Military personnel in the 1990–1991 Gulf War were exposed to combustion products and fuels from more than 600 burning oil wells in Kuwait, from fueling and exhaust from military vehicles and aircrafts, and from heaters in poorly ventilated tents. Post-9/11 military personnel were potentially also exposed to combustion products from fuels, exhaust, and heaters, but they were also exposed to smoke and emissions from the numerous burn pits used to dispose of waste in Iraq and Afghanistan. Gulf War and Post-9/11 veterans deployed to southwest Asia, including Iraq and Afghanistan, were often exposed to heavily dust-laden environments that contained hazardous airborne particles.
This chapter examines the reproductive, developmental, and generational health effects associated with combustion products—particulate matter (PM), polycyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/PCDFs), and several studies that looked specifically at exhaust—and fuels. Previous Gulf War and Health reports, authoritative reviews, and new literature are summarized below.
Several National Academies of Sciences, Engineering, and Medicine’s reports have examined the possible contribution of air pollution to adverse health effects among U.S. military personnel serving in the Middle East. In 2011 a report on the hazards of burn pit emissions identified PM, volatile organic compounds (VOCs, such as benzene and toluene), PAHs, and PCDD/PCDFs as the primary exposures of concerns at Joint Base Balad (JBB) in Iraq (IOM, 2011). The report concluded that
the greatest pollution concern at JBB may be the mixture of regional background and local sources—other than the burn pit—that contribute to high PM. This PM, which was characterized in a different study and at different locations at JBB, consists of substantial amounts of windblown dust combined with elemental carbon and metals that arise from transportation and industrial activities. On the basis of the
high concentrations in the previous studies of potentially toxic constituents of ambient PM, the air pollution literature that focused on PM and gaseous pollutant co-exposures was considered relevant to the potential morbidity of military personnel at JBB and at other sites in the Middle East. (IOM, 2011, p. 4)
Based on that report, the Volume 11 committee focused on several specific pollutant categories. The following section explores the literature pertaining to the reproductive, developmental, and transgenerational effects of PM, PAHs, and PCDDs/PCDFs. Given the overlap between VOCs and solvents, relevant VOCs are discussed in the chapter on solvents.
The Volume 3 committee reviewed the literature on combustion products through 2004. It approached its task by identifying the most common components in combustion emissions, including PM, sulfur dioxide (SO2), ozone (O3), nitrogen oxide (NO), nitrogen dioxide (NO2), carbon monoxide (CO), carbon dioxide (CO2), hydrogen sulfide (H2S), and VOCs and then assessing the possible health effects associated with each of these agents. The Volume 3 committee considered preterm birth, low birth weight, very low birth weight, intrauterine growth retardation, birth defects, and childhood cancers as relevant outcomes. These outcomes were assessed in epidemiologic studies of ambient air pollution—frequently conducted in or near heavily polluted cities—or in case-control studies based on residence or parental occupation. Generally, these studies were limited by a lack of control for other risk factors, such as smoking, and the use of overly broad exposure classifications. The Volume 3 committee concluded that there was:
- inadequate/insufficient evidence of an association between parental combustion product exposure and any childhood cancers or any birth defects,
- limited/suggestive evidence of an association between combustion product exposure during pregnancy and preterm birth,
- limited/suggestive evidence of an association between combustion product exposure during pregnancy and low birth weight or intrauterine growth retardation, and
- insufficient/inadequate evidence of an association at any specific time during pregnancy (e.g., first trimester) and preterm birth, low birth weight, or intrauterine growth retardation (IOM, 2005).
The 2011 IOM report Long-Term Health Consequences of Exposure to Burn Pits in Iraq and Afghanistan examined the epidemiologic literature on firefighters and people living or working near incinerators and health endpoints identified by authoritative reviews (e.g., the Environmental Protection Agency [EPA], the Agency for Toxic Substances and Disease Registry [ATSDR], the National Institute for Occupational Safety and Health, and the Hazardous Substances Database) of the 56 chemicals that were measured above the detection limit in burn pit environmental sampling data. Because data on the health of veterans or active-duty service members who were exposed to burn pits were scant, the committee reviewed studies of firefighters and incinerator workers because of their relatively similar exposure to combustion products. That committee found that there were known reproductive and developmental toxicants among the burn pit chemicals and that they should be further studied. The literature at the time included no “key” studies on reproductive or developmental outcomes, and “supporting” studies presented mixed results or showed no elevated risks. Thus, that committee concluded that there was “inadequate/insufficient evidence to determine whether an association exists between parental exposure to combustion products and adverse effects in their children and other adverse reproductive and perinatal outcomes in veterans, firefighters, or people living/working near incinerators” (IOM, 2011, p. 88).
Particulate matter is a general term used to describe a large group of chemically diverse substances that exist as particles in ambient air (Spektor, 1998). The particles are derived from both natural and anthropogenic sources; they may form in the atmosphere from a variety of sources or be emitted to the atmosphere with gaseous emissions. Coarse PM (10 μm or less in diameter; PM10) originates from sources such as mining, construction operations, and the resuspension of soil on roads and streets. Fine PM (less than 2.5 μm; PM2.5) and ultrafine PM (less than 0.1 μm) are found in emissions from the combustion of motor-vehicle fuel; smelters and steel mills; the combustion of wood, oil, and coal; and the atmospheric alteration of SO2, organics, and NO (NRC, 2010). The size of the particles is one of the major determinants of its toxicity. Only inhaled PM10 and smaller particles are deposited in the lower respiratory tract (i.e., below the vocal cords). Toxicity concerns are greatest for fine PM and ultrafine PM that may be deposited in smaller airways and alveoli (IOM, 2005). Ultrafine particles may even pass through the alveolar-capillary membrane into the circulatory system.
The daily exposures of service members to windblown dust and frequent dust storms increased their concerns about the possible health effects that might result from these exposures. In response to those concerns, the Department of Defense (DoD) designed and implemented the Enhanced Particulate Matter Surveillance Program to characterize and quantify the PM at 15 sites in the Central Command Area of Operations in the Middle East, beginning in Afghanistan in 2001 and in Iraq in 2003. PM sources in the Middle East included dust storms (which would carry the particles considerable distances), agricultural activities, emissions from vehicles and burn pits, battery-processing facilities, cooking stoves and fires, flaring at natural gas and oil extraction facilities, and fertilizer plants (NRC, 2010). DoD monitoring at JBB in Iraq in 2007 and 2009 found that PM10 levels ranged from 2 μg/m3 to 9,576 μg/m3 (with averages of 126 μg/m3 in 2007 and 709 μg/m3 in 2009) (IOM, 2011). Another study of three military sites in Kuwait in 2004–2005 found that the average PM10 concentrations fluctuated between 66 and 93 μg/m3 (Brown et al., 2008). In its review of the DoD Enhanced Particulate Matter Surveillance Program, a National Research Council committee determined that ambient PM10 concentrations in most areas of the Middle East exceeded the World Health Organization’s annual average air quality guideline of 20 μg/m3 (NRC, 2010). As a point of reference, PM exposure in the United States is subject to great variability as well, with some areas such as southern California, southern Arizona, and central New Mexico reporting average annual PM10 concentrations >50 μg/m3 (EPA, 2009).
In 2009, EPA issued an Integrated Science Assessment for Particulate Matter that described the human health effects associated with ambient PM through 2009. EPA concluded that the “epidemiologic and toxicologic evidence is suggestive of a causal relationship between long-term exposures to PM2.5 and reproductive and developmental outcomes” (p. 2-13). Epidemiologic data showed effects on birth weight, infant mortality, and fetal growth, with less consistent evidence for preterm birth, growth restriction, birth defects, and decreased sperm quality. The available toxicologic data also supported those associations but added little to the mechanistic understanding. EPA also noted that some PM studies had shown epigenetic effects (i.e., changes in DNA methylation), but their relevance to human health was unclear (EPA, 2009).
Although the data presented in this chapter focus on PM, it is just one of many constituents of air pollution determined by the EPA to cause adverse health effects, a so-called criteria pollutant. Other EPA criteria pollutants include O3, CO, NOX, and SO2 (EPA, 2018). Because air pollution is a mixture whose composition varies depending on the source, with emissions from combustion sources generally regarded as the most toxic, it is difficult to distinguish the contributions of specific pollutants from the effects of the overall mixture. Furthermore, PM concentrations are often correlated with other pollutants
emitted during combustion, and so the committee recognizes that PM toxicity does not exclude effects of co-pollutants.
