985), an abdominal descent phase during the first trimester and an inguinal scrotal descent phase during the third trimester; the latter phase is primarily influenced by testicular androgens (Clarnette et al. 1997). Based on the anti-androgenic activity of phthalate monoesters, associations with undescended or cryptorchid testes have been hypothesized. A greater than two-fold increase in the odds of incomplete testicular descent, adjusted for age and body weight percentile, was reported for a doubling in maternal urine MEHP (OR  2.38, p  0.048) among 12 cases compared to 107 controls in the aforementioned US AGD study (Swan 2008). Somewhat stronger adjusted odd ratios for incomplete testicular descent were reported for doublings in the levels of oxidative DEHP metabolites, including MEHHP (OR  2.67, p  0.054) and MEOHP (OR  2.55, p  0.064), albeit of “borderline” statistical significance. No effects were indicated for MEP, MBP, MiBP, MMP, MBzP, or MCPP. In a subsequent study, a higher prevalence of cryptorchidism was found in French children who had mothers with occupational phthalate exposure than among those who did not (OR  8.3, p  0.038); however, the

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total number of children whose parents were occupationally exposed to phthalates was small (seven cases and two controls; Wagner-Mahler et al. 2011). Case and control parents were interviewed with respect to workplace exposure to phthalates before and during the study pregnancy, but no exposure biomarker was employed. Two controls were matched to each of 95 cases by place and date of birth, gestational age, birth weight, and country of parental birth. The statistical analysis was adjusted for age and body weight, yet the authors did not stratify on the remaining matching factors. Unlike the cohort design, matching does not eliminate confounding under the case-control study design but rather enhances precision when an appropriate stratified analysis is used. Without stratification on the matching variables a selection bias is likely, although the direction is difficult to predict (Bloom et al. 2007). No significant difference in phthalate concentrations in breast milk was found in an earlier study of 62 cases of cryptorchidism vs. 68 controls (Main et al. 2006). Cases and controls were recruited from Denmark (n  29, n  36) and Finland (n  33, n  32), with Finnish controls matched to cases by maternal parity, smoking, diabetes, gestational age, and date of birth. The cases and control groups were combined for the analysis, yet the authors did not stratify by matching factors. In this study, phthalate exposure was quantified by measurement of DMP (MMP, Denmark median  0.10 ng/mL and Finland median  0.09 ng/mL), DEP (MEP, 0.93 and 0.97 ng/mL), DBP (MBP, 4.3 and 12 ng/mL), BBP (MBzP, 0.9 and 1.3 ng/mL), DEHP (MEHP, 9.5 and 13 ng/mL), and DiNP (MiNP, 101 and 89 ng/mL) metabolites in breast milk one month after birth and collected over several weeks. However, the use of a bioactive matrix (breast milk) to quantify phthalate exposure cannot exclude the possibility of contamination with parent phthalate diesters that were subsequently metabolized to the measured metabolites (Hogberg et al. 2008). Unlike the primary (hydrolytic) metabolites, secondary (oxidative) metabolites are not vulnerable to endogenous production from the parent compound during specimen collection and storage (Koch et al. 2006), yet these were not measured. Furthermore, exposure was quantified after birth. Although a single sample has been reported to be predictive of phthalate exposure in women of reproductive age over the month (Peck et al. 2010), the aforementioned study demonstrated substantial variation during pregnancy (Adibi et al. 2008). It is uncertain that breast milk samples can provide a realistic measure of exposure during critical periods of in utero development as women’s personal and eating habits change over the course of pregnancy and around delivery, as do pharmacokinetics (Anderson 2005). Therefore, measurement of exposure after birth may not be a valid way of estimating prenatal exposure. A recent case-control study nested within two French cohorts (EDEN and PELAGIE) reported no association between maternal occupational phthalate exposure and undescended testis (OR  0.3 [0.0–3.1]; Chevrier et al. 2012). The EDEN cohort comprised 2,002 women and their infants, recruited preterm in 2004–2006 from obstetrics and gynecology departments in Nancy and Poitiers, France (Drouillet et al. 2009). The PELAGIE cohort comprised 3,421 preterm mothers and their infants recruited in 2002–2006 from areas of Brittany in France (Garlantézec et al. 2009). Cases (n  50) were identified shortly after birth by a midwife or pediatrician. Three controls were

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DOI 10.3109/10408444.2013.875983

matched to each case by area of residence, gestational age at urine collection, and date of specimen collection. A stratified statistical analysis was used, and further adjusted for maternal age, parity, educational level, gestational duration, and urine creatinine, although based on only one exposed case and six exposed controls. In addition, morning maternal urine specimens were collected during the 6th–30th gestational week and levels of MEP (control group median  119 ng/mL), MBP (74 ng/mL), MiBP (84 ng/mL), MCPP (3.6 ng/mL), MBzP (40 ng/ mL), MEHP (13 ng/mL), MEOHP (46 ng/mL), MEHHP (64 ng/mL), mono(2-ethyl-5-carboxypentyl) phthalate (MECPP, 76 ng/mL), monocarboxy-isooctyl phthalate (MCiOP, 3.5 ng/ mL), and monocarboxy-isononyl phthalate (MCiNP, 3.9 ng/ mL) were determined. No differences were detected for individual urine phthalate mono-esters, or for sums of low molecular weight phthalates, high molecular weight phthalates, or DEHP phthalates operationalized as tertiles of the sample distribution. Yet a trend (p  0.06) toward a protective effect was reported for MEP (OR  0.50 [0.2–1.3] and OR  0.38 [0.1–1.1] for the 2nd and 3rd tertiles compared to the 1st tertile, respectively). On balance, only two studies have shown a potential link between cryptorchidism and phthalate exposure (Swan 2008, Wagner-Mahler et al. 2011). We note that while an appropriate matrix was used to directly assess exposure in the US study, the association between MEHP and cryptorchidism was of marginal statistical significance, and many independent statistical tests had been conducted (i.e., the result may reflect a chance occurrence). In the positive French study, personal interview may have led to a recall-related bias, in which case mothers had greater motivation for recall of workplace exposures than controls, consequently overestimating the effect (Werler et al. 1989). Furthermore, the frequently nonspecific nature of the workplace, in which multiple exposures may occur concurrently, is vulnerable to confounding by related compounds. We propose that to establish a link between phthalate exposure and cryptorchidism, exposure should be measured by quantification of phthalate metabolites in maternal urine during the third trimester of pregnancy, a period when inguinal scrotal testicular descent commences under the influence of testicular androgens. A second measure of phthalate metabolites in maternal urine during breast feeding should also be included as a measure linked in time with assessment of testicular descent. Consequently, we suggest that there is inadequate data to establish a link between phthalate exposure and cryptorchidism. Hypospadias Hypospadias results from incomplete urethral closure during development of the external genitalia, between the 8th and 12th gestational weeks (Hughes 2001). As a testosterone-mediated process occurring within a critical developmental window, exposure to anti-androgenic phthalates has been hypothesized to increase the risk. Four reports were found that explored a potential association between parental occupational exposure and hypospadias, three studies using a case-control design and one proportional-morbidity study. An increased risk for hypospadias was found with maternal occupational exposure quantified through job history and a

Reproductive toxicity of phthalates in males 473

job exposure matrix (OR  3.12 [1.04–11.46]), adjusted for income, birth weight, maternal smoking, and folate use during pregnancy (Ormond et al. 2009). The study sample comprised 417 cases referred for surgical correction and 490 randomly selected controls, all residing in southeastern England. However, only 14 cases and 4 controls were delivered to mothers with occupational exposure to phthalates. In addition, capture of isolated defects referred for surgical correction excluded mild and syndromic cases. In contrast, results of a prior birth defects registry-based study revealed no association between exposure to phthalates and the proportion of hypospadias among all congenital anomalies, adjusted for social class, year of birth, region, and maternal age (OR  0.90 [0.74–1.10]; Vrijheid et al. 2003). This proportional-morbidity analysis included all recorded cases of hypospadias in England and Wales from 1980 to 1996 as the numerator (n  3471) and all defects captured by the National Congenital Anomalies System (n  35,962) in the denominator. In a sensitivity analysis, all births reported by the Office for National Statistics (n  181,964) were included in the denominator with little change in the result. Maternal exposure was estimated through occupation as coded in malformations or birth registries and use of a validated job-exposure matrix developed for the study. Similarly, no association was found between maternal occupational exposure to phthalates and hypospadias (OR  1.19 [0.82–1.74]) or paternal exposure to phthalates and hypospadias (OR  1.16 [0.93–1.46]) in a large registry-based case-control study conducted in Western Australia (Nassar et al. 2010). Parental phthalate exposure was assessed using job-exposure matrix and employing birth certificate data, and cases were mild, moderate, or severe hypospadias diagnosed from 1980 to 2000 (n  1075). Controls were recruited from among all male infants delivered in the region without a diagnosis of hypospadias, matched to cases by birth year, in a 2:1 ratio (n  2289). The authors adjusted for maternal age, parity, race, residential location, marital and socioeconomic status, plurality of birth, fetal growth, and birth year. Results were similar for maternal and paternal exposure when stratified by severity of defect or isolated/multiple defect. More recently, the aforementioned nested case-control study of French boys reported no increase in risk associated with maternal workplace phthalate exposure (OR  4.2 [0.4–41.3]), although the results were based on only two exposed cases and two exposed controls (Chevrier et al. 2012). No associations were indicated for the 11 phthalate monoesters measured in maternal urine specimens, or for sums of low molecular weight phthalates, high molecular weight phthalates, or DEHP metabolites. In fact, a trend (p  0.06) toward a protective effect was reported for the sum of low molecular weight phthalates (OR  0.16 [0.03–0.9] and OR  0.20 [0.02–2.5] for the 2nd and 3rd tertiles compared to the 1st tertile, respectively). The only study to show a positive association between phthalate exposure and hypospadias employed too few exposed cases and controls (14 and 4, respectively) to draw meaningful conclusions, and was likely to have recruited only the most severe cases (Ormond et al. 2009). While the other studies were statistically more robust with overall sample sizes of 1202–3471 cases and 2583–32,491 controls (Nassar et al. 2010, Vrijheid et al. 2003), phthalate exposure in three of four

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studies was assessed through occupational history (using a job-exposure matrix). Exposure assessment using job-exposure matrix is expected to misclassify some participants due to uncertainty, yet bias is likely toward the null hypothesis (i.e., similar for cases and controls). On the other hand, use of a job-exposure matrix avoids issues related to recall that threaten data collected by parental interviews. An additional recent, but small study, reported no association using a direct measure of exposure, although half of the mothers provided urine after the 26th week of gestation, past the critical window for urethral development (Chevrier et al. 2012, Philippat et al. 2012). Thus, in the absence of direct measures of exposure at the critical window and the three null studies, we conclude that there is inadequate evidence to date in support of an association between phthalate exposure and hypospadias. However, we note that, in the absence of a well-designed robust epidemiological study with direct measures of phthalate exposure timed to the development of the external male genitalia, it is impossible to exclude phthalate exposure as a potential risk factor in the development of hypospadias. Puberty The age of pubertal onset has been reported to be occurring at younger ages in both girls and boys (Buck Louis et al. 2006). Although the mechanisms triggering puberty onset are poorly understood, gonadal steroids are key players in the process. Therefore, a link between exposure to phthalates and earlier pubertal onset has been explored. No association between DEHP exposure through extracorporeal membrane oxygenation (ECMO) as infants and the onset of puberty was found when compared to population reference values in 13 males (Rais-Bahrami et al. 2004). Blood is subject to substantial DEHP contamination while circulating through PVC tubing during ECMO. However, this study was very small and had several other limitations including: no direct measure of exposure, and discordance between the time of the exposure assessment (infancy) and age at puberty assessment. In addition to small study size, growth and hormonal parameters were within normal limits for all participants (with one exception related to a diagnosis of Marfan’s syndrome), and no comparison was made to a referent group recruited specifically from the case sampling frame. Androgen–estrogen imbalance can elicit ductal proliferation in male breast tissue during adolescence, prompting investigations of the potential relationship between phthalate exposure and pubertal gynecomastia. In a Turkish hospitalbased case-control study (Durmaz et al. 2010), circulating plasma concentrations of DEHP and MEHP were higher in cases (n  39) compared to age-matched controls (n  20), with odds ratios of 2.77 [1.48–5.21] and 24.76 [3.5–172.6] for 1 μg/mL increases of DEHP and MEHP, respectively. Higher DEHP (P  0.04) and MEHP (P  0.018) levels were also noted for nine cases presenting with severe pain compared to those without pain (Durmaz et al. 2010). The mean MEHP (1.37 μg/mL) level in controls was at least several orders of magnitude higher than reported for urine in the aforementioned studies, whereas DEHP was greater still (3.09 μg/mL). No adjustment was made for confounding and the results were not stratified by age, likely introducing a selection bias given

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the use of a matching strategy under the case-control study design as previously described (see Cryptorchidism). Unfortunately, phthalate exposure was measured in blood and thus potential contamination cannot be excluded. Controls were recruited as “healthy” children with no endocrinologic disorder and might thus be less likely to have experienced greater overall exposure to phthalates at the time of the assessment compared to the cases who were admitted to the hospital for treatment and might have experienced iatrogenic exposures. Additionally, no dose–response relationship with the amount of breast enlargement was found, nor was there any association between testis volume or stretched penis length and exposure. In a subsequent cross-sectional study, no age-adjusted association was found between contemporary urinary phthalate metabolite concentrations and the onset of puberty in 555 Danish boys assessed using Tanner pubic hair and genital stages (Mieritz et al. 2012). Analyses were adjusted for age and urine specimens, uncorrected for dilution, were collected on the morning of physical examination. Levels of 12 phthalate monoesters were determined, often at levels higher than those reported for 12- to 17-year olds residing in the US between 2007 and 2008, respectively, including: MEP (median  36.24 vs. 97.4 ng/mL), MiBP (74.88 vs. 10.5 ng/mL), MBP (45.14 vs. 27.3 ng/mL), MBzP (47.70 vs. 7.13 ng/mL), MEHP (5.05 vs. 2.30 ng/mL), MEHHP (47.36 vs. 19.6 ng/mL), MEOHP (25.06 vs. 10.7 ng/mL), MECPP (27.39 vs. 38.9 ng/mL), MiNP (0.65 ng/mL), mono(hydroxy-iso-nonyl) phthalate (MHiNP, 5.6 ng/mL), mono(oxo-iso-nono) phthalate (MOiNP, 3.29 ng/ mL), and MCiOP (7.66 ng/mL). Similarly, no difference in phthalate exposure was found in a case-control study of 38 boys with gynecomastia (breast enlargement) vs. 189 agematched controls, nested within the Danish cross-sectional study (Mieritz et al. 2012). However, that analysis was not stratified by age, despite the use of a matching strategy under the case-control study design possibly introducing a selection bias into the study results as previously described (see Cryptorchidism). Both of the recent phthalates-puberty studies used casecontrol designs with exposures assessed in prevalent (i.e., existing) cases, and so could only reflect the pubertal status of boys at one time point. Thus, these studies cannot exclude potential changes in the pace of pubertal development in relation to phthalates exposure over time. The latter point is important in view of a recent assessment of the secular change in pubertal onset which concluded that although pubertal onset does not appear to be occurring at an earlier age, the pace of pubertal transition has increased (Walvoord 2010). Contamination concerns undermine the positive results reported by the Turkish study of gynecomastia (Durmaz et al. 2010), and the large Danish study of exposure at phthalate levels higher than reported for biomonitoring studies detected no association (Mieritz et al. 2012). Hence, we conclude that the current literature is underwhelming and remains inadequate to determine an effect of phthalate exposure on puberty. Moving forward we suggest that future studies not only include measures of phthalate metabolites in urine throughout the pubertal transition but that longitudinal studies be considered that allow for assessment of the speed of the pubertal transition through the Tanner stages.