Like the Volume 3 committee, the Volume 11 committee did not examine the chemical and physical composition of particles because of the wide variation in particle components across geographic locations and sources, including differences in the relative amounts of elemental and organic carbon, sulfates, nitrates, pollen, microbial contaminants, and metals. Furthermore, the bioavailability and toxicity of such chemicals is affected by adsorption to the particle so that the toxicity of a chemical alone does not represent its toxicity when adsorbed on the particle (IOM, 2005). Rather, the Volume 11 committee focuses on studies of the toxicity of PM, not studies of the individual chemicals that might be adsorbed onto or incorporated into the PM.
The literature on the reproductive and developmental effects of PM is varied and extensive. The committee’s search identified 340 relevant publications of reproductive and developmental effects associated with PM published in 2004 or later (combustion products are reviewed in Volume 3 through 2004). Because EPA conducted an extensive review of the literature on the health effects of PM through 2009, the Volume 11 committee limited its review to studies published in 2009 or later, which totaled about 300 publications. No specific studies were identified on the reproductive or developmental effects of PM exposure in veterans. Given the volume of data, the committee paid special attention to evidence synthesized in systematic reviews and meta-analyses as well as to studies on preconception or first-trimester exposures because they were considered to be most relevant to potential veteran exposures.
Several recent publications have explored possible associations between PM and aspects of reproductive health or fertility in men and women—including endometriosis, fertility and fecundity, offspring sex ratios, and impacts on sperm. With the exception of fertility studies that rely on self-reports of contraception use, most of the studies on reproductive effects in women rely on outcomes documented in birth certificates, hospital discharge records, and registries.
Reproductive Effects in Men
A systematic review by Lafuente et al. (2016) evaluated 17 studies on the impacts of air pollution on sperm quality. Although the studies examined a range of windows of exposure before semen collection, all included the 90 days before sampling. The authors also noted a wide variability in the degree to which confounders such as age, smoking, body mass index (BMI), and socioeconomic status (SES) were controlled for. Thirteen studies measured air pollution exposure, six used biomarkers of air pollution exposure, and two did both. The authors concluded that PM2.5 was not associated with sperm DNA fragmentation, decreased sperm count, or alterations in sperm morphology.
Four cross-sectional studies examined the effects of PM exposure within 90 days of semen collection with mixed results, as discussed below (Hansen et al., 2010; Jurewicz et al., 2015; Wu et al., 2017; Zhou et al., 2014). Most of the studies had negative findings.
Significant detrimental effects on sperm count and concentration were reported by Wu et al. (2017) in association with PM2.5 exposure 70–90 days before sample collection and with the highest quintile of exposure over the whole exposure period of 0 to 90 days (sperm concentration β= –0.42, 95% confidence interval [CI] –0.67– –0.16; sperm count β= –0.40, 95% CI –0.66– –0.14; adjusted p-value for trend=0.018 for both). These data showed that there were no effects on sperm motility. The authors also noted no effect of PM10 on sperm centration, count, or morphology. Hansen et al. (2010) found that PM2.5
had no effect on sperm morphology, cytoplasmic droplets, DNA fragmentation, or immature sperm in an occupational cohort of 228 fertile men. Zhou et al. (2014) reported a positive association between PM10 and sperm concentration (β=0.08, 95% CI 0.02–0.11) and a negative association with sperm morphology (β= –0.21, 95% CI –0.80– –0.003) and some, but not all, measures of motility (motile sperm curve linear velocity β= –0.16, 95% CI –0.80 to 0.68, and motile sperm straight line velocity β= –0.21, 95% CI –0.82– –0.016); PM10 was not related to semen volume, progressive motility, total motility, or linearity.
Jurewicz et al. (2015) examined the relationship between PM and the percentage of sperm with chromosome disomy (X, Y, 13, 18, and 21). The authors observed significant relationships between PM2.5 and disomy Y (YY18 β=0.68, 95% CI 0.55–0.85), sex chromosome disomy (β=0.78, 95% CI 0.59–0.99), and chromosome 21 disomy (%2121 β=0.78, 95% CI 0.62–0.97), with no relationship to XY18, XX18, %1818, %1313, and total disomy. PM10 was also related to chromosome 21 disomy (%2121 β=0.58, 95% CI 0.46–0.72), but not to any of the other chromosome disomies.
Reproductive Effects in Women
One systematic review examined the human and animal literature published through early 2014 on fertility in association with PM (Frutos et al., 2015). Ten studies on live births, miscarriages, pregnancy, implantation, and embryo quality were assessed. The authors identified PM10 as having a detrimental effect on live births, implantation rates, and miscarriage, while PM2.5 was reported to adversely affect pregnancy rates. The authors also noted that impacts were greatest among sub-fertile women undergoing assisted reproduction.
Three cohort studies examined the association between PM exposure and the ability to conceive. The Nurses’ Health Study II, a large prospective cohort study of female nurses, examined associations between fertility (defined as attempted conception for ≥12 months without success) over 10 years and PM2.5, PM10–2.5, and PM10 levels based on residential address. Those investigators reported elevated hazard ratios for all three sizes of PM expressed as 2-year, 4-year, or cumulative averages, but all were nonsignificant after adjustment for confounders (Mahalingaiah et al., 2016). In a study that examined a cohort of 1,916 births in Teplice, Czech Republic, fecundability (the probability of pregnancy among noncontracepting couples each month) was found to be significantly related to local PM2.5 levels. Exposure in the previous 30 days was significantly related to decreased fecundability (fecundity ratio=0.86, 95% CI 0.77–0.97 per 10 μg/m3 increase). Fecundability was slightly reduced with PM2.5 exposure in the first month of unprotected intercourse, but not significantly. Both average PM2.5 in the previous 30-day period and the 2-month average exposure to PM2.5 were significantly related to decreased fecundability (adjusted fecundability ratios per 10 μg/m3=0.86, [95% CI 0.77–0.97] and 0.78, [95% CI 0.65–0.94], respectively) (Slama et al., 2013). Conversely, Nobles et al. (2018) reported on fecundability, over a year compared to modeled PM 5 days before and 10 days after ovulation in 501 couples living in Michigan and Texas but did not find evidence of adverse effects of PM2.5 or PM10 on fecundability. Nieuwenhuijsen et al. (2014) compared annual averages of PM2.5, PM10–2.5, and PM10 per interquartile range (IQR) change with birth rates at the census tract level and found a significantly decreased birth rate in association with PM10–2.5 that remained after adjustment for multiple covariates (incidence rate ratio [IRR]=0.88, 95% CI 0.83–0.94; PM2.5 IRR=0.98, 95% CI 0.95–1.02; PM10 IRR=0.99, 95% CI 0.97–1.02).
Endometriosis, which can impair fertility, was investigated in one publication of the Nurses’ Health Study II (Mahalingaiah et al., 2014). No associations between PM2.5, PM10–2.5, or PM10 and confirmed cases of endometriosis were reported in this cohort.
Merklinger-Gruchala et al. (2017) reported on associations between pollutants and changes in menstrual cycle length in 133 Polish women as a possible contributing factor to fertility problems. While
PM10 was not associated with overall cycle length or follicular phase length, mean PM10 was associated with a shortened luteal phase length (β= –0.02 day per 1μg/m3, 95% CI –0.04– –0.00).
Two studies of sex ratios indicated that increased exposure to PM favors female births. In a Chinese birth cohort study that examined concentrations of PM10 in the 13 days before conception versus sex ratios, Lin et al. (2015) found a significantly higher probability of female births with increased PM10 exposure 2 days before conception (excess risk = 0.64%, 95% CI 0.35–0.93 for each 10 μg/m3 increase in PM10), while controlling for confounders such as maternal age, parental education, and weather conditions. One ecologic time-series study found a negative association between the ratio of male to female births and annual average PM10 levels at the district level (β= –0.001, p=0.02), although no potential confounders or other pollutants were included as covariates in the analysis (El Khouri Miraglia et al., 2013).
Adverse Pregnancy Outcomes
The relationship between exposure to PM and adverse pregnancy outcomes has been studied by numerous researchers assessing a variety of maternal and fetal outcomes. The Volume 11 committee identified several studies of pregnancy outcomes that affect the health of the mother and child; these include 2 studies of placental problems, 3 studies on gestational diabetes, and 12 studies on preeclampsia1 and gestational hypertension. Several studies also examined the associations between maternal exposure to PM during pregnancy and effects on the uterus and placenta and also on fetal biomarkers that are thought to represent inflammation at the time of birth.