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Semen quality A decline in global semen quality over a 50-year span has been reported (Carlsen et al. 1992) leading to a plethora of papers and significant controversy. The impact of phthalate exposure on male fertility has also emerged as a concern from these observations. Of all of the health effects of phthalates, the impact of phthalate exposure on semen quality has received the greatest attention. Semen quality is not a single outcome measure but is an overall term that is comprised of sperm count, motility, and morphology. Investigators often dichotomize outcomes according to guidelines adopted by the World Health Organization, most recently including sperm concentration 15,000,000/mL or more, motility 40% or higher, and normal morphology 4% or higher (WHO 2010). In addition, several studies have examined the relationship between phthalate exposure and DNA damage of sperm and thus will be discussed. Sperm concentration, count, and morphology in general populations

Few studies of phthalate exposure and male reproductive health have been conducted among the general population. An early study (Murature et al. 1987) described sperm concentration and DBP measured in the seminal fluid of 21 US students recruited through a newspaper. Using a subgroup analysis, presumably defined by rate of DBP metabolism (i.e., n  12 low metabolizers vs. n  9 high metabolizers), DBP was associated with decreased sperm concentration. However, an overall analysis indicated a positive association between DBP and concentration. In a subsequent study (Jonsson et al. 2005), decreased sperm motility (8.8% fewer motile sperm [0.8–17]) was associated with increased urinary MEP concentrations (median  83 ng/mL) in a cross-sectional study of 234 Swedish military conscripts aged 18–21 years. Smoking and abstinence time were considered as confounding variables. No semen quality associations were noted for MBP (median  24 ng/mL), MBzP (4.4 ng/mL), or MEHP ( 15 ng/mL). A positive association between sperm motility and phthalic acid (9.4% more motile sperm [3.7–15]) suggested that an increased ability to metabolize phthalates may be protective (Jonsson et al. 2005); phthalic acid is a biomarker of phthalate diester metabolism. This study is highly relevant as it is one of the few reports in the literature to use healthy men as opposed to a highly selected population of men such as those seeking a vasectomy or attending an assisted reproductive therapy clinic. However, the participation rate was low (14%), raising the possibility for a selection bias, although the authors note this as unlikely. In addition, coefficients of variation for duplicate analysis on different days were high for some phthalates, indicating low analytic precision and raising the possibility for exposure measurement misclassification and undetected associations. A more recent publication suggested urine MBP above the sample median (18.72 ng/mL) was associated with decreased sperm concentration (OR  1.97 [0.96–4.08], p  0.10), adjusted for age and abstinence time in a general population sample of 232 men residing in a heavily industrialized urban area in China (Han et al. 2013). No associations were suggested for MEP (median  3.10 ng/mL) or MEHP (1.10 ng/mL), or for motility and morphology.

Reproductive toxicity of phthalates in males 475 Sperm concentration, count, and morphology in clinical populations

The majority of studies of phthalate exposure and semen quality have been conducted among infertile populations, frequently the male partners of couples undergoing infertility treatment. In 2006, higher urine MBP (median  17.7 ng/mL) was reported to be associated with decreased concentration and motility in 443 men undergoing infertility evaluation at the Massachusetts General Hospital, adjusted for age, abstinence time, and smoking status (OR  3.3 [1.2–8.5], OR  1.8 [1.1–3.2], respectively, p  0.04 for both; Hauser et al. 2006). No associations with semen quality were suggested for MEP (median  158 ng/mL), MMP (3.8 ng/mL), MEHP (7.9 ng/mL), MEOHP (32.1 ng/mL), or MEHHP (48.1 ng/mL). Urine specimens were spot collected and phthalate concentrations were corrected for specific gravity to accommodate dilution. The study participation rate was high (60%) and selection bias was unlikely. However, odds ratios overestimate underlying population relative risks when outcomes are common ( 10%; Zhang and Yu 1998), potentially biasing effect estimates away from the null hypothesis in studies conducted among subjects recruited from infertility clinics. Still, the results confirmed those for the top tertile of urine MBP and decreased sperm motility, adjusted for age, abstinence time, and smoking (OR  3.0 [1.2–7.6]) in an earlier investigation (n  163) conducted in the same study population (Duty et al. 2003a). The adjusted association for decreased sperm concentration and MBP did not achieve significance in the earlier report (OR  3.3 [0.9–12.6]), although an association for MBzP did (OR  5.5 [1.3–23.9]). Nonsignificant associations between urinary MMP and increased abnormal morphology (OR  3.2 [0.8–12.2]) as well as urinary MBzP and decreased sperm concentration (OR  2.7 [0.8–8.5]) were also described. A further report in 2004 using computer-aided sperm analysis (CASA), found that urinary MEHP, MBP, and MBzP were nonsignificantly associated with decreased motility between the third and first tertiles (for curvilinear velocity, β   2.93 [ 8.52 to 2.66], p  0.3, β   3.46 [ 9.09 to 2.18], p  0.2, β   2.45 [ 8.02 to 3.11], p  0.4), higher urinary MEP was associated with improved curvilinear velocity (β  6.36 [0.74– 11.99], p  0.03), and MMP had no association with semen quality (for curvilinear velocity β   1.27 [ 6.97 to 4.42], p  0.7) adjusted for age, smoking and abstinence time (Duty et al. 2004). Additional investigation in this study population suggested that associations between MBP, MBzP, and sperm motility were potentiated by concurrent exposure to polychlorinated biphenyls (PCBs; Hauser et al. 2005b), ubiquitous and persistent pollutants were also considered for association with semen quality (Phillips and Tanphaichitr 2008). The association between MBP and decreased sperm motility appears consistent, as demonstrated in a second series of studies conducted among couples undergoing infertility evaluation in India. Phthalate diesters were measured in the semen of 300 men, and correlations were detected between DBP, DEHP, and decreased motility (r   0.18 for both, p  0.05) in unadjusted analyses (Pant et al. 2008). Unadjusted associations were also reported between DBP, DEHP, DEP, and decreased sperm concentration (r   0.20, r   0.25, r   0.19, p  0.05), and for DEHP and increased abnormal morphology (r  0.18, p  0.05). No associations were

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reported for semen DMP or di-n-octyl phthalate (DnOP). In a subsequent cross-sectional study conducted at the same site, the median semen DBP concentration (1.23 μg/mL) was greater in 65 oligoasthenospermic men vs. 50 fertile men (0.07 μg/mL), and median DEHP concentration (0.15 μg/mL) was also greater in oligoasthenospermic men vs. in fertile men (0.05 μg/ml), although differences were not tested (Pant et al. 2011). In addition, sperm motility was inversely associated with DBP (r   0.25, p  0.01 and r   0.20, p  0.01) and DEHP (r   0.30, p  0.001 and r   0.25, p  0.01) in oligoasthenospermic (n  65) and asthenospermic (n  65) men, respectively. However, the associations were not adjusted for potential confounding variables. The assessment of parent phthalate diesters at μg/mL concentrations raises serious concerns with respect to exogenous contamination during specimen storage and analysis, an issue that was not addressed. Another cross-sectional study recruited 97 men undergoing infertility treatment in China, and found the top tertile of urinary MBP (median  10.1 ng/mL) to be strongly associated with decreased sperm concentration (OR  12.0 [1.01–143], p  0.05), adjusted for age, abstinence time, body mass index (BMI), smoking, alcohol consumption, and education (Liu et al. 2012). Phthalate concentrations were corrected for urine creatinine to accommodate dilution. There were no associations reported for MMP (median  15.4 ng/mL), MEP (12.6 ng/mL), MBzP ( 0.15 ng/mL), MEHP (0.50 ng/mL), or MEOHP (median  1.82 ng/mL). Two urine specimens were collected from each participant, several days apart and presumably averaged to reduce intra-individual variability. However, the results were based on only 11 cases and were thus imprecise. As noted earlier, reduced sperm concentration may not be infrequent among members of infertile populations and thus the odds ratio is likely to overestimate the true population relative risk. Urinary MEP (median  108 ng/mL), collected as first morning voids, was associated with decreased concentration in 45 men being investigated at a Michigan infertility clinic (OR  6.5 [1.0–43.6]), adjusted for race and urine specific gravity, and with abnormal morphology (OR  3.4 [0.9–13.8]), adjusted for specific gravity (Wirth et al. 2008). Urinary mono-3-carboxypropyl phthalate (MCPP) was also associated with an increased proportion of morphologically abnormal sperm (OR  7.6 [1.7–33.3]), and an association was suggested for urinary MEHP (median  10.1 ng/mL) with low sperm concentration (OR  5.4 [0.9–30.8]). However, no associations were indicated for MMP (median  2.4 ng/mL), MBP (24.7 ng/mL), MiBP (5.8 ng/mL), MBzP (17.4 ng/mL), MEHP (10.1 ng/mL), MEOHP (37.6 ng/mL), and MEHHP (56.6 ng/mL). In contrast to previous studies, sperm concentration was higher in men (p  0.05 and p  0.01, respectively) with greater urinary concentrations of MBP (median  7.22 ng/ mL) and MBzP (9.18 ng/mL) in a cross-sectional study of 41 men of subfertile couples in Tokyo (Toshima et al 2012). In a multiple regression analysis, specific gravity-corrected urinary MBP was positively and significantly associated with increased sperm concentration (β  0.294, p  0.05), adjusted for equol detectability, age, BMI, abstinence period, smoking, and frequency of consuming fruits, vegetables, and coffee. Morning urine specimens were collected from participants at their homes, just prior to semen sample collection. No associations

Crit Rev Toxicol, 2014; 44(6): 467–498

were reported for urinary MMP (median  7.22 ng/mL), MEP (10.7 ng/mL), MEHP (5.94 ng/mL), MEHHP (11.5 ng/mL), and MEOHP (7.93 ng/mL), or for phthalates and motility. Another study of 349 men recruited from a German andrology clinic reported no associations between urinary DEHP metabolites collected as spot samples (median MEHP  4.35 ng/mL; MEHHP  12.66 ng/mL; MEOHP  9.02; 5-carboxymono(2-ethylhexyl)phthalate (5cx-MEPP)  14.53 ng/mL), and any parameter used to measure semen quality, adjusted for age, smoking, abstinence period, and urine creatinine (Herr et al. 2009). Furthermore, total semen phthalates were higher among 21 men with an infertility diagnosis compared to 32 men without (p  0.05), in a cross-sectional study of infertility clinic patients in India (Rozati et al. 2002). Among the infertile men, seminal phthalates were associated with decreased morphologic normality (r   0.769, p  0.001), but not with count or motility. Finally, no associations were detected for sperm count or morphology and semen concentrations of DEP (mean  0.47 ng/mL), DEHP (0.28 ng/mL), and DBP (0.16 ng/mL) in a second study of 52 men attending a reproduction clinic in Shanghai (Zhang et al. 2006). Most recently, a large study (n  269) of men under 45 years of age and attending a Polish infertility clinic measured MEHP (median  8.8 ng/mL), MEHHP (18.4 ng/mL), MiNP (1.1 ng/mL), MBzP (5.2 ng/mL), MBP (83.4 ng/mL), and MEP (45.2 ng/mL) in morning voids (Jurewicz et al. 2013). Although attending an infertility clinic, men recruited to the study had semen parameters within or close to WHO 1999 reference limits. Urine phthalate levels were higher than reported for semen quality studies conducted elsewhere in Europe or in Asia, yet lower than those reported for the earlier described Massachusetts General Hospital Study. Inverse associations were detected for percentage sperm motility with log-transformed MEHP (β   3.85%, p  0.001), MEHHP (β   3.94, p  0.003), and MiNP (β   9.05%, p  0.003), and for straight line (β   4.11%, p  0.007), and curvilinear (β   6.56%, p  0.009) velocities with MBP, adjusted for age, smoking, abstinence period, past diseases and creatinine as confounding covariates. No adjusted associations were detected for sperm concentration, morphology, and linearity, or for urine MBzP and MEP. Sperm DNA damage

Damage to the male genetic complement contained within sperm may lead to delayed fertilization, compromised implantation, and infertility (Sakkas and Alvarez 2010). In the aforementioned Massachusetts General Hospital study, urinary MEP was associated with increased comet extent and tail-distributed moment in 379 men (β  6.6 μm per interquartile range (IQR) increase [0.94–12.3] and β  2.72 μm/ IQR [ 0.46 to 5.00], p  0.02), adjusted for age and smoking (Hauser et al. 2007). Increased values for neutral comet assay parameters, a single cell gel electrophoresis procedure, indicate reduced DNA integrity (Singh et al. 1988). The results confirmed those for an earlier investigation of 141 men from the same study population, in which an adjusted association was also described between urinary MEP and DNA damage (for comet extent β  3.5 μm/IQR [0.74–6.47]); (Duty et al. 2003b). Sperm DNA damage was re-examined in the larger

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sample and MBP was associated with an increase in the percentage of DNA in the comet tail (β  1.63 [0.20–3.08]) and MBzP was associated with increased comet extent and tail-distributed moment (β  5.12 μm/IQR [0.98–9.25] and β  2.49 μm/IQR [0.82–4.13], respectively; Hauser et al. 2007). MEHP was also associated with an increased percentage of DNA in the comet tail (β  3.06 μm/IQR [1.33–4.79], p  0.0006). The association between percentage MEHP (the percentage of the DEHP metabolites consisting of MEHP on a molar basis) and DNA damage was also examined, with positive associations detected for comet extent (β  15.0 μm/IQR [7.9–22.0]), tail-distributed moment (β  5.54 μm/IQR [2.76–8.32]), and percentage of DNA in the comet tail (β  3.11 μm/IQR [0.65–17.2; Hauser et al. 2007). Hauser (2008) theorized that the ability to metabolize DEHP to MEHP was protective. This theory agrees with Jonsson et al.’s finding that greater concentrations of phthalic acid, perhaps indicative of faster metabolism, were associated with improved semen quality (Jonsson et al. 2005). These results suggest that efficient elimination of phthalate diesters and metabolites has a protective effect. Supporting this observation, the reproductive consequences of phthalate exposure were more severe in rats with induced kidney failure (Nabae et al. 2006, Tsutsumi et al. 2004). These studies indicate that renal clearance of phthalates plays an important role in preventing phthalate-induced damage. Therefore, we suggest that some populations may be at a greater risk of adverse effects from phthalate exposure. Hence, it may be important to examine the concentrations of phthalates and any adverse effects in dialysis patients who could be exposed to higher concentrations of phthalates due to medical treatments and who have kidney failure. Additional evidence also indicates compromised sperm DNA quality in association with phthalates exposure. The aforementioned study of 53 men attending an Indian infertility clinic reported an increased percentage of single-strand DNA (r  0.855, p  0.001) in association with increased semen phthalates (Rozati et al. 2002). An occupational study of 45 PVC plant employees receiving sperm chromatin structure assay (SCSA) reported a positive association between sperm DNA damage (the tendency for DNA denaturation, β  0.038, p  0.015 and DNA fragmentation index, β  0.014, p  0.010), and negatively correlated with motility (β   0.227, p  0.044) with the concentration of DEHP measured in the personal breathing zone at each employee’s work station (Huang et al. 2011). Associations were adjusted for age, smoking, and coffee drinking. No associations were detected for sperm concentration or morphologic abnormality. Unfortunately biomarkers of exposure were not included. The aforementioned Polish study reported a positive association for covariate-adjusted log urine MBP and log DNA fragmentation index (β  0.20, p  0.047), although no association for the other phthalates was reported (Jurewicz et al. 2013). However, fluorescent in situ hybridization analysis indicated significantly increased sperm aneuploidy rates, in association with covariate-adjusted MBzP, MBP, MEHP, and MEP. Overall, the studies examining phthalate metabolites in urine and parameters of semen quality yielded inconsistent results, although there are some common themes to the data. In general, MBP and MEHP were often associated with decreased semen

Reproductive toxicity of phthalates in males 477

quality to some extent and MBzP was occasionally correlated with decreased semen quality. Some differences between the studies could be due to a difference in the population sampled, for example, healthy compared to subfertile men or men in different areas of the world with different exposures. Differences could also be due to chance findings resulting from small sample sizes and multiple statistical tests. Studies assessing exposure to phthalate diesters in semen may reflect exogenous contamination, and assessment of monoesters may be biased by in situ esterase activity post collection. Most research has been conducted among members of infertile populations; a group perhaps more sensitive to toxic insult than the general population. In addition, participation rates were generally low and men volunteering for semen quality studies may differ from those who do not, a potential source of bias, although the impact is difficult to predict (Cohn et al. 2002, Eustache et al. 2004). Fertility and fecundity Associations between fertility, defined as the production of live offspring (e.g., a live birth), and fecundity, defined as the ability to produce live offspring (e.g., conception, clinical pregnancy), offer indicators of overall reproductive fitness (Porta 2008). A recent study collected urine for analysis of five phthalate monoesters in 56 infertile couples attending an Italian reproductive health clinic, and 56 control couples with demonstrated fertility residing in the same area (Tranfo et al. 2012). Median urinary concentrations of phthalate monoesters were higher among cases than among controls, respectively, including MEP (198.62 vs. 51.76 μg/g creatinine, p  0.001), MEHP  MEHHP (17.2 vs. 15.14 μg/g creatinine, p  0.009), MBzP (12.37 vs. 8.8 μg/g creatinine, p  0.009), and MBP (53.76 vs. 31.36 μg/g creatinine, p  0.001). Stratified by sex, significant differences remained for MEP and MBP among men, with a similar trend for MEHHP (p  0.076) and MBzP (p  0.053). However, couples undergoing infertility treatment comprise a highly selected group of individuals unable to have a child and who seek medical attention; in particular if in vitro fertilization (IVF) is involved. Inherent socioeconomic and behavioral differences related to phthalates, between those seeking and not seeking infertility treatment might have biased study results and limit generalizability to non-IVF couples (Wilcox 2010). In an earlier retrospective cohort study of occupational exposures and fecundity, there were no associations between increased time-to-pregnancy, and paternal occupational DEHP exposure (based on employee’s job and known exposure levels within the factory) among 153 couples adjusted for maternal and paternal age at time of pregnancy, and for length of recall (Modigh et al. 2002). Data related to study pregnancies were collected up to a decade after the event, through telephone interview or mailed questionnaire. Contrary to earlier evidence (Joffe et al. 1995), recent work suggests this approach may misclassify study outcomes (Cooney et al. 2009). Female employees who were also exposed to phthalates were more likely to be included in the reference group than that in the exposed groups, potentially introducing a bias toward the null hypothesis. In addition, women in the workforce tend to have a higher probability for infertility

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than women not employed outside of the home (Joffe 1985), also a potential source of bias toward the null hypothesis. However, a sensitivity analysis excluding reference group pregnancies occurring during maternal employment resulted in little change in the results, suggesting minimal impact from these sources (Modigh et al. 2002). Yet, exclusion of couples attempting, yet not achieving, pregnancy might have biased study results toward the null hypothesis, if sensitive groups or those with higher exposures were less likely to conceive a clinical pregnancy during follow-up. Given the methodological limitations to the few epidemiologic studies conducted to date, we conclude that there is sparse human evidence for associations between male phthalate exposure, fertility, and fecundity. Reproductive hormones The results of 12 epidemiologic studies evaluating serum reproductive hormones in association with exposure to phthalates are summarized in Table 2. Sex-steroid hormones were assessed in association with urinary MEP, MBP, MBzP, MEHP, MEHHP, and MEOHP levels among 425 men participating in the earlier described Massachusetts’s General Hospital Study (Meeker et al. 2009). The molar sum of DEHP metabolites, and percentage MEHP were assessed as well. An IQR increase in MEHP was associated with decreased serum estradiol (β   2.04 pg/mL [ 3.37 to  0.73]) and testosterone (β   14.9 ng/dL [ 27.5 to  2.30]), adjusted for age, BMI, smoking, and time of specimen collection. A significant association between percentage MEHP and an increase in the testosterone:estradiol ratio (β  1.14 [1.04–1.25], p  0.007) was also reported. Adjusted decreases in the free androgen index (FAI), which describes the ratio of testosterone to sex-hormone binding globulin (SHGB), were also reported for MEHHP, MEOHP, and the sum of DEHP metabolites (p  0.05). No associations were reported for MBP, MBzP, serum prolactin, FSH, or LH. An earlier publication concerning 295 men from the same study population reported an association between an IQR increase in urine MBzP and decreased FSH (multiplicative factor  0.90 [0.84–0.96], p  0.003) and a nonsignificant association between an IQR increase in MEHP and decreased testosterone (β   0.47 [ 1.03 to 0.10], p  0.1), adjusted for age, BMI, and time of specimen collection (Duty et al. 2005). Furthermore, MBP was nonsignificantly associated with increased inhibin B after adjustment for covariates (β  7.33 pg/mL [ 0.55 to 15.21], p  0.07). This study contradicts the result from a study of 118 men seeking infertility treatment in Shanghai, China (Li et al. 2011). Li et al. reported no significant association between DEHP concentrations in serum (median  0.05 μg/mL) or semen (median  0.02 μg/mL) and estradiol or testosterone, after adjustment for age, BMI, educational background, smoking, and alcohol consumption. Although, unadjusted correlations were detected for DEHP and increased estradiol (r  0.24, p  0.03) and semen DBP (0.02 μg/mL) with decreased testosterone (r   0.21, p  0.02). On the other hand, associations between semen DEHP exposure and the testosterone:estradiol ratio (β   0.08 [ 0.15 to 0.00], p  0.05) as well as positive associations between prolactin and serum DEHP, semen DEHP, and serum DBP (0.05 μg/mL) were detected after cor-

Crit Rev Toxicol, 2014; 44(6): 467–498

rection for potential confounders (Li et al. 2011). However, the former was only identified following exclusion of 34 subjects with DEHP below the limit of detection, an approach likely to introduce bias (Schisterman et al. 2006). Yet, a dose-dependent increase in prolactin was also associated with DEHP and DBP, corresponding to differences of 20–26% for a ten-fold increase in exposure (Li et al. 2011). Nonetheless, the use of serum and semen to quantify phthalate exposure makes interpretation of these data difficult, and decreases the validity of the results due to the possibility of contamination. Other studies did not examine prolactin but reported similar decreases in testosterone. One cross-sectional study in China quantified exposure using urinary metabolites in 74 occupationally exposed workers and 63 unexposed referent workers, matched by age and smoking (Pan et al. 2006). Decreases in free testosterone were associated with MEHP (median exposed vs. unexposed  562.3 vs. 5.4 μg/g creatinine) and MBP (548.4 vs. 113.5 μg/g creatinine) exposure in all the men (r   0.242, p  0.005 and r   0.237, p  0.006), adjusted for age and alcohol consumption (Pan et al. 2006). In addition, concentrations of follicle stimulating hormone (FSH) were nonsignificantly decreased (r   0.103, p  0.107 and 0.135 for MBP and MEHP, respectively; Pan et al. 2006). In a cross-sectional study of 363 fertile men participating in a multicenter US study, inverse associations were described for urine DEHP metabolites (median MEHP, MEHHP, MEOHP, and MECPP  3.2, 23.7, 12.9, and 32.3 ng/mL, respectively) and FAI. The associations included MEHP (β   0.05 [ 0.08 to  0.02], p  0.01), MEHHP (β   0.04 [ 0.07 to  0.004], p  0.05), and MEOHP (β   0.04 [ 0.07 to  0.01], p  0.05) and were adjusted for age, BMI, smoking, ethnicity, urine creatinine, and time of sample collection (Mendiola et al. 2011a). A positive association was also detected for MEHP and SHBG, and the FAI:luteinizing hormone (LH) ratio after adjustment for covariates (β  0.05 [0.02–0.09], p  0.01). However, no associations were reported for FSH, estradiol, or LH. The earlier discussed study of Swedish conscripts reported a cross-sectional association between urine MEP and decreased LH (mean difference  0.7 IU [0.1–1.2]) in 234 healthy men when comparing the quartile of highest exposure to the lowest (Jonsson et al. 2005). However, no associations between concentrations of the other phthalate metabolites and concentrations of the other hormones, including testosterone, estradiol, FSH, inhibin B, and SHBG were found. A recent study compiled data from two investigations of phthalate exposure and reproductive hormones; combining men from the Massachusetts General Hospital and multicenter US studies described above (Mendiola et al. 2012). In a combined total of 783 men there were no associations with urinary concentrations of MEP, MBP, or MBzP and any reproductive hormone measured. However, metabolites of DEHP were associated with decreased free testosterone (MEHP β   0.02 [ 0.04 to  0.004], MEHHP β   0.02 [ 0.04 to  0.001], MEOHP β   0.02 [ 0.04 to  0.001]), decreased FAI (MEHP β   0.02 [ 0.04 to  0.01], MEHHP β   0.04 [ 0.06 to  0.01], MEOHP β   0.04 [ 0.06 to  0.01]), decreased estradiol (MEHP β   7.9 [ 12.4 to  3.5]), and increased SHBG (MEHHP β  0.03 [0.002–0.06], MEOHP β  0.03 [0.005–

↓ ΣDEHP, ΣDiNP a ↑ MEP ↓ MEHP

a



↓ MEHP 

↓ MBP



↓ MEHP

↓ MEHP, %MEHP –











E2



↑ DEHP, DBP

Prolactin

b

b

↑ ↑ MEHP, MEHHP, MEHP, MEHHP, MEOHP MEOHP – –













↓ MBzP –

FSH





↑ MiNP



↓ MEP



LH –













↓ MEHP



↑ %MEHP

b

↓ DEHP –

Inhibin B T:E2 ratio –

↓ MEHP, MEHHP, MEOHP

↑ MEHHP, MEOHP, ΣDEHP ↓ MEHP, MEHHP, MEOHP, MECPP, ΣDEHP



FAI –







↑ MMP, MEP, MBP



LH:T or LH:FT ratio

↓ MEHP

FAI:LH ratio

NOTE: ↑ signifies a significant increase, ↓ signifies a significant decrease, and  signifies no significant change in concentrations of the hormone with an increase in the phthalate metabolite concentration; empty cell indicates that the association was not considered. aNo association when adjusted for age. bNo association when adjusted for confounding variables. DEHP, diethyl hexyl phthalate; ΣDEHP, sum of DEHP metabolites; DiNP, di-isononyl phthalate; ΣDiNP, sum of DiNP metabolites; E2, estradiol; FAI, free androgen index; FSH, follicle-stimulating hormone; FT, free testosterone; LH, luteinizing hormone; MBP, mono-butyl phthalate; MBzP, mono-benzyl phthalate; MECPP, mono(2-ethyl-5-carboxypentyl) phthalate; MEHHP, mono-2-ethyl-hydroxy-hexyl phthalate; MEHP, mono ethyl-hexyl phthalate; %MEHP, percent of DEHP metabolites consisting of MEHP; MEOHP, mono-2-ethyl-oxo-hexyl phthalate; MEP, mono-ethyl phthalate; MMP, mono-methyl phthalate; SHBG, sex hormone-binding globulin; T, testosterone

↑ MEHHP, MEOHP

↑ MEHP



↑ MEP, MBP





SHBG – –



FT





T

b b Mendiola et al. ↓ ↓ (2011) MEOHP, MECPP MEHP, MEHHP, MEOHP, MECPP, ΣDEHP b Mendiola et al. ↓ ↓ (2012) MEHP, MEHHP, MEHP, MEHHP, MEOHP MEOHP Pan et al. ↓ (2006) MEHP, MBP

Meeker, et al. (2009)

Mieritz et al. (2012)

Lin et al. (2011) Main et al. (2006)

References Duty et al. (2005) Janjua et al. (2007) Jonsson et al. (2005) Jurewicz et al. (2013) Li et al. (2011)

Table 2. Summary of the effects of phthalates on serum reproductive hormone concentrations in men and boys as reported in the literature.