The Volume 11 committee then focused on the effects of PM on birth outcomes such as preterm birth, low birth weight (less than 2,500 g), and fetal death. In the sections on preterm birth and birth weight, the committee evaluated numerous systematic reviews and meta-analyses. Systematic reviews and meta-analyses were not available for outcomes such as miscarriage, stillbirths (fetal loss after 20 weeks of gestation), and spontaneous abortion (fetal loss before 20 weeks of gestation), and therefore the committee considered individual studies for these events. For the most part, the committee considers studies that include appropriate statistical adjustments for potential confounders and indicates instances when a study may not have accounted for common confounders.
Understanding the effects of prenatal PM on birth outcomes requires an appreciation of several confounding issues, such as SES and maternal health. Erickson et al. (2016) found that 49% of the between-neighborhood differences in birth weight were related to neighborhood SES and the interaction between SES and PM2.5. SES modified the effect of PM2.5 on birth weight in such a way that the most pronounced effect was in neighborhoods with lower SES. Mendola et al. (2016) reported significantly increased risks of preterm birth among mothers with asthma exposed to PM10 in the last 6 weeks of pregnancy but no association between preterm birth and PM10 among women without asthma.
Placental Problems Two studies examined placental problems and prenatal PM exposure. In Japan,
1Preeclampsia is a pregnancy complication related to high blood pressure during pregnancy (gestational hypertension) that can cause damage to organ systems, specifically kidney and liver, and lead to the need for early delivery of the baby and, in extreme cases, seizures (eclampsia) in the mother (Mayo Clinic, 2018d).
Michikawa et al. (2016) found a significantly elevated risk of placenta accreta2 and placenta previa associated with exposure to suspended PM (equivalent to PM7) in the first 4 weeks of pregnancy (placenta previa odds ratio [OR]=1.12, 95% CI 1.01–1.23; placenta accreta OR=1.33, 95% CI 1.07–1.66, per 10 μg/m3), but no associations were found for exposure during gestation weeks 5–12 or in the second trimester. Ibrahimou et al. (2017) examined placental abruption as recorded on Florida birth certificates (884 placental abruptions among 100,495 births) and found an increased risk was associated with PM2.5 exposure throughout pregnancy or during the first trimester. Slightly elevated and significant risks were identified for first-trimester exposure to specific components of PM2.5 (aluminum and elemental carbon).
Gestational Diabetes Gestational diabetes was the focus of three U.S. cohort studies. In a large cohort study of 219,952 births, Robledo et al. (2015) found no association between gestational diabetes and PM2.5 or PM10 in the 3 months before pregnancy or in the first trimester. However, Hu et al. (2015) reported an increased risk of gestational diabetes with PM2.5 exposure in the first trimester (OR=1.11, 95% CI 1.06–1.16), but not in the second trimester, after controlling for other pollutants although not for prepregnancy weight, an important risk factor for gestational diabetes. Investigating PM2.5 exposure in the second trimester, Fleisch et al. (2014) found a significant increase in impaired glucose tolerance when comparing the highest and lowest quartiles of exposure (OR=2.63, 95% CI 1.15–6.01) but no association with gestational diabetes as a clinical diagnosis.
Preeclampsia and Gestational Hypertension One systematic review and meta-analysis of literature published between 2009 and 2013 examined the association between PM and preeclampsia and gestational hypertension, with seven studies on exposure and five on PM10 exposure (Pedersen et al., 2014). Pooled estimates showed significant increases in all hypertensive disorders of pregnancy, combined, with each 5 μg/m3 increase of PM2.5 (OR=1.57, 95% CI 1.26–1.96) and each 10 μg/m3 increase of PM10 (OR=1.13, 95% CI 1.02–1.26).
Among the studies discussed below, one found increased risks of preeclampsia with PM10 exposure (Lee et al., 2013), while four did not (Dadvand et al., 2013; Mobasher et al., 2013; van den Hooven et al., 2011; Wu et al., 2011). Four studies of PM2.5 found positive associations with preeclampsia (Dadvand et al., 2013; Lee et al., 2013; Mobasher et al., 2013; Wu et al., 2011), and two studies did not (Rudra et al., 2011; Savitz et al., 2015). One study of preconception exposure by Rudra et al. (2011) found no relationship between PM2.5 and preeclampsia.
Lee et al. (2013) found no significant relationship between preeclampsia and first-trimester PM10 or PM2.5 exposure. However, among mothers who did not smoke during pregnancy, the risks of gestational hypertension were elevated with PM exposure (PM10 OR=1.12, 95% CI 1.01–1.25; PM2.5 OR=1.11, 95% CI 0.99–1.24, per IQR change); this elevation was not seen among mothers who smoked during pregnancy.
In a cohort study in Spain of 8,398 pregnancies (including 103 cases of preeclampsia), Dadvand et al. (2013) found no association between first- or second-trimester PM exposures and preeclampsia or between third-trimester PM10 exposures and preeclampsia. However, IQR increases of PM2.5 in the third trimester (OR=1.51, 95% CI 1.13–2.01) and over the entire pregnancy (OR=1.32, 95% CI 1.02–1.71) were both related to preeclampsia, as were increases among late-onset (OR=1.42, 95% CI 1.01–2.00) but not early-onset cases.
2Placenta accreta is a serious pregnancy condition that occurs when the placenta grows too deeply into the uterine wall. While the placenta usually detaches from the uterine wall after childbirth, with placenta accreta it remains attached and can cause severe blood loss after delivery. Placenta previa is when the placenta partially or totally covers the mother’s cervix, requiring a cesarean section to deliver the baby. It can cause severe bleeding during pregnancy and delivery. Placental abruption is when the placenta separates from the the uterus before delivery, an action that can decrease the baby’s oxygen and nutrient supply and cause heavy bleeding in the mother (Mayo Clinic, 2018a,b,c).
Among 81,186 births in Los Angeles and Orange counties, Wu et al. (2011) found no elevated risks of preeclampsia with PM2.5 or PM10 exposure throughout the entire pregnancy using regression models and PM measured at air quality monitoring stations. However, when an air dispersion model (CALINE 4 model) was used for PM2.5 exposure, significantly elevated risks were found (LA County OR=1.08, 95% CI 1.02–1.15; Orange County OR=1.10, 95% CI 1.04–1.17). Another study in LA County used a case-control design (136 cases, 162 controls) to examine the roles of pollutant exposure and BMI on preeclampsia and found a significant interaction (p=0.02). Results showed an increased risk of preeclampsia with PM2.5 exposure in the first trimester among women with BMI <30 (OR=8.63, 95% CI 3.10–24.1) but no significant increase among obese women (BMI >30) or related to PM10 exposure (Mobasher et al., 2013). A cohort study of pregnant women in the Netherlands showed no relationship between PM10 exposure in any trimester and preeclampsia (van den Hooven et al., 2011).
Rudra et al. (2011) examined PM2.5 exposure from prepregnancy through the last month of pregnancy in 3,509 Washington State births and found no association between PM2.5 and preeclampsia. Savitz et al. (2015) also examined PM2.5 exposure in the first and second trimesters and found no association with gestational hypertension or with mild or severe preeclampsia among 268,601 New York City births.
Several studies discussed below reported a positive association between maternal blood pressure or gestational hypertension and exposure to PM (Lee et al., 2012; van den Hooven et al., 2011; Vinikoor-Imler et al., 2012; Xu et al., 2014); however, one study found no such effects (Hampel et al., 2011).
Xu et al. (2014) reported on the risk of hypertensive disorders of pregnancy (including gestational hypertension, preeclampsia, and eclampsia) among 22,041 pregnancies in Florida. Although there was no association between hypertensive disorders of pregnancy and PM2.5 exposure in the first trimester, elevated risks were reported for exposure in the second trimester (OR=1.28, 95% CI 1.13–1.46) and across the entire pregnancy (OR=1.24, 95% CI 1.08–1.43).
In a cohort of 1,684 pregnant women in Pennsylvania, investigators compared blood pressure measurements taken in the first 20 weeks of gestation with those measured late in pregnancy and found increases that were associated with first-trimester PM2.5 exposure (systolic β=0.84 mmHg, 95% CI −0.33–2.00, and diastolic β=0.60 mmHg, 95% CI –0.27–1.47, per IQR) and PM10 exposure (systolic β=1.88 mmHg, 95% CI 0.84–2.93, and diastolic β=0.63 mmHg, 95% CI –0.50–1.76). No associations were seen with exposures during the second or third trimesters (Lee et al., 2012).