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0.06]). Finally, the earlier described study of Polish infertility clinic patients reported reduced testosterone in association with increased covariate-adjusted log urine MEHP (β   0.29, p  0.038), although there was no association for estradiol or other measured phthalates (Jurewicz et al. 2013). The results from studies conducted in children were also inconsistent. An earlier described study of 96 Danish and Finnish infants, with and without cryptorchidism, examined a possible association between hormone concentrations at three months and phthalate concentrations in maternal breast milk (Main et al. 2006). Associations with decreases in free testosterone concentrations with MBP (r   0.220, p  0.033) as well as with increases in SHBG with MBP and MEP (r  0.272, p  0.01 and r  0.323, p  0.002), LH with MiNP (r  0.243, p  0.019), and the LH:free testosterone ratio with MBP, MEP, and MMP (r  0.282, p  0.006, r  0.323, p  0.002, and r  0.210, p  0.044) were found (Main et al. 2006). While breast feeding is likely to be an important exposure source for infants, measurements made in breast milk may be less accurate due to contamination with environmental phthalate diesters (Latini et al. 2004). On the other hand, another study examining in utero exposure and hormones in the cord blood found no associations between age-adjusted MMP (median  34.6 ng/mL), MEP (34.6 ng/mL), MBP (65.5 ng/mL), MBzP (8.85 ng/mL), MEHP (11.7 ng/mL), MEOHP (17.2 ng/mL), and MEHHP (11.8 ng/mL), measured in third trimester maternal urine and sex-steroid hormones in 81 male Taiwanese infants (Lin et al. 2011). The aforementioned puberty study in 555 Danish boys 6–19 years of age, reported no association between phthalate exposure and testosterone after adjustment for age (Mieritz et al. 2012). MBP and MEHP were associated with decreased testosterone in males (Jurewicz et al. 2013, Main et al. 2006, Meeker et al. 2009, Mendiola et al. 2012, 2011a, Pan et al. 2006), although some studies found no correlation (Jonsson et al. 2005, Lin et al. 2011, Mieritz et al. 2012). There exists limited evidence to suggest an association between circulating testosterone and exposure to phthalates. Most studies found no association between phthalates and FSH or LH (Li et al. 2011, Meeker et al. 2009, Mendiola et al. 2012, 2011a, Pan et al. 2006), although a decrease in FSH (Duty et al. 2005), a decrease in LH (Jonsson et al. 2005), and an increase in LH (Main et al. 2006) have been reported. Therefore, we suggest that the existing literature is inconclusive and the association between phthalate exposure and circulating concentrations of LH and FSH is still unresolved. Similarly, most studies reported no association between urinary concentrations of phthalates and estradiol (Jonsson et al. 2005, Jurewicz et al. 2013, Lin et al. 2011, Mendiola et al. 2011a, Pan et al. 2006). Only two studies reported a decrease in estradiol (Meeker et al. 2009, Mendiola et al. 2012). Based on the preponderance of null results reported to date, we conclude that phthalate exposure is not associated with circulating estradiol. Some disparity in the results from epidemiologic studies of phthalates and reproductive hormones could be due to the fact that not all phthalates were examined in every study. Methodological limitations might also introduce some discrepancies. For example, Li et al. measured concentrations of the phthalate diesters in serum and semen and thus potential contamination cannot be excluded (Li et al. 2011). Similarly, although Main

Crit Rev Toxicol, 2014; 44(6): 467–498

et al. measured concentrations of the phthalate metabolites, quantification was done in breast milk, which is bioactive and thus phthalate metabolites could be artificially increased by contamination with the parent diesters (Main et al. 2006). One placebo-controlled cross-over trial examining the potential association between phthalate exposure and reproductive hormone concentrations was found (Janjua et al. 2007). Men were topically exposed to a cream containing DBP and DEP daily for one week followed by a control week using an uncontaminated cream. The estimated daily exposure was 10 mg/kg/day DBP and DEP, and concentrations of reproductive hormones in serum were measured daily. Statistically significant but minor changes in inhibin B, estradiol, and LH at different time points were found, but the authors attributed these changes to chance as they did not occur consistently (Janjua et al. 2007). Thus, there was no clinically significant association between DBP or DEP and any hormone studied. Overall, the results of this randomized trial contradict those from positive epidemiological studies, raising the possibility that the latter may have been produced by unrecognized biases or unmeasured confounding. Furthermore, the results suggest that human reproductive hormones may not be adversely affected by exposure to phthalates, as the level of exposure in the study was higher than that of the general population. However, these results are interpreted with caution because the sample size of the trial was small, only 26 men, investigators were not blinded to the treatment, assignment of treatment and placebo arms may not have been random, and the exposure period was only one week (Janjua et al. 2007). It is possible that longer exposure periods, or longer washouts between the treatment and control periods, may result in different outcomes. Additionally, the phthalates were absorbed through the skin instead of ingested (the main route of exposure). Despite these concerns, the high exposure level suggests that phthalates do not affect hormone concentrations in a biologically meaningful manner at doses higher than typical human exposure. In summary, the results from the epidemiological studies are mixed, suggesting that the evidence of an association between phthalates and reproductive hormones is weak. Studies in human tissue In vitro studies using human testis explants (both adult and fetal) have also been used to examine the effect of DEHP on testosterone production. However, these studies have produced contradictory results. One found that 10  5 M DEHP and MEHP caused a decrease in production of testosterone from adult testis explants with no change in apoptosis (Desdoits-Lethimonier et al. 2012). On the other hand, in another study, 10  4 M MEHP increased apoptosis but had no effect on testosterone production in fetal human testes (Lambrot et al. 2009). Similarly, 10  5 M MEHP increased apoptosis of germ cells in human fetal testes as demonstrated by increased caspase-3, although testosterone production was not examined (Muczynski et al. 2012). A further two studies which examined the testosterone production of fetal human testes xenografts after treatment with phthalates reported no effect. One found no change in testosterone production from fetal human testes grafted into mice after treatment with DBP, although the testosterone production of explanted fetal rat testes was

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affected (Mitchell et al. 2012). Likewise, the second study grafted rat, mouse, and human fetal testes into rats which were subsequently treated with DBP (Heger et al. 2012). Only the rat fetal testes were affected with respect to testosterone production (Heger et al. 2012); however, multinuclear germ cells were induced in mouse, rat, and human fetal testes (Heger et al. 2012). Although these studies suggest that testosterone production from human fetal testes may be unaffected by phthalate exposure, evidence of multinuclear germ cells and germ cell apoptosis suggests caution is needed.

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Animal studies While epidemiological and biomonitoring studies are essential, animal studies are also important because they permit dosing of genetically homogenous experimental subjects at defined developmental stages through routes of administration relevant to human exposure, to known concentrations of pure test compounds, for defined periods of time, followed by the collection of body fluids and target tissues at key developmental ages; all elements that are frequently unknown or cannot be replicated in human studies. Thus, animal studies are an indispensable tool providing insight into biological plausibility and potential mechanisms of action. Accordingly, animal studies provide experimental data from which no-observed-adverseeffect levels (NOAEL) and lowest-observed-adverse-effect levels (LOAEL) have been established (Table 3) across multiple endpoints. We note that there is a very limited data set for DMP, DEP, dipropyl phthalate, di-n-hexyl phthalate (DnHP), di-isohexyl phthalate, di-n-heptyl-phthalate, di-n-pentyl phthalate (DPP), DnOP, DiNP, DiDP, di-(C(7)-C(9) alkyl) phthalate, and di-(C(9)-C(11) alkyl) phthalate. Several prior exhaustive reviews have documented the adverse reproductive and developmental effects of these phthalates in animals (Kavlock et al. 2006, National Academy of Sciences 2008) and thus will not be discussed herein. Although a robust literature is available for BBP, very few studies have been added to the literature since previous reviews and thus will not be discussed further. In contrast, in recent years, several studies have substantially expanded the DBP, DiBP, and DEHP literature and thus are discussed in detail in the sections that follow. We also note that several papers have been brought forward in the literature describing the reproductive and developmental consequences of exposure to phthalate mixtures, and thus are also captured in the present review.

Physical development and in utero exposure Common outcome measures examined in animal studies included AGD, nipple retention, weight of the sexual organs, and gross malformations of the reproductive tract. AGD was decreased in response to exposure to all phthalates studied with the exception of DMP, DEP, DnOP, and DiDP where no change was found (Table 3). Of the phthalates studied, BBP, DPP, and DnHP were the most potent with AGD being the most sensitive outcome measure of developmental toxicity producing LOAELs of between 100 and 125 mg/kg/day and NOAELs of between 20 and 50 mg/kg/day. Recent studies also reveal that AGD was decreased in Wistar and SpragueDawley rats after in utero exposure to DEHP (Andrade et al.

Reproductive toxicity of phthalates in males 481

2006b, Christiansen et al. 2010, Culty et al. 2008, Gray et al. 2000, 2009, Jarfelt et al. 2005, Moore et al. 2001, Vo et al. 2009, Wilson et al. 2007), DiBP (Boberg et al. 2008, Borch et al. 2006, Saillenfait et al. 2008, 2006, Zhu et al. 2010), and DBP (Scott et al. 2008, Struve et al. 2009, Kim et al. 2010, MacLeod et al. 2010, Jiang et al. 2011). DEHP was more potent with effects at oral doses as low as 100 mg/kg/day on gestational day (GD) 11–21 but not after exposure to 10 mg/ kg/day (Vo et al. 2009). The same dose decreased AGD in another study (Christiansen et al. 2010), whilst reductions in AGD were only detected at higher oral doses (750 or 1,250 mg/kg/day) by other groups (Culty et al. 2008, Moore et al. 2001, Wilson et al. 2007). Developmental exposure to 1–100 mg DBP/kg/day had no effect on AGD in Sprague-Dawley and Wistar rats (Hoshi and Ohtsuka 2009, Scarano et al. 2010). A recent study reported no effect of DEHP on AGD (Pocar et al. 2012) in mice; however, a maximum dose of only 5 mg/kg/day in the diet was used. AGD was also decreased in rats exposed to 250 mg DiBP/kg/day (Boberg et al. 2008, Borch et al. 2006, Saillenfait et al. 2008). Therefore, taken together, the literature shows that AGD is decreased via phthalate exposure with DiBP and DEHP being the most potent. Similar to AGD, phthalate exposure consistently showed an increase in nipple retention (Table 3) in Wistar and Sprague-Dawley rats. Of the phthalates studied, the most potent inducer of nipple retention was DBP with a LOAEL of 100 mg/kg/day. Increased nipple retention was observed after doses of 250 mg DiBP/kg/day (Saillenfait et al. 2008). DEHP and DiBP exposure induced increased nipple retention with LOAELs of 3 and 250 mg/ kg/day, respectively (Andrade et al. 2006b, Gray et al. 2009, Jarfelt et al. 2005). The incidence of abnormal testes in Wistar rats increased at doses of 300 mg DEHP/kg/day or greater (Christiansen et al. 2010). Furthermore, this study reported increased dysgenesis of the external genitalia starting at doses of 3 mg DEHP/kg/ day (Christiansen et al. 2010). Increased incidence of abnormal testes was found in Sprague-Dawley rats following gestational exposure to 100 mg/kg/day DEHP or higher (Gray et al. 2009). Similarly, cryptorchidism was increased after exposure to DEHP (Andrade et al. 2006a, Culty et al. 2008, Moore et al. 2001). However, the doses at which this outcome occurred varied widely. In one study using Wistar rats, three cases of undescended testes were seen after doses as low as 5 mg/kg/ day DEHP, though no increase in the incidence of hypospadias was observed (Andrade et al. 2006a). In another study using Sprague-Dawley rats, cryptorchidism occurred after doses of 938 mg/kg/day DEHP (Culty et al. 2008). Similarly, a dose of 750 mg/kg/day DEHP during gestation increased the incidence of cryptorchidism in Sprague-Dawley rats (Gray et al. 2000). In contrast, increased incidence of cryptorchidism and hypospadias was found in Sprague-Dawley rats exposed to only 500 mg/kg/day DEHP on GD11–21 (Vo et al. 2009). In one comparative study, DEHP was more likely to cause epididymal agenesis in Sprague-Dawley rats and conversely gubernacular malformations in Wistar rats (Wilson et al. 2007), suggesting that DEHP may affect rat strains differently. Finally, DEHP may cause adverse effects at lower doses in subsequent generations. One multigenerational study found that the number of reproductive tract malformations were increased and occurred at lower doses in the second and third generations compared

125 250 ↓ rat G

20 100 ↓ rat C

33 100 ↓ rat G

50 125 ↓ rat G

DiBP

BBP

DPP

DnHP

DBP

DPrP

DEP

AGD 750 – NC rat G 1000 – NC rat C 500 1000 ↓ rat G 100 250 ↓ rat G

Phthalate DMP

rat G 125 250 ↑ rat G

50 125 ↑ rat G

NR

250 750 ↑ rat C

250 500 ↑ rat G

1500 – NC rat G – 250 ↑ rat G,W

Hypospadias 750 – NC rat G NR

50 125 ↑ rat G

NR

250 1000 ↓ rat A,O

200 400 ↓ rat C

167c 250c ↑ rat G NR

300 500 ↓ rat JO

Testes Weight 750 – NC rat G 1000* – NC mouse C 4060* 8600* ↓ mouse C – 3* ↓ rat G,W

250 500 ↑ rat G

1000 1500 ↑ rat G 125b 250b ↑ rat G

Cryptorchidism 750 – NC rat G NR

NR

NR

NR

125 250 ↓ rat G

Prostate Weight 750 – NC rat G 1770* 3640* ↑ mouse C 4060* 8600* ↓ mouse C 50 250 ↓ rat G,W

NR

NR

200 400 ↑ rat C

250 500 ↑ rat G

125 250 ↑ rat G

250 1000 ↑ rat A,O

200 400 ↑ rat C

– 125 ↑ rat G

3* 14–28* ↑ rat G,W

12 50 ↑ rat C

1000 – NC rat C NR

Testes Histology 750 – NC rat G 1000 – NC rat C NR

Sexual Maturitya NR

800* 1670* ↓ mouse C

– 760* ↓ mouse C

200 400 ↓ rat C

NR

Sperm Count 750 – NC rat G 1770* 3640* ↓ mouse C 4060* 8600* ↓ mouse C – 3* ↓ rat G,W

800* 1670* ↓ mouse C

NR

NR

NR

1000 – NC rat C 4060* 8600* ↓ mouse C 50 250 ↓ rat G,W

Sperm Motility NR

NR

250 1000 ↑ rat A,O

NR

NR

40 200 ↑ rat C 4060* 8600* ↑ mouse C – 400 ↑ rabbit G

Sperm Morphology NR

NR

NR

100 500 ↓ rat C

NR

– 250 ↓ rat G

Serum Testosterone – 1000* ↓ rat J 40 200 ↓ rat C NR

NR

– 33 ↓ rat G

NR

– 600 ↓ rat G

10 50 ↓ rat G

Testicular Testosterone – 1000* ↓ rat J – 1000* ↓ rat J NR

NR

100 ppm – NC rat A NR

NR

– 100 ↑ rat G,W

NR

NR

Signs of cancer NR

(Continued)

Lamb et al. (1997c), Saillenfait et al. (2009b)

Lamb et al. (1997b), Hannas et al. (2011a), Heindel et al. (1989), Lindstrom et al. (1988)

Higuchi et al. (2003), Jiang et al. (2007), Lee et al. (2004), Lehmann et al. (2004), Mylchreest et al. (1998), Mylchreest et al. (1999), Mylchreest et al. (2000), Salazar et al. (2004), Scarano et al. (2009), Shono et al. (2003), Zhang et al. (2004) Boberg et al. (2008), Borch et al. (2006), Saillenfait et al. (2006), Saillenfait et al. (2008), Zhu et al. (2010) Aso et al. (2005), Ema et al. (2003), Kohno et al. (2004), Nagao et al. (2000), Tyl et al. (2004);

Lamb et al. (1997a), Heindel et al. (1989), Saillenfait et al. (2011b)

Lamb et al. (1997d), Fujii et al. (2005), Lamb et al. (1987), Oishi et al. (1980)

References Gray et al. (2000), Oishi et al. (1980)

V. R. Kay et al.

100 300 ↑

250 750 ↑ rat C

125 250 ↑ rat G

50 100 ↑ rat G

NR

Nipple Retention 750 – NC rat G NR

Table 3. Summary of the NOAELs and LOAELs in mg/kg/day and the direction of change reported (in that order) for each phthalate and outcome. The animal model and the period of exposure are also noted.

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482 Crit Rev Toxicol, 2014; 44(6): 467–498

220 400 ↓ rat G 500 1000 ↓ rat G 1000 – NC rat G 750 900 ↓ rat G,W 600* – NC rat C NR

DiHP

D79P

DiDP

DiNP

DnOP

DHPP

DEHP

AGD – 250 ↓ rat G 3 10 ↓ rat G,W

Phthalate DcHP

600 750 ↑ rat G,W 600* – NC rat C NR

NR

220 400 ↑ rat G NR

Nipple Retention 100 500 ↑ rat G,W 3 10 ↑ rat G,W

Table 3. (Continued)

NR

NR

NR

NR

NR

220 400 ↑ rat G 1000 – NC rat G NR

Cryptorchidism 750 – NC rat G 1.215 5 ↑ rat G,W

NR

NR

220 400 ↑ rat G NR

Hypospadias 100 500 ↑ rat G,W – 3 ↑e rat G,W

37 375 ↓ rat A 900 – NC rat G,W 300* 600* ↑i rat G,W 350* 800* ↑ rat C

220 400 ↓ rat G NR

– 0.05 ↓ mouse G,W

Testes Weight NR

900 – NC rat G,W 600* – NC rat G,W 350* 800* ↓ rat C

NR

220 400 ↓ rat G NR

Prostate Weight 100 500 ↓ rat G,W 3 10 ↓ rat G,W

300 600 ↑ rat G,W NR

1000* – NC rat G,W 600* – NC rat G,W 800* – NCa,e rat G NR

NR

50 220 ↑ rat G NR

Testes Histology 100 500 ↑ rat G,W – 3* ↑ rat G,W

NR

220 400 ↑ rat G NR

Sexual Maturitya 100 500 ↑ rat G,W 5 15 ↑ rat G,W

400 – NC rat G NR

7500* – NC mouse C 300 600 ↓ rat G,W 600* – NC rat G,W 800* – NC rat G

7500* – NC mouse C 900 – NC rat G,W 600* – NC rat G,W 800* – NC rat G

0.015 0.045 ↓ rat G,W

Sperm Motility NR

– 220 ↓ rat G NR

Sperm Count 15–17* 80–90* ↓ rat C 0.405 1.215 ↓ rat G,W

600* – NC rat G,W 800* – NC rat G

7500* – NC mouse C NR

400 – NC rat G NR

0.015 0.045 ↑ rat G,W

Sperm Morphology NR

NR

– 750 ↓ rat G,W NR

NR

NR

NR

Serum Testosterone 15–17* 80–90* ↑d rat C – 100 ↓f rat G

NR

– 1000* ↓ rat J – 500 ↓ rat G NR

100 300h ↓ rat G NR

– 2000 ↓g rat J

Testicular Testosterone NR

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NR

NR

NR

NR

NR

NR

95 300 ↑ rat A,C

Signs of cancer NR

(Continued)

Willoughby et al. (2000)

Heindel et al. (1989), Oishi et al. (1980), Poon et al. (1997), Saillenfait et al. (2011a) Boberg et al. (2011), Borch et al. (2004), Hannas et al. (2011b), Masutomi et al. (2003); Hushka et al. (2001)

Saillenfait et al. (2011a)

Andrade et al. (2006a), Andrade et al. (2006b), Arcadi et al. (1998), Christiansen et al. (2010), Martinez-Arguelles et al. (2009), Oishi 1985, Pocar et al. (2012), Voss et al. (2005) Hannas et al. (2011b), McKee et al. (2006a)

References Hoshino et al. (2005), Saillenfait et al. (2009a), Yamasaki et al. (2009)

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Reproductive toxicity of phthalates in males 483

V. R. Kay et al.

Crit Rev Toxicol, 2014; 44(6): 467–498

*Dose in mg/kg/day estimated from dietary exposures. aWhere an increase (↑) means that there was a delay in the onset of preputial separation. bMBP, the metabolite of DBP, was administered (Shono et al. 2003). cMBzP, a metabolite of BBP, was administered (Ema et al. 2003). dTrend for an increase in serum testosterone, reached statistical significance (P  0.05) at a dose of 1200 ppm in the diet. eThe outcome measured was mild genital dysgenesis, including only a minor preputial cleft. fIncreases in serum testosterone reported at 0.045, 0.405 and 405 mg/kg/day (Andrade et al. 2006a), as well as at 10 mg/kg/day (Ge et al. 2007), and after inhalation of 1 mg/kg/day (Kurahashi et al. 2005). gTesticular testosterone increased at doses from 10 mg/kg/day PND21–48 (Akingbemi et al. 2001). hEC50 is 443 mg/kg/day. iRelative weight as opposed to absolute weight. jDelayed but no statistically significant difference from controls. A, exposure of adult animals; AGD, anogenital distance; BBP, butylbenzyl phthalate; C, continuous exposure; DBP, dibutyl phthalate; DcHP, dicyclohexyl phthalate; DEHP, diethyl hexyl phthalate; DEP, diethyl phthalate; DHPP, di-n-heptyl phthalate; DiBP, di-isobutyl phthalate; DiDP, di-isodecyl phthalate; DiHP, di-isohexyl phthalate; DiNP, di-isononyl phthalate; DMP, dimethyl phthalate; DnHP, di-n-hexyl phthalate; DnOP, di-n-octyl-phthalate; DPP, di-n-pentyl phthalate; DPrP, dipropyl phthalate; D79P, di-(C(7)-C(9) alkyl) phthalate; D911P, di-(C(9)-C(11) alkyl) phthalate; EC50, half maximum effective concentration; G, gestational exposure; J, exposure of juvenile animals; LOAEL, lowest observed adverse effect level; MBP, mono-butyl phthalate; NOAEL, no observed adverse effect level; NC, no change; NR, not reported; O, one dose; PND, postnatal day; W, exposure of animals before weaning

Nipple Retention NR AGD NR Phthalate D911P

Table 3. (Continued)

Hypospadias NR

Cryptorchidism NR

Testes Weight 300* 700* ↓i rat G

Prostate Weight 700* – NC rat G

Sexual Maturitya 700* – NCj rat G

Testes Histology NR

Sperm Count – 60* ↑ rat G

Sperm Motility 700* – NC rat G

Sperm Morphology 700* – NC rat G

Serum Testosterone NR

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Testicular Testosterone NR

Signs of cancer NR

References Willoughby et al. (2000)

484

with the first (Blystone et al. 2010). In comparison to DEHP, the incidence of undescended testes was increased at doses of 500 and 750 mg DiBP/kg/day in Sprague-Dawley rats (Saillenfait et al. 2006, 2008). Hypospadias was also increased after gestational exposure to 500 mg DiBP/kg/day (Saillenfait et al. 2008). Taken together these data suggest that DiBP affects the male reproductive system in an anti-androgenic manner similar to that of the other phthalates. In comparison to the rat studies, developmental exposure to 500 mg/kg/day of DBP had little effect on the male reproductive tract in marmosets (McKinnell et al. 2009) suggesting that the rat may be more sensitive to adverse effects of phthalates than nonhuman primates. Multiple studies found changes in the weight of the testes in rats and mice after gestational exposure to DEHP (Andrade et al. 2006b, Arcadi et al. 1998, Christiansen et al. 2010, Culty et al. 2008, Gray et al. 2000, 2009, Moore et al. 2001, Pocar et al. 2012, Wilson et al. 2007) and DBP (Johnson et al. 2008, Boekelheide et al. 2009, Kim et al. 2010, MacLeod et al. 2010). Most studies reported a reduction in testes weight in Sprague-Dawley and Wistar rats after DEHP was administered orally to the pregnant rat (Christiansen et al. 2010, Gray et al. 2009, Moore et al. 2001). Testes weight was decreased at 750 mg DEHP/kg/day administered on GD14–18, the only dose used, in SpragueDawley and Wistar rats (Wilson et al. 2007) whereas decreased testes weight was found at 300 mg DEHP/kg/ day administered on GD8-postnatal day (PND) 17 as well as on GD8 to necropsy at PND63–65, but not at lower doses (Gray et al. 2009). Reduced testes weights in LongEvans rats and mice were found at lower doses such as 32.5 ul/L (0.197 0.031 μg/ml DEHP observed in maternal plasma) or even 0.05 mg/kg/day DEHP in the drinking water and in the diet, respectively (Arcadi et al. 1998, Pocar et al. 2012). The lowest dose to decrease testes weight in these studies varied widely from 0.05 to 750 mg/kg/day. Several studies have also reported no change in testes weight after exposure to DEHP in rats at doses of between 500 and 750 mg/kg/day, respectively (Culty et al. 2008, Dalsenter et al. 2006). While divergent results can arise from sample size limitations as well as differences in study design, species and strain of animal used, diet, and housing conditions of the animals, we suggest that age of exposure and the interval between the end of dosing and measurement of effects are critical. For example, no difference in testes weight was found when rats were examined at PND190, although testes weight was decreased at PND22 in the group treated with 750 mg/ kg/day (Jarfelt et al. 2005). Alternatively, it has been suggested that decreased testes weight could also be accounted for by an increased prevalence of cryptorchidism (Mahood et al. 2007). Decreased testes weight was documented in orl rats, a model predisposed to cryptorchidism, given 500 mg/ kg/day DBP whereas wild-type rats were unaffected (Johnson et al. 2008). Finally, one study reported no change in testes weights when Wistar rats were exposed during gestation and examined on PND144 (Andrade et al. 2006a). Furthermore, a study in Wistar rats found an increase in testes weights at low doses up to 135 mg/kg/day in addition to the expected decrease in testes weight at a higher dose of 405 mg/kg/day (Andrade et al. 2006b).