Hampel et al. (2011) measured blood pressure longitudinally in a cohort of 1,500 pregnant French women and assessed associations with PM10 exposures in the preceding 7 days. No significant changes for systolic and diastolic pressure changes were associated with exposure on each day in the week before blood pressure was measured. However, a significant decrease in systolic pressure was observed when PM10 exposure was averaged across the 7 days before measurement (–0.3 mmHg, 95% CI –0.6–0.0).
Among 7,006 Dutch women, van den Hooven et al. (2011) reported that 10 μg/m3 increases in PM10 concentrations were associated with a 1.11 mmHg increase in systolic blood pressure in the second trimester (95% CI 0.43–1.79) and a 2.11 mmHg increase in the third trimester (95% CI 1.34–2.89), but there were no associations with first-trimester exposure or diastolic blood pressure. The authors reported a significantly increased risk of pregnancy-related hypertension (OR=1.72, 95% CI 1.12–2.63) for each 10 μg/m3 increase in PM10.
Vinikoor-Imler et al. (2012) found increased risks of gestational hypertension with IQR increases in PM2.5 and PM10 averaged to represent the entire pregnancy in 222,775 North Carolina births (PM2.5 RR=1.11, 95% CI 1.08–1.15, and PM10 RR=1.07, 95% CI 1.04–1.11). The associations remained significant after adjustments for neighborhood poverty. However, this study did not control for maternal weight gain, prepregnancy weight, or hypertension.
Prenatal exposure to PM2.5 has also been associated with an increased systolic blood pressure in the child at birth. Each IQR increase in PM2.5 estimated for the 2–90 days before birth was associated with a 1.0 mmHg increase in systolic blood pressure (95% CI 0.1–1.8) (van Rossem et al., 2015).
Most of the studies cited above controlled for prepregnancy weight, BMI, or maternal weight gain, all of which are important risk factors for gestational hypertension and preeclampsia (Dadvand et al., 2013; Hampel et al., 2011; Lee et al., 2012; Mobasher et al., 2013; Rudra et al., 2011; Savitz et al., 2015; van den Hooven et al., 2011). Three studies did not adjust for maternal BMI (Lee et al., 2013; Vinikoor-Imler et al., 2012; Wu et al., 2011), thus limiting their usefulness.
Biomarkers of Oxidative Stress and Inflammation at Birth Thirteen studies identified by the committee investigated various measures and markers of stress and inflammation in the uterus, placenta, and cord blood in relation to prenatal PM exposure. Toxicologic data suggests that adverse health effects in offspring may be mediated by a greater susceptibility of the mother to the oxidative and inflammatory effects of PM during pregnancy; for example, inflammatory or immune response in pregnant mice has been associated with health effects in the offspring (EPA, 2009).
Effects of PM on the uterus and placenta have been studied in four cohorts. In a large cohort of 5,059 mother–child pairs in Boston, increased risks of intrauterine inflammation (intrapartum fever and placental pathology) were reported with the highest quartile of PM2.5 exposure compared with the lowest quartile in the 3 months before conception (OR=1.52, 95% CI 1.22–1.89), first trimester (OR=1.93, 95% CI 1.55–2.40), second trimester (OR=1.67, 95% CI 1.35–2.08), and third trimester (OR=1.53, 95% CI 1.24–1.90) (Nachman et al., 2016). Placental 3-nitrotyrosine (3-NTp; a marker of oxidative stress) levels increased with each IQR increase in PM2.5 (29%, 95% CI 4.9–58.6) and black carbon (23.6%, 95% CI 4.4–46.4) in the first trimester in a birth cohort in Belgium (N=90) (Saenen et al., 2016). A study of 3,614 Italian women did not find an association between PM10 and placental weight (Giovanni et al., 2018).
To investigate whether placental oxidative stress could affect fetal growth and maternal hypertensive complications, van den Hooven et al. (2012b) measured growth factors that are important for placental growth in maternal and cord blood as well as placental vascular resistance and placental weight in 7,801 pregnant women in the Netherlands (known as the Generation R Study). Throughout pregnancy, PM10 exposure (10 μg/m3) was associated with decreased maternal growth factor levels as measured by soluble fms-like tyrosine kinase 1 (sFlt-1= –4.5%, 95% CI –8.4– –0.5) and changes in fetal growth factors (sFlt-1= 35.8%, 95% CI 25.6–45.9; and placental growth factor= –16.3%, 95% CI –21.9– –10.7%). Although PM10 was associated with the umbilical artery pulsatility index in the second trimester (difference= –0.10, 95% CI –0.17– –0.03), it was not associated with the uterine artery pulsatility index, uterine artery notching, or placental weight (van den Hooven et al., 2012b). These findings indicate that PM may affect placental and fetal growth.
Several studies documented changes in immune markers in cord blood associated with PM exposure. Ashley-Martin et al. (2016) reported an association between the highest quartile of exposure to PM2.5 in the first trimester and immunoglobulin E (IgE) in cord blood among 582 female births (OR=2.37, 95% CI 1.23–4.58) in Canada, although there was no association among 671 male births or with other markers (IL-33 and TSLP). The prevalence ratios of elevated cord serum IgE for each μg/m3 increase in PM2.5 indicated decreased risks in the first month of pregnancy and elevated risks in months 4 through 7 in a study of 459 births in the Czech Republic; the results were modified by maternal IgE levels and cigarette smoke (Herr et al., 2011).
Levels of C-reactive protein, a marker of inflammation, have been studied in association with PM exposure. Lee et al. (2011) found an increased risk of having elevated C-reactive protein (≥8 ng/ml) in maternal blood collected before the 22nd week of gestation and PM in the month before sample collection (PM10 9.2 μg/m3 increase OR=1.41, 95% CI 0.99–2.00; PM2.5 4.6 μg/m3 increase OR=1.47, 95% CI
1.05–2.06) in a cohort of nonsmoking pregnant women in Pennsylvania. In the Netherlands, the highest quartile of PM10 exposure in the 2 weeks before maternal blood sampling had increased risks of elevated C-reactive protein levels (≥8 ng/ml) among pregnant women (OR=1.32, 95% CI 1.08–1.61), but there was no association at 1 and 4 weeks before sampling. This study also found that high PM10 throughout the pregnancy was related to elevated fetal C-reactive protein at delivery (>1 mg/L) (OR=2.18, 95% CI 1.08–4.38) (van den Hooven et al., 2012a).
Changes in newborn cord blood lymphocyte subpopulations associated with PM10 exposure before and during pregnancy were reported among 370 women giving birth at a hospital in France (Baiz et al., 2011). Preconception exposure to PM10 (per 10 μg/m3 increase) was associated with increases in CD8+ cells (β=1.52, 95% CI 0.05–2.99) and decreases in CD4+CD25+ cells (β= –0.82, 95% CI –1.45– –0.19). First-trimester exposure was associated with decreases in CD3+ cells (β= –3.10, 95% CI –6.02– –0.18) and increases in natural killer (NK) cells (β=0.22, 95% CI 0.008–0.43), and exposure across all three trimesters was associated with decreases in CD4+CD5+ cells (1st β= –0.71, 95% CI –1.37– –0.05; 2nd β= –0.88, 95% CI –1.30– –0.17; 3rd β= –0.59, 95% CI –1.15– –0.04) (Baiz et al., 2011). Another study of 1,397 births in the Czech Republic also showed changes in lymphocyte subpopulations with PM exposure in early and late pregnancy. First-trimester PM2.5 exposure was associated with increases in CD3+ and CD4+ fractions and decreases in CD19+ fraction and NK cells; whereas, third-trimester exposure was associated with decreased CD3+ and CD4+ fractions and increased CD19+ fraction and NK cells (Herr et al., 2010).
In the cord blood of 197 newborns from the Belgian ENVIRONAGE (Environmental influence ON early AGEing) study, Martens et al. (2017a) measured oxylipins as a marker of the oxidation of fatty lipids. The investigators found that oxylipins, representing the lipoxygenase metabolic pathway for inflammation, were significantly associated with PM2.5 in the second trimesters for both the 5-LOX and the 12/15-LOX pathways, but only the 5-LOX pathway was significantly associated with PM2.5 exposure throughout the pregnancy. Reductions in telomere length, a marker of oxidative stress and inflammation, were studied in 730 newborns from the same cohort study (Martens et al., 2017b). Each 5 μg/m3 increase in PM2.5 in the second trimester was significantly associated with shorter telomere length in cord-blood leukocytes (-9.4%, 95% CI –13.1– –5.6%) and placental tissue (-0.1%, 95% CI –11.6– –2.4%).