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Exposure to DEHP as a juvenile can also affect testes weight. In one study, exposure to 500 mg/kg/day on PND21– 48 reduced testes weight (Ge et al. 2007). In another study, testes weight was decreased by 750 mg/kg/day administered for 30 days to 21-day-old Wistar rats (Botelho et al. 2009b). Similarly, testes weight was decreased by exposure to 900 mg/kg/day DEHP on PND22–58 in Sprague-Dawley, but not Long-Evans rats (Noriega et al. 2009), findings that are harmonious with previous studies (Agarwal et al. 1986, Gray et al. 1977, Nabae et al. 2006, Cammack et al. 2003, David et al. 2000a, 2000b, Oishi and Hiraga 1980, Oishi 1985, 1986, Park et al. 2002, Parmar et al. 1986). On the other hand, testes weights were significantly increased following 500 mg DiBP/kg/day in rats but only at a dose of 1000 mg/ kg/day in mice suggesting that the sensitivity of the species differs (Zhu et al. 2010). In another study, testes weight was unaffected by doses up to 625 mg/kg/day DiBP when cryptorchid testes were excluded (Saillenfait et al. 2008). Still, decreases in the weight of the prostate, epididymis and seminal vesicles were noted at doses of 500 mg/kg/day (Saillenfait et al. 2008). Decreases in the weights of the accessory sex organs were also reported in rats (Christiansen et al. 2010, Dalsenter et al. 2006, Gray et al. 2000, 2009, Moore et al. 2001, Pocar et al. 2012, Wilson et al. 2007). The seminal vesicle weight in mice was decreased after in utero exposure to doses as low as 0.05 mg/kg/day DEHP in the diet (Pocar et al. 2012). In contrast, one study reported an increase in bulbourethral gland weight and maturity in three-week-old pigs treated with DEHP for four weeks (300 mg/kg/day thrice a week; Ljungvall et al. 2008). However, this was an isolated statistical difference and the study used t-tests to examine multiple outcomes and thus the level of significance employed was inadequately conservative. Decreases in androgen sensitive target tissues including the levator ani and bulbocavernosus (LABC) muscle, prostate, and epididymis were found in rats after exposure to 300 or 500 mg/kg/day DEHP during gestation (Gray et al. 2009). Conversely, a study in Wistar rats reported no changes in the weight of the epididymis at doses up to 405 mg/kg/day DEHP (Andrade et al. 2006b). Similarly, there was no change in the weights of the epididymis or prostate, although the seminal vesicle weight was decreased, in adult rats exposed to 405 mg/kg/day DEHP during gestation (Andrade et al. 2006a). In 21-day-old Wistar rats administered 750 mg/kg/day DEHP for 30 days, seminal vesicle and LABC weights decreased while epididymis weights increased (Botelho et al. 2009b). Exposure to 750 mg/kg/day DEHP for 28 days starting at PND21 also reduced the weight of the prostate (Ge et al. 2007). Likewise, exposure starting at PND22 decreased the weights of the epididymis, seminal vesicles, Cowper’s glands, prostate and LABC at doses ranging from 100 to 900 mg/kg/ day DEHP (Noriega et al. 2009), findings that agree with prior studies (Agarwal et al. 1986, Oishi 1985, 1986). We took note of a study which revealed that exposure to 25,000 ppm DEHP in the diet in conjunction with kidney failure (34.5 18.6 μg/ml MEHP measured in blood) decreased the weights of the epididymis, prostate, and seminal vesicles (Nabae et al. 2006). These data suggest that metabolism of DEHP together with impaired clearance of metabolites enhances phthalate toxicity.

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Sexual maturation Although epidemiological studies in humans have shown no change in puberty as a result of phthalate exposures including exposure to DEHP, animal studies reveal delayed sexual maturation at high oral doses of DEHP (300 mg/kg/day and greater) and DBP (50–250 mg/kg/day) as measured by preputial separation in Wistar, Sprague-Dawley, and Long-Evans rats (Andrade et al. 2006b, Gray et al. 2009, Noriega et al. 2009, Kim et al. 2010). DiBP also caused a delay in the onset of sexual maturity, as measured by the day of preputial separation, in Sprague-Dawley rats starting at oral doses of 500 mg/kg/day (Saillenfait et al. 2008). Preputial separation was delayed at 300 mg/kg/day DEHP in Long-Evans rats but at 900 mg/kg/day in Sprague-Dawley rats (Noriega et al. 2009). A trend for delayed preputial separation has also been reported (Botelho et al. 2009b) in 21-day-old Wistar rats dosed with 250–750 mg/kg/day DEHP for 30 days. Interestingly, in pups exposed only during gestation and lactation, there was no effect on preputial separation while delayed maturation was found in pups with ongoing exposure (Gray et al. 2009). In contrast, no effect of 20–500 mg/kg/day DEHP on the onset of maturity in Wistar rats was found despite other effects on the male reproductive system (Dalsenter et al. 2006) which is harmonious with prior studies using doses of DEHP up to 750 mg/kg/day (Gray et al. 2000, Jarfelt et al. 2005, Moore et al. 2001). Interestingly, one study found that sexual maturity occurred earlier when low doses of DEHP (10 mg/kg/day) were administered to Long-Evans rats (Ge et al. 2007). However, two other studies contradicted this result, revealing that low oral doses of DEHP (15 mg/kg/day and greater) delayed the onset of maturity in Wistar rats whilst the other found that low doses of DEHP (10 mg/kg/day) had no effect on the onset of maturity in Sprague-Dawley and Long-Evans rats (Andrade et al. 2006b, Noriega et al. 2009). Overall, while the majority of the studies suggest that sexual maturation is delayed at high doses, the effect of low doses of DEHP is ambiguous.

Semen quality Several studies examined the effect of fetal and neonatal exposure to DEHP on sperm production later in life. Declines in daily sperm production have been described in Wistar rats and in cauda epididymal sperm counts in mice at low doses including 1.215 and 0.05 mg/kg/day (Andrade et al. 2006a, Pocar et al. 2012). A dose of 0.045 mg/kg/day, but not higher doses, also resulted in an increased number of animals with abnormal sperm (Andrade et al. 2006a). However, we considered this to be a potentially spurious finding owing to the lack of an effect at the higher doses. More interestingly, sperm taken from mice treated with 0.05 and 5 mg/kg/day DEHP were less able to form viable embryos in vitro (Pocar et al. 2012). A significant decrease in cauda epididymal sperm counts in Sprague-Dawley rats after doses of 750 mg/kg/day DEHP and greater, as well as nonsignificant decreases in daily sperm production after a dose of 375 mg/kg/day has also been reported (Moore et al. 2001). Another study in Sprague-Dawley rats reported decreased sperm concentration, motility and viability after exposure to 10 and 500 mg/kg/day but not 100 mg/kg/ day DEHP on GD11–21 (Vo et al. 2009). Lastly, decreased

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epididymal sperm counts and daily sperm production was described in Wistar rats after exposure to 500 mg/kg/day DEHP but no change in the occurrence of abnormal morphology (Dalsenter et al. 2006). Spermatogenesis was unaffected by oral doses of DEHP up to 1000 mg/kg/day when the number of germ cells at each stage of spermatogenesis in the testes in Sprague-Dawley rats was counted (Shirota et al. 2005). Additionally, no effect of DEHP on either caudal epididymal or testicular sperm counts in Sprague-Dawley rats was detected after doses of DEHP up to 600 mg/kg/day following both intravenous and oral routes of exposure (Cammack et al. 2003). Similarly, no effect of either 300 or 750 mg/kg/day DEHP in Wistar rats on sperm counts or motility was reported when a small number of severely affected outliers were excluded (Jarfelt et al. 2005). The above studies all exposed male rats to DEHP prenatally and/or neonatally. This period of exposure may result in more severe effects as the reproductive system is still developing; however, as the outcomes can only be measured after maturity, the treated animals may recover before the target outcomes can be measured. For example, a separate series of studies in boars reported no effect of DEHP exposure from 3 to 7 weeks of age on sperm counts, motility, morphology and function measured at 6 months of age (Spjuth et al. 2006a, 2006b, 2007a, 2007b), results likely due to the length of time between exposure and measurement of outcomes. Therefore, the lack of an effect reported by Spjuth et al. is likely the consequence of recovery. A study conducted in marmosets also found no effect of DEHP on the histology of the testes or on testicular sperm counts (Tomonari et al. 2006). This result may be due to the species used and is encouraging as marmosets are potentially a better model species than rats for effects in humans. On the whole, however, it is apparent that DEHP can negatively affect spermatogenesis and fertility in animals at high doses, while the effects of low doses are ambiguous. Testicular morphology Wistar rats exposed to 600 mg/kg/day DiBP during gestation had an increased number of gonocytes, Sertoli cell vacuolization, and Leydig cell clustering at GD21 (Borch et al. 2006). Similarly, the number of Sertoli cells was decreased and Sertoli cell vacuolization was seen in rats treated with DBP (100–750 mg/kg/day) during gestation (Alam et al. 2010a, 2010b, Auharek et al. 2010, Bao et al. 2011, Chen et al. 2011, Ryu et al. 2007, 2008, Scott et al. 2008), whereas no differences were seen in marmosets with exposure up to 500 mg/ kg/day (McKinnell et al. 2009). Increased apoptosis of spermatogenic cells as well as an increase in the disorganization of vimentin filaments in the testes was found in prepubertal male rats treated orally with 500 mg/kg/day DiBP (Zhu et al. 2010). In mice, however, there was no increase in apoptosis of spermatogenic cells at doses up to 1000 mg/kg/day suggesting that mice are less sensitive to the effects of DiBP (Zhu et al. 2010). There was a dose-dependent increase in damage to the seminiferous tubules leading to oligospermia and azoospermia reported at all doses used (125, 250, 500, and 625 mg/kg/day DiBP) in adult Sprague-Dawley rats (Saillenfait et al. 2008). Degeneration of the seminiferous epithelium and disorganization of the seminiferous tubules was also documented in

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DBP-treated rats (Kim et al. 2010, Kondo et al. 2006, Ryu et al. 2007, 2008, Zhou et al. 2010). Taken together, we suggest that DiBP and DBP exposure can adversely affect the testes leading to decreased sperm counts. Studies using DBP add evidence that phthalates negatively affect the cellular structure of the testes with adverse effects on spermatogenesis and potentially on tumor-initiation. One study reported decreased proliferation, delayed differentiation, and aggregation of germ cells after in utero exposure to DBP (Jobling et al. 2011). The changes to differentiation suggest similarities with carcinoma-in-situ cells related to testicular cancer. In a second study, DBP induced multinucleated germ cells (Saffarini et al. 2012). These cells persisted only in the adulthood of p53 null mice (Saffarini et al. 2012); however, despite the absence of p53, the cells did not display markers of testicular cancer (Saffarini et al. 2012). In addition to its direct effect on germ cells, DBP also affects the Sertoli cells. DBP caused disorganization of the vimentin filaments leading to detachment and apoptosis of the spermatogenic cells (Alam et al. 2010a). Hyperplasia and aggregation of cells in the testes were also outcomes of interest. In Sprague-Dawley rats, DEHP increased hyperplasia of interstitial cells and aggregation of multinucleated germ cells at doses of 250 mg/kg/day administered on GD7–18 (Shirota et al. 2005). This study also reported damage to seminiferous tubules after in utero exposure to 1000 mg/kg/ day and an increase in multinucleated germ cells after doses of 500 mg/kg/day or higher (Shirota et al. 2005). Another study using Sprague-Dawley rats reported hyperplasia and clustering of Leydig cells after doses of 938 mg/kg/day DEHP or higher (Culty et al. 2008). Similarly, Leydig cell hyperplasia and clustering was also reported in Wistar rats treated with 500 mg/kg/day DBP either during gestation or during the prepubertal period (Mahood et al. 2007, Scarano et al. 2010, Bao et al. 2011). Furthermore, an increase in the number of multinucleated germ cells was reported in Wistar rats after in utero exposure to 135 and 405 mg/kg/day DEHP (Andrade et al. 2006b). However, in marmosets, no effect of 5000 mg/kg/day DEHP on the histology of the testes was seen (Kurata et al. 1998). In mice, increased apoptosis and sloughing of germ cells in the seminiferous tubules after treatment with 700 and 800 mg/kg/day MEHP has been reported (Tay et al. 2007a). The disappearance of most germ cells after treatment with 2% DEHP in the diet was also reported in mice (∼26 mg/kg/day; Miura et al. 2007). In rats, 300 mg/kg/day DEHP decreased the diameter of the seminiferous tubules and caused an apparent delay in testes development (Christiansen et al. 2010). In separate study, 2000 mg/kg/day DEHP damaged the seminiferous tubule epithelium, resulting in stalled spermatogenesis but had no effect on Leydig cells (Park et al. 2002). A decrease in Sertoli cell proliferation and an increase in multinuclear gonocytes were clearly demonstrated in one study using PND3 SpragueDawley rats (Li et al. 2000). Decreased expression of cyclin D2 was reported in the affected Sertoli cell population, despite an absence of change in serum FSH concentrations (Li et al. 2000). Importantly, signs of increased proliferation were noted 48 hours after dosing ceased, perhaps indicating recovery (Li et al. 2000). We postulate that the divergent effects likely arise from differences in test animals, strain, and age at the time of exposure, and the outcome measures used. For example,

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one study reported a decreased number of Sertoli cells after treating young rats but not after treating adults (Dostal et al. 1988). Despite ambiguity about the doses which negatively affect the testes, these studies show that DEHP is capable of producing adverse effects on testicular cells which ultimately affect spermatogenesis.