Fetal Growth and Preterm Birth There is an extensive literature surrounding birth outcomes (preterm birth, low birth weight, and fetal death) associated with PM, often with conflicting results. The committee’s search identified 96 studies on preterm birth and birth weight (70 studies on birth weight, small for gestational age, and intrauterine growth retardation, and 44 studies on preterm birth) published since 2009. To best describe this volume of literature, the committee examined relevant systematic reviews and meta-analyses and selected publications that described special issues of interest such as confounders and preconception exposure. The systematic reviews and meta-analyses reviewed here are summarized in Table 6-1. Fetal death, including stillbirths and miscarriages, associated with PM exposure has been the subject of relatively few extensive reviews and meta-analyses, and those studies are reviewed individually.
Eight systematic reviews and meta-analyses have assessed the consistency of the literature on adverse birth outcomes associated with PM exposure. These assessments cover the literature on preterm birth and birth weight through 2015 and small for gestational age through 2014. All these reviews noted difficulties in assessing the evidence due to the various studies’ heterogeneity in study design, exposure assessment methods, windows of exposure assessment (each trimester, only specific trimesters or time periods, or the entire duration of the pregnancy), the ability to adjust for confounders (e.g., SES, smoking, and maternal nutrition), and the level of increase reported in analyses (IQR=1, 10, 50 or 100 μg/m3; low versus high exposure; distance from an air quality monitoring station, etc.). Furthermore, there are
TABLE 6-1 Meta-Analyses and Systematic Reviews of PM and Preterm Birth, Birth Weight/Low Birth Weight, and Small for Gestational Age
|Study||Approach||Exposure/Pooled Result or Positive Studies/Total Studies||Comments|
|Lai et al. (2013)||SR* of Chinese population-based studies, though 2012.||PM10: 1/1||Data on birth outcomes are insufficient for meta-analyses.|
|Shah et al. (2011)||SR of women exposed to pollutants, through 2009.||PM2.5: 1/1
|PM2.5: 2/4||PM2.5: 3/4||Heterogeneity of associations across studies. PM2.5 associated with LBW, PTB, and SGA.|
|Bosetti et al. (2010)||SR of maternal exposure to pollutants, through 2009.||PM2.5: 3/4
|No summary estimates provided due to heterogeneity. Does not provide convincing evidence of association with risk of PTB, LBW, or SGA.|
|Jacobs et al. (2017)||SR of Chinese studies, through 2015.||PM10: 5/10||PM10: 4/7||Associations between birth weight and PM10 are not as consistent as SO2.|
|Sun et al. (2015)||MA of prenatal exposure to PM2.5, through 2014.||13 studies OR=1.13 (95% CI 1.03–1.24)||Clear association between PM2.5 and PTB but with significant heterogeneity. Estimates by trimester were not significant.|
|Sun et al. (2016)||MA of prenatal exposure to PM2.5, through 2015.||32 studies birth weight β= –15.9 (95% CI –26.8– –5.0) LBW OR=1.09 (95% CI 1.03–1.15)||Significant heterogeneity. PM2.5 significantly associated with decreased birth weight and risk of LBW with second and third trimesters both significant.|
TABLE 6-1 Continued
|Study||Approach||Exposure/Pooled Result or Positive Studies/Total Studies||Comments|
|Zhu et al. (2015a)||MA of birth outcomes and PM2.5, through 2014.||8 studies OR=1.10 (95% CI 1.03–1.18)||12 studies birth weight β= –14.58 (95% CI –19.31– –9.86) LBW OR=1.05 (95% CI 1.02–1.07)||6 studies OR=1.15 (95% CI 1.10–1.20)||Significant heterogeneity. Effects by trimester not significant for PTB, second and third trimester exposure significant for birth weight, all trimesters significant for SGA.|
|Sapkota et al. (2012)||MA of birth outcomes and PM2.5 or PM10, through 2009.||6 studies on PM2.5 and 8 studies on PM10.
PM2.5: OR=1.15 (95% CI 1.14–1.16)
PM10: OR=1.02 (95% CI 0.99–1.04)
|4 studies on PM2.5 and 11 studies on PM10.
PM2.5: OR=1.09 (95% CI 0.9–1.32)
PM10: OR=1.02 (95% CI 0.99–1.05)
|Third trimester exposure significantly related to PTB and LBW. Significant heterogeneity for PM10 but not PM2.5.|
|Lamichhane et al. (2015)||MA of birth outcomes and PM2.5, or PM10, through 2015.||7 studies on PM2.5, 3 studies on PM10
PM2.5: OR=1.14 (95% CI 1.06–1.22)
PM10: OR=1.23 (95% CI 1.04–1.41)
|8 studies on PM2.5, 16 studies on PM10
PM2.5: β= –13.88 (95% CI –15.7––12.06)
PM10: β= –6.50 (95% CI –10.94––2.5)
|Analyses stratified by adjustment for smoking. Pooled estimates of studies adjusted for smoking were much higher than those that did not. Pooled trimester estimates not reported.|
NOTE: CI=confidence interval; LBW=low birth weight; MA=meta-analysis; OR=odds ratio; PM=particulate matter; PTB=preterm birth; SGA=small for gestational age; SO2=sulfur dioxide.
* Systematic reviews (SRs) defined as those that reported databases searched, search terms, inclusion and exclusion criteria, and approach for assessing biases. Positive study=any statistically significant results (p<0.05) at any level of exposure at any time period during pregnancy (entire pregnancy, first, second or third trimesters).
inconsistencies among trimester-specific results. All the meta-analyses express pooled estimates as ORs per 10 μg/m3 increases in PM. These inconsistencies make the interpretation of the results challenging and highlight the need for additional work in this area.
Preterm birth associated with PM2.5 was the subject of two systematic reviews and four meta-analyses. The two reviews reported significantly elevated risks of preterm birth in association with prenatal PM2.5 exposure (Bosetti et al., 2010; Shah et al., 2011). Of the four meta-analyses that reported pooled estimates of preterm birth associated with prenatal exposure to PM2.5 throughout the gestational period, all reported slight but significantly elevated ORs (Lamichhane et al., 2015; Sapkota et al., 2012; Sun et al., 2015; Zhu et al., 2015a), which ranged from 1.10 (95% CI 1.03–1.18), based on eight studies (Zhu, et al., 2015a), to 1.15 (95% CI 1.14–1.16), based on six studies (Sapkota et al., 2012). First-, second-, and third-trimester-specific estimates were less consistent. Sapkota et al. (2012) found positive associations in all three trimesters (ORs=1.04, 95% CI 0.73–1.34; 1.07, 95% CI 1.00–1.15; and 1.15, 95% CI 1.15–1.16, respectively), but only the associations for the second and third trimesters were significant. By contrast, Zhu et al. (2015a) found that none of the trimester-specific estimates were elevated (ORs=0.96, 95% CI 0.77–1.21; 0.90, 95% CI 0.79–1.03; and 0.97, 95% CI 0.89–1.05, respectively), and Sun et al. (2015) reported elevated but not statistically significant estimates (ORs=1.08, 95% CI 0.92–1.26; 1.09, 95% CI 0.82–1.44; and 1.08, 95% CI 0.99–1.17, respectively). Lamichhane et al. (2015) did not report trimester-specific estimates for PM2.5.
Fewer studies have examined the risk of preterm birth in association with prenatal exposure to PM10. Four systematic reviews reported 1 of 1, 1 of 2, 3 of 9, and 5 of 10 studies that found statistically significant elevated risks (Bosetti et al., 2010; Jacobs et al., 2017; Lai et al., 2013; Shah et al., 2011). Such heterogeneity leaves unanswered questions related to the effects of PM on the risk of preterm birth. Two of the meta-analyses calculated pooled estimates for PM10 throughout the gestational period and preterm birth; the first found an OR=1.02 (95% CI 0.99–1.04) based on eight studies (Sapkota et al., 2012), while the second calculated an OR=1.23 (95% CI 1.04–1.41) based on three studies (Lamichhane et al., 2015). Regarding trimester-specific pooled estimates, Sapkota et al. (2012) found significantly increased ORs for the second and third trimesters (OR=1.02, 95% CI 0.97–1.06; OR=1.02 95% CI 1.01–1.03, respectively), whereas Lamichhane et al. (2015) found no significant increase with PM10 exposure in the first trimester (OR=0.98, 95% CI 0.94–1.03), a significant decrease in the second trimester (OR=0.97, 95% CI 0.95–0.99), and significant increase in the third trimester (OR=1.03, 95% CI 1.01–1.05).
Two systematic reviews and four meta-analyses assessed decreases in birth weight and low birth weight in association with prenatal exposure to PM2.5. The systematic reviews identified two (out of four) and two (out of three) studies that reported effects on birth weight in association with PM2.5 (Bosetti et al., 2010; Shah et al., 2011). Three of the meta-analyses calculated pooled estimates of decreases in birth weight associated with 10 μg/m3 changes in PM2.5 exposure throughout gestation with statistically significant results: β= –15.9 (95% CI –26.8– –5.0), based on 32 studies (Sun et al., 2016); β= –14.58 (95% CI –19.31– –9.86), based on 12 studies (Zhu et al., 2015a); and β= –13.88 (95% CI –15.7– –12.06), based on 8 studies (Lamichhane et al., 2015). Of those that reported trimester-specific pooled estimates, Sun et al. (2016) found significant decreases in birth weight for the second and third trimesters (βs= –12.6 and –10.0, respectively) as did Zhu et al. (βs= –8.0 and –14.91, respectively). Pooled estimates of the association between low birth weight and prenatal exposure to PM2.5 throughout gestation showed significant increases: OR=1.09 (95% CI 1.03–1.15), based on 19 studies (Sun et al., 2016); OR=1.05 (95% CI 1.02–1.07), based on 6 studies (Zhu et al., 2015a); and OR=1.09 (95% CI 0.9–1.32), based on 4 studies (Sapkota et al., 2012). By trimesters, Sun et al. (2015) reported increased but not statistically significant risks of low birth weight for PM10 exposure in all three trimesters (ORs=1.026, 1.035, 1.233,
respectively). The Volume 11 committee notes that these findings are mixed and indicate the need for additional work in this area.
Changes in birth weight and low birth weight associated with PM10 throughout gestation were examined in two systematic reviews that reported statistically significant effects in 4 out of 12 studies (Bosetti et al., 2010) and four out of seven studies (Jacobs et al., 2017) as well as in two meta-analyses (Lamichhane et al., 2015; Sapkota et al., 2012). Lamichhane et al. (2015) examined changes in birth weight in 16 studies and reported β= –6.50 (95% CI –10.94– –2.5). The authors examined trimester-specific risks stratified by whether the study adjusted for smoking in its analyses and found that among smoking-adjusted studies the risks were significantly elevated for the third trimester (trimester effect sizes –1.43, –6.50, and –5.11), while the opposite was true for smoking-unadjusted studies: the first and second trimesters had significantly elevated risks (trimester effect sizes –3.31, –1.24, 1.36). Residual confounding by smoking was about 6.73%. Sapkota et al. (2012) reported a pooled estimated OR for low birth weight of 1.02 (95% CI 0.99–1.05), based on 11 studies, with trimester-specific estimates close to 1 (ORs=1.0, 1.0, 1.02, respectively).
Small for gestational age was assessed in two systematic reviews, both of which had three studies (out of three and four studies, respectively) finding significant effects associated with PM2.5 (Bosetti et al., 2010; Shah et al., 2011); Bosetti et al. (2010) found one (out of three studies) that reported significant effects associated with PM10. The Zhu et al. (2015a) meta-analysis found an association between small for gestational age and PM2.5 exposure throughout gestation, with an OR=1.15 (95% CI 1.10–1.20) based on six studies. Trimester-specific estimates were all slightly elevated and statistically significant (ORs 1.07, 1.06, 1.06, respectively). Collectively, the data suggest that PM may have a slight but significant effect on small-for-gestational-age pregnancy outcomes.
One recent publication that reported on 499 participants of the Belgian ENVIRONAGE study found that PM2.5 was related to thyroid hormone levels in cord blood, which may mediate effects on birth weight. For an IQR increase in PM2.5 (8.2 μg/m3) in the third trimester, cord blood levels of thyroid-stimulating hormone (TSH) declined 11.6% (95% CI –21.8– –0.1), and the free thyroxin/triiodothyronine ratio (FT4/FT3) decreased by 62.7% (95% CI –91.6– –33.8%). Only FT4 was significantly associated with decreases in birth weight, and the analysis estimated that FT4 levels explained 21% of the association (-19g, 95% CI –37– –1) between PM2.5 and birth weight (Janssen et al., 2017).
Fetal Death, Stillbirth, and Miscarriage Fetal mortality was not adequately captured by the systematic reviews and meta-analyses. Several studies have examined the associations between spontaneous abortion/miscarriage and stillbirth and PM exposure using hospital discharge data or birth records and registries. A number of studies on the relationship between exposure to traffic-related air pollutants (which includes PM2.5) and fetal death (spontaneous abortion and stillbirth) have reported significant associations (Arroyo et al., 2016; DeFranco et al., 2015; Enkhmaa et al., 2014; Faiz et al., 2012, 2013; Green et al., 2015; Hwang et al., 2011; Moridi et al., 2014).
Two studies examined the risk of miscarriage/spontaneous abortion with PM exposure. Candela et al. (2015) examined the risk of miscarriage (1,375 miscarriages among 11,875 pregnancies) in association with PM10 from local incinerators in Northern Italy. Although ORs reflected an increasing risk of spontaneous abortions with increasing quartiles of PM10 exposure (p-value for trend=0.042), none of the estimates were significantly elevated (highest quintile [PM10>1.33 ng/m3] versus no exposure OR=1.29, 95% CI 0.97–1.72). Although there was no significant interaction (p=0.19), risks of spontaneous abortion with high levels of PM10 exposure were observed among women with no history of previous spontaneous abortion (4th quintile OR=1.45, 95% CI 1.08–1.94; 5th quintile OR=1.44, 95% CI 1.06–1.96; p-value for trend=0.009). Enkhmaa et al. (2014) examined the correlations between monthly
hospitalizations for spontaneous abortions in 1,219 women and monthly PM2.5 and PM10 levels in Mongolia and found them to be significantly correlated (R=0.92 and 0.64, respectively, both p<0.001).
Stillbirth was reported by five independent studies. Faiz et al. (2013) examined 1,719 stillbirths in New Jersey in a cohort study (n=343,077) using a case-crossover design to investigate the effects of pollutants in the week before delivery; it found no significant associations with PM2.5. Green et al. (2015) reported on 13,999 stillbirths in several areas of California (compared with more than 3 million live births) and found no significant association between PM2.5 and stillbirths in any trimester. However, stratification by maternal demographic variables revealed significant effects of maternal age (25–34 years OR=1.07, 95% CI 1.0–1.16; ≥35 years OR=1.13, 95% CI 1.03–1.24) and maternal educational level (college or beyond OR=1.32, 95% CI 1.19–1.47). DeFranco et al. (2015) examined Ohio births (349,188 live births, 1,848 still births) within 10 km of an air monitoring station and found that high levels of exposure to PM2.5 (≥mean PM2.5 + IQR range) were associated with stillbirth in the third trimester (OR=1.42, 95% CI 1.06–1.91), with the association stronger for births within a 5-km radius (OR=1.54, 95% CI 1.08–2.20); however, no significant associations were found for the first or second trimesters or estimates when PM2.5 was analyzed as a continuous variable. A case-control study (9,325 cases and 93,250 controls) in Taiwan found that the risk of stillbirth was only associated with first-trimester PM10 exposure among preterm births (OR=1.03, 95% CI 1.0–1.07) after controlling for other pollutants (Hwang et al., 2011). Black smoke (PM <4 μg/m3) exposure during any trimester of pregnancy, which was studied by Pearce et al. (2010) among 812 stillbirths and 90,537 live births in the United Kingdom, was not associated with stillbirth. One study of cause-specific stillbirths (n=5,377) in California found no association between PM2.5 total mass and congenital malformations, fetal disorders, or maternal complications; however, PM2.5 was significantly associated with stillbirths caused by fetal growth (OR=1.23, 95% CI 1.06–1.44) and obstetric complications (OR=1.13, 95% CI 1.03–1.24), although not for those caused by maternal complications, congenital malformations, or fetal disorders (Ebisu et al., 2018).
In a study of 385 infant deaths before the age of 1 year among 465,682 births in Massachusetts, no association was found between PM2.5 throughout gestation or in each trimester and respiratory death, sudden infant death syndrome, or all causes of death (Son et al., 2017).