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Reproductive hormones Decreased serum testosterone concentrations were reported in Sprague-Dawley rats and rabbits treated with high concentrations of DBP (250–750 mg/kg/day) during gestation (Jiang et al. 2007, Nair et al. 2008, Kim et al. 2010, Bao et al. 2011); however, others have been unable to document changes in circulating testosterone concentrations albeit with lower concentrations of DBP and measurement at later developmental time points (Ryu et al. 2007, Scarano et al. 2010). DBP treatment also decreased intratesticular testosterone concentrations in Sprague-Dawley and Wistar rats (Mahood et al. 2007, Scott et al. 2008, Struve et al. 2009, MacLeod et al. 2010). Developmental exposure to 100 mg/kg/day DBP had no effect on intratesticular testosterone concentrations measured on PND90 (Scarano et al. 2010) potentially owing to either the low dose of DBP used or recovery of the testis. In comparison, gestational exposure to DBP had no effect on circulating concentrations of testosterone measured at birth in marmosets (McKinnell et al. 2009). The effect of DEHP at different concentrations on circulating testosterone concentrations is unclear. Several studies have reported decreases in testosterone after either gestational (Akingbemi et al. 2001, Culty et al. 2008, Martinez-Arguelles et al. 2009, Vo et al. 2009, Wilson et al. 2007) or adult exposure to DEHP. Serum testosterone decreased after gestational exposure to DEHP (100 mg/kg/day GD14–19 or GD12–21) in Long-Evans and Sprague-Dawley rats (Akingbemi et al. 2001, MartinezArguelles et al. 2009). Similarly, serum testosterone was decreased after doses of 234 mg/kg/day or higher, and testosterone production was reduced by doses of 938 mg/kg/day or higher while serum estradiol concentrations were unaffected in Sprague-Dawley rats exposed from GD14 to birth (Culty et al. 2008). Decreased serum testosterone was also found after exposure to 500 mg/kg/day but not 10 or 100 mg/kg/ day on GD11–21 in Sprague-Dawley rats (Vo et al. 2009). In contrast to the above findings, serum testosterone concentrations were increased in response to developmental exposure (GD6–PND21) to oral doses of 0.045, 0.405, and 405 mg/kg/ day but the measures were made in adult Wistar rats (Andrade et al. 2006a). Thus, temporal disconnection between exposure and measurement of circulating testosterone concentrations complicates comparisons across studies. In adult animals exposed to DEHP, decreased serum testosterone concentrations have also been reported (Agarwal et al. 1986, Miura et al. 2007, Oishi 1985, 1986, Akingbemi et al. 2001). In Wistar rats, the concentrations of serum testosterone and testicular testosterone were not significantly different but there was a trend for decreased serum testosterone after oral exposure to 750 mg/kg/day DEHP for 30 days (Botelho et al. 2009b). While these studies clearly suggest a decrease in testosterone and therefore an anti-androgenic effect of phthalates, other studies have reported the opposite effect. Circulating testosterone concentrations were

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decreased by postnatal exposure (PND21–48) to 750 mg DEHP/ kg/day but increased at a dose of 10 mg/kg/day in Long-Evans rats (Ge et al. 2007). Exposure to MEHP decreased testosterone production in vitro by a dose of 10 mM MEHP whilst doses of 100 μM and 1 mM MEHP increased testosterone production in vitro (Ge et al. 2007). These studies suggest that DEHP may have different effects at high and low doses. However, the effect of high and low DEHP doses on serum testosterone concentrations yielded nonsignificant decreases at 300 mg/kg/day and 900 mg/kg/day but no increase at lower doses (10 mg/kg/day) in both Long-Evans and Sprague-Dawley rats exposed PND22 until necropsy (Noriega et al. 2009). This study also reported significant decreases in testosterone production after doses of 300 and 900 mg/kg/day (Noriega et al. 2009) adding to the evidence that high doses result in decreased circulating testosterone concentrations. Based on these studies, the effect of low doses of DEHP on testosterone levels is unclear; however, the majority of studies agree that high doses of DEHP cause decreases in circulating testosterone concentrations in animals exposed during development and as juveniles or adults. Two studies examined the changes in steroidogenic proteins and testicular testosterone in pups after in utero exposure to DiBP. Both studies reported a decrease in the expression of proteins important to the production of testosterone, including steroidogenic acute regulatory protein (StAR) and P450scc, as well as decreases in testicular testosterone and testosterone production after treating Wistar rats with 600 mg/kg/day DiBP during gestation (Boberg et al. 2008, Borch et al. 2006). In contrast, other studies have found no effect of DEHP on testosterone concentrations (Andrade et al. 2006b, Dalsenter et al. 2006, Gray et al. 2009, Kurata et al. 1998, Ljungvall et al. 2005, 2006, Stroheker et al. 2006, Tomonari et al. 2006). For example, no changes in testicular serum testosterone concentrations after Wistar rats were exposed to doses as high as 405 mg/kg/day during gestation and before weaning were reported (Andrade et al. 2006b). Two of these studies were conducted in marmosets and found no change in testosterone concentrations, potentially because marmosets are less sensitive to phthalates (Kurata et al. 1998, Tomonari et al. 2006). Interestingly, in a study conducted in boars (Ljungvall et al. 2005), there was no change in testosterone or LH concentrations during exposure but an increase in plasma testosterone at 7.5 months suggesting that the effects of DEHP may be reversible. A recovery period before examination may also explain the lack of effect on testosterone in one of the other studies (Dalsenter et al. 2006). The effects of DEHP on FSH and LH are equally uncertain. One study reported increases in FSH and LH mRNA in the pituitaries of mice after exposure to 5 mg/kg/day DEHP in the diet (Pocar et al. 2012). Furthermore serum LH and FSH were increased in rats exposed to 20,000 ppm in the diet (∼1200 mg/kg/day; Agarwal et al. 1986). Similarly, an increase in LH was found in Sprague-Dawley but not Long-Evans rats treated with 900 mg/kg/day DEHP (Noriega et al. 2009) consistent with a previous study (Akingbemi et al. 2004). Conversely, decreased LH was found after exposure to 100 mg/kg/day on GD12–21 and after exposure to 10 mg/kg/day or greater on PND21–48 (Akingbemi et al. 2001). In this study, LH was unchanged by exposure up to 200 mg/kg/day when administered on PND1–21, PND21–34, PND35–48, or PND62–89

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(Akingbemi et al. 2001) suggesting that fetal development is a critical window for the adverse effects of DEHP. Another study reported decreased serum LH after exposure to 500 mg/ kg/day on GD11–21 in Sprague-Dawley rats (Vo et al. 2009). No change in the amount of LH mRNA in the pituitary was found in rats treated with doses of DEHP up to 750 mg/kg/ day (Ge et al. 2007). Overall, the effects of DEHP on FSH and LH concentrations in serum are ambiguous, potentially owing to differences in window of exposure, doses used, route of exposure, and animal model employed. We suggest that DEHP exposure, at high doses administered during fetal development decreases serum LH concentrations whereas exposure of adult animals results in ambiguous effects on serum FSH and LH concentrations.

Mixtures Up to this point, the phthalates and their effects on the reproductive system have been discussed separately. However, the different diesters and their metabolites result in similar effects and likely share mechanisms of action. Additionally, humans are typically exposed to many phthalates concurrently (Blount et al. 2000, Calafat and McKee 2006b, Koch and Calafat 2009, Wittassek et al. 2011). Thus, the effects of phthalate mixtures are important. Five studies have examined the effects of combinations of phthalates compared to those of the phthalates alone. Two examined a combination of DBP and DEHP (Howdeshell et al. 2007, Martino-Andrade et al. 2009). In one, the effects of 500 mg/kg/day DEHP, 500 mg/kg/day DBP and a mixture of 500 mg/kg/day DBP and 500 mg/kg/day DEHP were compared (Howdeshell et al. 2007). Although all three treatments affected the male reproductive system, the effects on AGD and nipple retention were more pronounced in the animals treated with the mixture (Howdeshell et al. 2007). Furthermore, the incidence of hypospadias, undescended testes, or other malformations of the male reproductive system was highest in the group treated with the mixture (Howdeshell et al. 2007). Finally, although testosterone was decreased by all treatments, the largest decrease occurred with the combination mixture (Howdeshell et al. 2007). Similarly, the other study compared a mixture of 150 mg/kg/day DEHP and 100 mg/ kg/day DBP to 150 mg/kg/day DEHP, 100 mg/kg/day DBP, and 500 mg/kg/day DBP individually (Martino-Andrade et al. 2009). In this study, the mixture of DBP and DEHP caused a decrease in testosterone and changes in the histology of the testes similar to the effects caused by 500 mg/kg/day DBP (Martino-Andrade et al. 2009). On the other hand, the mixture and the lower doses of DBP and DEHP did not affect AGD or nipple retention, which were decreased and increased respectively at the 500 mg/kg/day DBP dose (Martino-Andrade et al. 2009). Based on these two studies, DBP and DEHP can act in an additive manner for some, but not all endpoints. A third study also reported a tendency for DEHP and DiNP to act additively when administered as a binary mixture (Borch et al. 2004). Furthermore, a fourth study compared the effective dose in at least 50% of animals (ED50) for a mixture of DEHP, DBP, DPP, DiBP, and BBP to the ED50s for the individual phthalates (Howdeshell et al. 2008). The ED50s identified for DEHP, DBP, DiBP, and BBP were 300 mg/kg/day while the ED50 for DPP was 100 mg/kg/day (Howdeshell et al. 2008).

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The ED50 for the mixture was 260 mg/kg/day showing that the effects of the phthalates were modestly additive (Howdeshell et al. 2008). Finally, one study examined a mixture of nine phthalates and reported significant decreases in testicular testosterone production at a 17% dilution of the 650 mg/kg/ day mixture (Hannas et al. 2011b). Overall, these studies show that a combination of phthalates can negatively affect the reproductive system in an additive fashion. Although human exposure to individual phthalates is low, the combined effects of the different phthalates may be able to cause biologically important effects. As these studies suggest that even the less potent phthalates can contribute to adverse effects on the reproductive system when combined with other phthalates, human exposure to all phthalates must be considered. Additional animal studies are needed to determine the LOAELs and NOAELs for mixtures of the common phthalates in order to predict the effects of exposure in humans. Mixtures could be formulated in several ways. For example, phthalates could be combined in a one-to-one ratio, proportionately based on the ED50 or proportionately based on the concentrations of phthalates to which humans are typically exposed. Ideally, the mixture studied should reflect the relative concentrations of phthalates seen in human biomonitoring studies. The likelihood that phthalates act additively with each other and with other anti-androgen compounds is comprehensively covered in a monograph from the U.S. National Academy of Sciences with the conclusion that the effects of mixtures needs to be a focus of future research and considered when evaluating the risks of phthalate exposure (2008).

Perspective The above studies provide further support for the biological plausibility that exposure to phthalates can adversely affect the development of the male reproductive tract and the morphology of the testes, delay the onset of sexual maturity, adversely affect spermatogenesis, and induce changes in circulating concentrations of reproductive hormones. Consistency of results across endpoints in the animal literature provides an added measure of confidence of the anti-androgenic effects of phthalates at high doses. Of the reported adverse effects of phthalates, the effects on spermatogenesis were the most sensitive and the literature points to the developing fetus as a critical window for phthalate exposure. Furthermore, the recent literature suggests that DEHP is the most potent phthalate with effects on spermatogenesis the critical outcome measure. However, the animal literature suggests that the adverse effects of phthalates on the reproductive tract are potentially reversible. Modes of action The mode of action and mechanism for the effects of phthalates on male reproductive health has not yet been fully elucidated. While the mechanisms are mainly unknown, there is evidence implicating several unique mechanisms. These distinct mechanisms may account for the effect of phthalate exposure on multiple aspects of the reproductive system leading to a complex syndrome (like TDS) which consists of seemingly unrelated outcomes. The early literature observed phthalate-induced changes in zinc metabolism after exposure to high concentrations of phthalates including DEHP,

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DBP, and DPP with decreased testicular zinc concentrations (Foster et al. 1980), increased urinary excretion of zinc (Cater et al. 1976, 1977, Foster et al. 1981, 1982), and decreased half-life of zinc in the testes after phthalate exposure (Cater et al. 1976, 1977). Zinc is known to be essential to spermatogenesis (Yamaguchi et al. 2009) and moderate zinc deficiency has been shown to affect spermatogenesis in both mice and humans (Abbasi et al. 1979, Croxford et al. 2011). Although recognized as an essential trace element important in spermatogenesis, the mechanisms underlying the adverse effects of phthalate exposure on testicular zinc concentrations and potentially other trace elements are unclear. Furthermore, the protective effect of essential trace elements may be important in regions of the world where intake is inadequate. For example, the adverse effects of DEHP on testicular morphology, spermatogenesis, and circulating testosterone concentrations were more pronounced in selenium deficient rats whereas selenium supplementation was found to be protective (Erkekoglu et al. 2011, 2012). These data suggest that trace element supplementation could be important in at risk populations and could also aid in understanding divergent results across studies. Decreased steroidogenesis and circulating testosterone concentrations are thought to be central to phthalate-induced reproductive toxicity. Phthalate-induced changes in reactive oxygen species (ROS) have been proposed to cause changes in steroidogenesis of Leydig cells (Zhao et al. 2012). Decreased testosterone production and increases in ROS were reported after treatment with 20–2000 μM MEHP (Zhao et al. 2012), which is consistent with the findings of earlier studies (Oishi 1990, Saxena et al. 1985). Increased superoxide dismutase (SOD) activity, decreased glutathione peroxidase (GPx) activity, and increased malondialdehyde in the epididymis after exposure to 500 mg/kg/day DBP provide additional support for the importance of oxidative stress in the mode of phthalate actions (Zhou et al. 2011). Likewise, SOD and GPx activities were decreased while malondialdehyde was increased in the testes after exposure to 250 mg/kg/day DBP (Zhou et al. 2010). Similarly, mitochondrial peroxiredoxin and cyclooxygenase-2 expression are increased in germ cells after phthalate exposure, suggesting a response to oxidative stress (Onorato et al. 2008). Increased ROS were also reported in Leydig cells after MEHP or DEHP treatment with a protective antioxidant effect of selenium supplementation (Erkekoglu et al. 2010). Interestingly, no protective effect but rather an adverse effect of concurrent DEHP and vitamin C treatment was reported (Botelho et al. 2009a). The mechanisms underlying phthalateinduced changes in ROS enzyme expression and activity are unclear. Numerous studies have provided evidence for decreased expression of proteins important in the steroidogenic pathway including scavenger receptor class B member 1(SR-B1), StAR, p450scc, 3-beta-hydroxysteroid dehydrogenase (3βHSD), 17-beta-hydroxysteroid dehydrogenase (17βHSD), and cytochrome P450 17A1 (CYP17a1) among others (Akingbemi et al. 2001, Barlow et al. 2003, Borch et al. 2006, Chauvigne et al. 2011, Gunnarsson et al. 2008, Johnson et al. 2011, Kuhl et al. 2007, Lehmann et al. 2004, Wang et al. 2007). On the other hand, several studies have also reported increased expression of steroidogenic proteins,