In EPA’s review of PM (EPA, 2009), toxicological evidence supported an association between PM and adverse reproductive outcomes. However, there were few data on biological plausibility to explain the link between PM exposure and reproductive outcomes, although systemic inflammation and oxidative stress were implicated. The Volume 11 committee identified several animal studies that examined PM exposure and reproductive effects that have appeared since the 2009 EPA review. These studies provide some additional mechanistic information and are discussed below.
Cao et al. (2015) showed decreased fertility (reduced sperm count, increased rates of sperm abnormalities, and testicular damage by histomorphology) with evidence of oxidative stress mediated by the PI3K/Akt pathway in male rats exposed to PM2.5. An additional study by Cao et al. (2017) indicated that PM2.5 causes reproductive injury by compromising the blood–testes barrier integrity in such a way that affects mating success, sperm parameters, epididymal morphology, the expression of markers of spermatogenesis, superoxide dismutase levels in testes, and the expression of blood–testes barrier junction proteins. Similarly, Liu et al. (2017a) showed that PM2.5, particularly in the winter time, adversely affects sperm motility, sperm malformations, testicular tissue injury, and testicular apoptosis—effects believed to be induced by endoplasmic reticulum stress.
Reduced gestational length and birth weight were reported by Blum et al. (2017) in mice exposed to ambient PM2.5 throughout gestation (6h/day on gestational day [GD] 0.5–16.5). This period aligns with milestones in human development.
An in vitro study found that cytotoxicity and genotoxicity differed by the size of the PM (<0.49, 0.49–0.97, 0.97–3 and >3 μm), with the smallest particles having the greatest bioactivity (Velali et al., 2016). Another in vitro study found that the chemical constituents bound to ambient PM, particularly PAHs, affected the cytotoxicity of PM in JEG-3 human placental cells (van Drooge et al., 2017). The PAHs were associated with the burning of biomass in rural areas.
The Volume 11 committee considered a wide spectrum of developmental effects that are evident in infants at or after birth and in children as they continue to develop. This section includes any developmental outcomes in infants and children, including birth defects, childhood cancer, autism spectrum disorder (ASD), neurodevelopment and behavioral outcomes, and respiratory effects, in relation to parental preconception or prenatal exposures that were reported in the literature. No studies of outcomes in adult offspring were identified by the committee.
A number of studies have investigated the associations between in utero exposure to PM and congenital anomalies, with heterogeneous findings (Chen et al., 2014; Vrijheid et al., 2011). Because of the wealth of data on PM exposures and birth defects, the text below summarizes the results of epidemiologic studies; the details about these studies (design, population, and exposures and outcomes) are summarized in Table 6-2.
Among studies that reported on congenital anomalies, two conducted systematic reviews and meta-analyses (Chen et al., 2014; Vrijheid et al., 2011). Vrijheid et al. (2011) conducted a systematic review and meta-analysis of 10 studies of congenital anomalies and air pollution published between 2002 and 2011. The authors reported an elevated summary estimate for atrial septal defects (OR=1.14, 95% CI 1.01–1.28) with continuous exposure to PM10, but the association was not significant in a binary analysis comparing high- to low-exposure groups (OR=1.23, 95% CI 0.91–1.67). Other defects (ventricular septal defects, coarctation of the aorta, and tetralogy of Fallot) were not associated with PM10 exposure. A second systematic review and meta-analysis by Chen et al. (2014) of 17 studies published between 2011 and 2014 found no significant association between any congenital anomaly and PM10.
The committee reviewed 25 publications from 22 individual studies that investigated associations between PM exposure and congenital anomalies, particularly congenital heart defects. Most studies examined exposure to PM10, used registries of birth defects, and were large population-based, case-control studies from countries around the world—China (4), Spain, Australia, Israel (2), Italy, Taiwan, and England (4)—as well as from several states in the United States (California , Massachusetts, New Jersey, Texas, and North Carolina) and the National Birth Defects Prevention Study (NBDPS) (3). Exposures were modeled on the basis of residential distance to air monitors or roadways, or on land use. Many studies controlled for other risk factors such as SES, maternal smoking, and alcohol use; none of the studies assessed any paternal or occupational exposures. All the studies examined exposure in the first trimester, generally weeks 2 or 3 to week 8 of gestation. Two studies examined preconception exposures (Yao et al., 2016; Zhu et al., 2015b). The results of the studies are briefly summarized below, and the details of the location, study population and size, and other characteristics are presented in Table 6-2.
TABLE 6-2 Characteristics of Studies of Birth Defects
|Agay-Shay et al. (2013)||Tel-Aviv, Israel||2000–2006||case-control||1,260 cases and 130,402 controls||PM2.5 and PM10||1st: weeks 3–8 of pregnancy||Congenital heart defects|
|Dadvand et al. (2011b)||Northern Congenital Abnormality Survey/England||1985–1996||case-control||2,713 cases and 9,975 controls||Black smoke||1st: weeks 3–8 of pregnancy||Congenital heart defects|
|Dadvand et al. (2011a)||Northern Congenital Abnormality Survey/England||1993–2003||case-control||2,140 case and 14,256 controls||PM10||1st: weeks 3–8 of pregnancy||Congenital heart defects|
|Dolk et al. (2010)||Northern Congenital Abnormality Survey/England||1991–1999||ecologic||9,085 cases||PM10—1996 annual mean at census ward level||Not specific||Congenital anomalies|
|Farhi et al. (2014)||Israel||1997–2004||cohort||216,730 infants||PM10||1st, 2nd, and entire pregnancy||Congenital anomalies|
|Gianicolo et al. (2014)||Italy||2001–2010||case-control||189 cases and 756 controls||TSP||1st: weeks 3–8 of pregnancy||Congenital anomalies|
|Girguis et al. (2016)||Massachusetts||2001–2008||case-control||3,461 cases, 7,816 controls||PM2.5||1st trimester||Cardiac, neural tube, and orofacial defects|
|Hansen et al. (2009)||Brisbane, Australia||1998–2004||case-control||150,308 births||PM10||1st: weeks 3–8 of pregnancy||Congenital heart defects and orofacial clefts|
|Hwang et al. (2015)||Taiwan||2001–2007||case-control||1,087 cases and 10,870 controls||PM10||Entire pregnancy||Cardiac defects|
|Liang et al. (2014)||Hainan province, China||2009–2011||case-control||509 cases and 63,391 controls||PM10||1st, 2nd, 3rd, and entire pregnancy||Congenital anomalies|
|Liu et al. (2017b)||Fuzhou, China||2007–2013||case-control||662 cases and 3,972 controls||PM10||1st: days 11–60||Congenital heart defects|
TABLE 6-2 Continued
|Marshall et al. (2010)||New Jersey||1998–2003||case-control||717 cases and 12,925 controls||PM2.5 and PM10||1st: weeks 3–8 of pregnancy||Orofacial clefts|
|Padula et al. (2013a)||San Joaquin Valley, CA||1997–2006||case-control||806 cases and 849 controls||PM2.5 and PM10||1st||Neural tube defects, orofacial clefts, gastroschisis|
|Padula et al. (2013b)||San Joaquin Valley, CA||1997–2006||case-control||822 cases and 849 controls||PM2.5 and PM10||1st||Congenital heart defects|
|Padula et al. (2013c)||San Joaquin Valley, CA||1997–2006||case-control||874 cases and 849 controls||PM2.5 and PM10||1st||Congenital anomalies, except congenital heart defects, neural tube defects, orofacial clefts, and gastroschisis|
|Padula et al. (2015)||San Joaquin Valley, CA||1997–2006||case-control||215 cases and 849 controls||PM2.5 and PM10||1st||Neural tube defects|
|Rankin et al. (2009)||Northern Congenital Abnormality Survey/England||1985–1990||case-control||3,197 cases and 15,000 controls||Black smoke||1st||Congenital anomalies|
|Schembari et al. (2014)||Barcelona, Spain||1994–2006||case-control||2,247 cases and 2,991 controls||PM2.5, PM10, and PMcoarse (10–2.5 μm)||1st: weeks 3–8 of pregnancy||Congenital anomalies|
|Stingone et al. (2014)||NBDPS||1997–2006||case-control||Not provided||PM2.5 and PM10||1st: weeks 2–8 of pregnancy||Congenital heart defects|
|Vinikoor-Imler et al. (2013)||North Carolina||2003–2005||cohort||322,969 live births||PM2.5||1st: weeks 3–8 of pregnancy||Congenital anomalies|
|Vinikoor-Imler et al. (2015)||Texas||2002–2006||case-control||21,060 cases and 1,401,611 controls||PM2.5||1st||Congenital anomalies|
|Warren et al. (2016)||NBDPS||2001–2006||case-control||1,986 cases and 2,741 controls||PM2.5||1st: weeks 2–8 of pregnancy||Congenital heart defects|
|Yao et al. (2016)||Anqing City, China||2010–2012||cohort||16,332||PM10||Preconception, 1st, 2nd, 3rd trimesters, and entire pregnancy||Congenital anomalies|
|Zhou et al. (2017)||NBDPS||2001–2007||case-control||7,035 cases and 4,697,523 births||PM2.5||1st: weeks 5–10 of pregnancy||Orofacial clefts|
|Zhu et al. (2015b)||Consortium on Safe Labor||2002–2008||retrospective cohort||188,102 births and fetal deaths||PM2.5 and PM10||Preconception and 1st: weeks 3–8 of pregnancy||Orofacial clefts|
NOTE: NBDPS=National Birth Defects Prevention Network (Arizona, Florida, New York [excluding New York City], and Texas); PM=particulate matter; TSP=total suspended particulate.