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with or without a concurrent decrease in steroid hormone production (Chauvigne et al. 2011, Ryu et al. 2007a, Wang et al. 2006). Overall, the evidence indicates that steroidogenesis is affected by phthalate exposure although the cellular and molecular mechanisms remain uncertain. Studies have variously suggested a glucocorticoid-dependent mechanism, action through sterol-regulatory element-binding protein (SREBP2) or activation of peroxisome proliferator-activated receptor gamma (PPARγ; Gunnarsson et al. 2008, Johnson et al. 2011, Kobayashi et al. 2003, Xiao-feng et al. 2009). Studies examining gene expression after exposure to DBP recently identified changes in the pathways responsible for cholesterol transport and metabolism as important (Liu et al. 2005, Plummer et al. 2007, Euling et al. 2013, Ovacik et al. 2013, Plummer et al. 2013). Other pathways, ranging from the cell cycle to energy metabolism were also identified as affected by treatment (Ovacik et al. 2013). Increased expression of growth factors, including insulin-like growth factor 1 (IGF1) and mast/stem cell growth factor receptor (c-KIT) ligand, has also been suggested as a potential upstream mechanism of changes in steroidogenesis (Lin et al. 2008). Furthermore, increased aggregation and decreased numbers of fetal Leydig cells were noted in this study (Lin et al. 2008). Leydig cell aggregation was linked to increased expression of leukemia inhibitory factor (Lin et al. 2008). In a separate study, hyperplasia of Leydig cells was connected to increased expression of phospholipase D and to phosphorylated ERK1/2 (Lee et al. 2011). Overall, the major effect on Leydig cells is a change in steroidogenesis. However, the mechanism leading to these changes is unclear. Changes in steroidogenesis, in particular a decrease in testosterone synthesis, are implicated in malformations of the reproductive system after phthalate exposure. Changes in the expression of other proteins involved in development may also be important. Decreased expression of 5α-reductase but also androgen receptor and sonic hedgehog were reported after DBP exposure (Kim et al. 2010). Decreased expression of insulin-like factor 3 was found in MBP-treated cultures of fetal rat testes (Chauvigne et al. 2011). A decrease in these proteins could contribute to impaired development of the male reproductive tract. An effect of phthalates on germ cell apoptosis has also been suggested. Increased membrane localization of Fas and apoptosis of germ cells after DEHP exposure has been reported (Ichimura et al. 2003, Richburg et al. 1999). Furthermore, these apoptotic germ cells were observed to associate with FasL expressing Sertoli cells (Ichimura et al. 2003). Subsequently, gld mice (expressing a mutant FasL that cannot bind Fas) were partially protected from DEHP- or MEHPinduced germ cell apoptosis (Giammona et al. 2002, Lin et al. 2010) further implicating the apoptosis pathway in phthalateinduced toxicity. Detachment of the germ cells from the seminiferous epithelium is another common outcome observed after phthalate exposure. Interestingly, it has been shown that the detachment of the germ cells is independent from apoptosis (Yao et al. 2010). The detachment of the cells is likely related to the loss of cellular adhesions between Sertoli cells. Multiple recent studies have shown changes in the expression and localization of structural proteins including connexin 43, cluadin-11, occludin, zonula occludens as well as some cadherins after phthalate

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exposure (Chiba et al. 2012, Kang et al. 2002, Sobarzo et al. 2006, 2009, Yao et al. 2010, Zhang et al. 2008). Additionally, changes in vimentin including decreased expression and length have been reported as a result of phthalate treatment (Tay et al. 2007b). Decreases in cluadin-11 and occludin expression have been proposed to be the result of p44p42MAPK activation (Chiba et al. 2012), whilst increased matrix metalloprotease (MMP2) activity has been suggested to degrade Sertoli cell tight junctions (Yao et al. 2010). Increased MMP2 activity was also linked to decreased expression of a tissue inhibitor of metalloproteases (TIMP2; Yao et al. 2011). Moreover, interference with FSH-signaling was implicated in the decrease of TIMP2 expression in addition to activation of MYC, an inhibitory transcription factor (Yao et al. 2011). Inhibition of FSHsignaling by MEHP also played a role in decreasing Sertoli cell proliferation during rat development (Li et al. 1998). In this study, MEHP was also able to interfere with proliferation as a result of exogenous cAMP (Li et al. 1998). While several studies have shown impaired FSH-induced cAMP signaling in Sertoli cells as decreased cAMP accumulation after in vitro treatment with MEHP (as well as MBP and MPP but not MEP, MMP, or MPrP; Grasso et al. 1993, Heindel and Chapin 1989, Heindel and Powell 1992, Lloyd and Foster 1988), the mechanism was reported as impaired binding of FSH to Sertoli cell membranes (Grasso et al. 1993). Limitations of animal studies To date, the animal studies clearly show that phthalates can negatively impact the male reproductive system. However, differences in comparative physiology and endocrinology cannot be overlooked when translating results from rodents to humans. In rats and mice, DEHP and other phthalates act as peroxisome proliferators by activating the peroxisome proliferator-activated receptor alpha (PPARα) pathway, which may be less relevant in humans (Ito and Nakajima 2008, Latini et al. 2008, Rusyn and Corton 2011). For example, while rats experience liver damage and tumors after treatment, this is not thought to occur in humans (Ito and Nakajima 2008, Rusyn and Corton 2011). Likewise, PPARα activation may be important in the reproductive toxicity of phthalates (Latini et al. 2008). However, in PPARα knockout mice, there were no effects in the liver after exposure to DEHP although testes weight was still decreased (Ward et al. 1998). Interestingly, the time at which histological changes were noted in the testes differed between wild-type and knock-out mice (Ward et al. 1998). There may be other pathways, including PPARγ activation, leading to reproductive toxicity (Ryu et al. 2007b). Therefore, the reproductive effects of DEHP seen in animal studies may be relevant to humans, despite the differences in sensitivity to peroxisome proliferators. It is noteworthy that the effects of phthalate seen in marmosets are typically less severe than those seen in rats. Therefore, it may be concluded that marmosets, and by extension humans, are less sensitive to phthalates. Compared to rats and mice, marmosets had lower constitutive expression of PPARα in the liver (Ito et al. 2007). The difference in PPARα expression supports the use of marmosets as a model for effects in humans. Still, whether the absorption of phthalates in marmosets closely resembles the absorption in humans is an important question, as the increased resistance to phthalates could

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be due to a decreased concentration of phthalate metabolites reaching the reproductive system. One study on metabolism in humans reported 90% of the 0.047 mg/kg dose was excreted as metabolites in urine in the first 24 hours (Anderson et al. 2011), while a study examining metabolism in juvenile marmosets reported only 18% of a 100 mg/kg dose excreted in urine (Kurata et al. 2012a). While the difference in excretion could be due to differences in dose, it could also indicate a difference in absorption, distribution, metabolism, and excretion processes in marmosets compared to those of humans. Conversely, marmosets and humans are similar in that they excrete glucuronide-conjugated metabolites while rats do not (Kurata et al. 2012b). Overall, marmosets may be a better model of the true effects of phthalates in humans. Critical windows of exposure Animals are more susceptible to the effects of phthalates if exposed in utero during development of the reproductive system (Blystone et al. 2010, Chapin et al. 1997, Dobrzynska et al. 2011, Higuchi et al. 2003, Hoshino et al. 2005, McKee et al. 2006, van den Driesche et al. 2012, Wine et al. 1997). Similarly, immature rats are more vulnerable to the effects of phthalates than mature rats (Dostal et al. 1988, Kondo et al. 2006, Sjoberg et al. 1986). The effect of exposure at different developmental stages in rabbits also highlights the importance of attention to critical windows of exposure (Higuchi et al. 2003). In this study, 400 mg/kg/day DBP was administered in utero (GD15–29), from postnatal week 4–12 and for 12 weeks in 6–8 months old rabbits. The total number of sperm per ejaculate and the volume of the ejaculate were decreased by in utero exposure but unchanged by exposure later in life (Higuchi et al. 2003). These results indicate that the developing fetus and possibly infants and children are more vulnerable to the adverse effects of phthalate exposure than adults. Transgenerational effects In two-generation studies, the first generation (F0) is dosed beginning in adulthood, the second is exposed in utero, and throughout life and the offspring of the second generation are examined as pups. These studies note greater effects, particularly in developmental outcomes, in the second generation (F1) compared to those of the first generation (Blystone et al. 2010, Chapin et al. 1997, Dobrzynska et al. 2011, Hoshino et al. 2005, Wine et al. 1997). However, there may also be an increase in effects between the second generation and their offspring (F2) in animal studies. Therefore, the effects of phthalates in one generation may be transmitted to the next in addition to the effects of in utero exposure. Indeed, transgenerational effects on testicular germ cell organization and spermatogonal stem cell function was demonstrated via in utero exposure to DEHP (Doyle et al. 2013). Furthermore, the possibility of epigenetic changes has been suggested by reports of increased methylation of DNA and increased expression of DNA methyltransferases (Wu et al. 2010a, 2010b). However, identifying epigenetic effects is difficult as the association must be made across four generations. The first through third generations are directly exposed by the initial treatment of the pregnant animals (Jirtle and Skinner 2007, Skinner 2008). During exposure of the first generation pregnant female, the

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second generation is exposed as embryos and the third generation is exposed as germ cells (Jirtle and Skinner 2007, Skinner 2008). Similarly, in humans, the primordial follicles form in utero, meaning that a pregnant woman’s exposure directly impacts both her child and her grandchildren (Sarraj and Drummond 2012). In order for a true epigenetic effect to be demonstrated, epigenetic markers in the fourth generation must be correlated with exposure in the first generation (Perera and Herbstman 2011). A recent study in CD-1 mice provides evidence for a true transgenerational effect of phthalate exposure. Abnormal histology of the seminiferous tubules was reported in the F1, F2, F3, and F4 generations after treatment with 500 mg/kg/day DEHP GD7.5–14.5 in the F0 generation (Doyle et al. 2013). Decreased sperm counts were also reported in all four generations (Doyle et al. 2013). Additionally, a delay in preputial separation was noted in the F1 generation as well as in the F2 and F3 generations descending from exposed males (Doyle et al. 2013). Further long-term studies examining multiple generations are needed to corroborate this phenomenon and clarify any changes in the NOAEL and LOAEL between generations.

Conclusions In conclusion, although the literature has been expanded with many new epidemiological studies in recent years, there is insufficient evidence to determine causal associations with hypospadias or cryptorchidism. The current evidence also suggests no causal association between phthalate exposure and changes in the timing of puberty onset. Furthermore, associations found with AGD, and concentrations of reproductive hormones are weak due to conflicting results, although the evidence for reduced testosterone is more persuasive. There exists greater weight of evidence in terms of the consistency of the epidemiologic literature for a causal association between phthalate exposure and semen quality, although the clinical relevance remains to be determined. We found an expanded and robust animal literature for DEHP, DiBP, and DBP which supports the biological plausibility for adverse effects of phthalates on development of the male reproductive tract and suggests that phthalate-induced Leydig cell dysfunction is central to the documented suppression of steroidogenic enzymes, circulating and intratesticular testosterone concentrations, and semen quality. Several genomic and proteomic studies have identified changes in genes important to cholesterol transport and steroidogenesis as potentially central pathways underlying the toxic effects of phthalates. Furthermore, exposure to high concentrations of phthalates are associated with morphological changes of the testis including loss of Sertoli cells, disruption of seminiferous tubules, and formation of multinucleated germ cells; all potentially linked to increased apoptosis of germ cells and decreased semen quality. While the majority of the documented effects occur at higher concentrations, decreased semen quality appears within an order of magnitude of human exposure. Recently, a high concentration of MBP was detected in the urine of an adult male exposed through the use of a medication delivered in a slow-release capsule (Hauser et al. 2004). This case highlights the concern about potential effects on male reproductive health in highly exposed populations. These concerns

are further elevated by evidence of effects across generations and the potential for additive effects of phthalates in chemical mixtures. However, we note that throughout the animal literature mice were more resistant to the adverse effects of phthalates than rats suggesting differences in species sensitivity to phthalates. Moreover, in xenografts of rat, mouse, and human testes multinucleated germ cells were found but attenuated steroidogenesis was only found in xenografts of rat testes whereas the human xenografts were resistant (Heger et al. 2012). In addition, nonhuman primates (i.e., marmosets) also were much more resistant to the adverse effects of phthalate exposure possibly owing to differences in PPAR activation across mammalian species. Taken together, we suggest that while the rat is the most sensitive animal model, differences in PPAR expression and activation limit the relevance of these data for human health risk assessment. Therefore, clarification of the molecular mechanisms underlying the toxic effects of phthalates and assessment of the transgenerational effects of phthalates remain a high priority for further research.

Acknowledgments The authors gratefully acknowledge the administrative support of Ms. Margaret Talbot during the preparation of this manuscript. Declaration of interest The employment affiliation of the authors is shown on the cover page. Funding has been provided by Ferring Pharmaceutical to WGF for the qualification of phthalate exposure in patients prescribed medications containing phthalates in slow-release capsules. No external funding support was provided for the preparation of this manuscript. The manuscript was prepared during the normal course of the authors’ employment. The authors have not participated in any legal or regulatory proceedings related to the subject matter of the manuscript. The authors have sole responsibility for the preparation of the manuscript and the views and opinions expressed in the manuscript.

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Notice of correction The version of this article published online ahead of print on 06 June 2014 contained two minor typographical errors. These have been corrected for this version.

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Reproductive and developmental effects of phthalate diesters in males.

Phthalate diesters are a diverse group of chemicals used to make plastics flexible and are found in personal care products, medical equipment, and med...
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