Among studies that examined PM in association with any birth defect, no increase was reported with PM2.5, PM10, or black smoke in the first trimester (Liang et al., 2014; Rankin et al., 2009; Schembari et al., 2014). Yao et al. (2016) examined the relationship between 398 children with birth defects among 16,332 births in Anqing City, China, and PM10 exposure during preconception and in the first, second, and third trimesters and found no associations. However, Liang et al. (2014) reported that birth defects in Hainan province, China (509 birth defects and 63,391 controls), were associated with PM10 exposure in the second and third trimesters (second-trimester OR=1.039, 95% CI 1.016–1.063; third-trimester OR=1.066, 95% CI 1.043–1.090).
Congenital heart defects were the most widely studied group of congenital anomalies, but the results were not consistent across studies. PM10 exposure throughout pregnancy was associated with significantly increased risks of all congenital heart defects (OR=1.07, 95% CI 1.0–1.14) (Farhi et al., 2014), ventricular septal defects (OR=1.18, 95% CI 1.01–1.39) (Farhi et al., 2014), and atrial septal defects (OR=2.52, 95% CI 1.44–4.42) (Hwang et al., 2015). Exposure to PM10 in the first 3 to 8 weeks of pregnancy was associated with multiple congenital heart defects (OR=1.05, 95% CI 1.01–1.10) (Agay-Shay et al., 2013). Another study that examined 10-day intervals in early pregnancy found significant increased risks of ventricular septal defects, atrial septal defects, and patent ductus arteriosus with the second quartile of PM10 exposure, but not the third or fourth quartiles, for days 31–40 and 41–50 (ORs ranging from 1.19 to 2.17) (Liu et al., 2017b), indicating that the response was not monotonic. Other studies reported no significantly increased risk of these outcomes in association with PM10 (Agay-Shay et al., 2013; Dadvand et al., 2011a; Gianicolo et al., 2014; Hansen et al., 2009; Hwang et al., 2015; Padula et al., 2013b; Schembari et al., 2014; Stingone et al., 2014; Yao et al., 2016) or a significantly reduced risk of ventricular septal defects with exposure in the first 3 to 8 weeks of pregnancy (OR=0.88, 95% CI 0.81–0.97; Schembari et al., 2014). Thus, the Volume 11 committee found the results to be mixed, and no clear explanation for these increases could be derived from the reports.
For PM2.5, seven studies reported no increased risk for any congenital heart defect or for most of the specific heart defects examined (Agay-Shay et al., 2013; Girguis et al., 2016; Padula et al., 2013b; Schembari et al., 2014; Stingone et al., 2014; Vinikoor-Imler et al., 2013, 2015; Warren et al., 2016). However, PM2.5 exposure was associated with increased risks of certain specific defects: namely, dextro-transposition of the great arteries with first-trimester exposure (3rd quartile of exposure OR=2.6, 95% CI 1.1–6.5) (Padula et al., 2013b), pulmonary valve stenosis with exposure on day 53 of gestation, and tetralogy of Fallot on days 50–51 of gestation (estimates not reported) (Warren, et al., 2016). PM2.5 exposure in the first 3 to 8 weeks of gestation was associated with reduced risks for ventricular septal defects (OR=0.83, 95% CI 0.72–0.97) (Schembari et al., 2014), septal heart defects (OR=0.79, 95% CI 0.76–0.83) (Vinikoor-Imler et al., 2015), obstructive heart defects (OR=0.88, 95% CI 0.79–0.98) (Vinikoor-Imler et al., 2015), and patent ductus arteriosus (OR=0.78, 95% CI 0.68–0.91) (Agay-Shay et al., 2013). Two studies found that exposure to black smoke was not associated with significant risk of congenial heart defects (Dadvand et al., 2011b; Rankin et al., 2009). Together this mixed evidence suggests that additional work on the risk of prenatal exposure PM for heart defects is needed.
Many studies investigated the associations between orofacial clefts and PM2.5 or PM10, and most found no significantly elevated risks (Farhi et al., 2014; Girguis et al., 2016; Hansen et al., 2009; Marshall et al., 2010; Padula et al., 2013a; Rankin et al., 2009; Schembari et al., 2014; Vinikoor-Imler et al., 2013, 2015). Two studies noted an association between PM and cleft palate but not cleft lip with or without cleft palate during certain windows of development. Zhou et al. (2017) reported an OR=1.43 (95% CI 1.11–1.86) per 10 μg/m3 increase in PM2.5 between 5 and 10 weeks of gestation among 2,609 cases and more than 4 million control births in the NBDPS. In a study of 159 cases and 188,102 live birth controls in the Consortium on Safe Labor cohort in the United States, Zhu et al. (2015b) observed an association
between cleft palate only and PM2.5 exposure in the first 3 to 8 weeks of gestation (OR=1.74, 95% CI 1.15–2.64), but the association did not extend to exposure to PM2.5 in the 3 months before conception. Conversely, Hansen et al. (2009) reported a reduced risk of cleft palate with PM10 exposure in the first 3 to 8 weeks of gestation (OR=0.69, 95% CI 0.50–0.93) among 145 cases and 150,308 controls in Brisbane, Australia. Yao et al. (2016) found no elevated risk of cleft palate or cleft lip with PM10 exposure in the 3 months before pregnancy among 16,332 births in Anqing City, China.
No increased risk of neural tube defects in association with PM has been reported for the U.S. general population. One study in California reported an interaction between PM10 and socioeconomic factors for the risk of spina bifida, with the risk elevated among people with low income and low education (Padula et al., 2015). Other studies have reported null findings for PM exposure and neural tube defects (Dolk et al., 2010; Girguis et al., 2016; Padula et al., 2013a; Schembari et al., 2014; Vinikoor-Imler et al., 2013, 2015).
Two studies examined chromosomal defects (specifically, trisomy 21, the cause of Down syndrome) in association with PM10. Farhi et al. (2014) found an increased risk in the first trimester (OR=1.14, 95% CI 1.01–1.28), whereas Dolk et al. (2010) did not. No obvious difference between the two studies has been identified.
Long lists of other congenital anomalies have been studied, with no significant effects reported in more than one study. Increased risks have been reported for abdominal wall defects with both PM10 and PM2.5 in the first 3 to 8 weeks of gestation (PM2.5 OR=1.33, 95% CI 1.0–1.76 and PM10 OR=1.36, 95% CI 1.14–1.62) (Schembari et al., 2014); esophageal atresia with quartiles of ambient PM10 concentration in the first trimester (2nd quartile OR=4.1, 95% CI 1.2–14.8; 3rd quartile OR=3.9, 95% CI 1.1–14.0; 4th quartile OR=4.9, 95% CI 1.4–17.2) (Padula et al., 2013c); nervous system anomalies with black smoke in the first trimester (OR=1.10, 95% CI 1.03–1.18) (Rankin et al., 2009); and omphalocele (a gastrointestinal defect) with PM10 exposure (RR=2.17, 95% CI 1.0–4.71) (Dolk et al., 2010). Padula et al. (2013c) reported decreased risks of hydrocephaly for specific quartiles of both PM2.5 and PM10 exposure in the first trimester: PM2.5 2nd quartile OR=0.2 (95% CI 0.1–0.7) and PM10 4th quartile OR=0.3 (95% CI 0.1–0.9). Also, a negative association (OR=0.78, 95% CI 0.64–0.96) between craniosynostosis and PM2.5 exposure in the first 3 to 8 weeks of gestation was reported by Vinikoor-Imler et al. (2015). Other studies reported no significant associations for these outcomes and others with exposure to PM10 (Dolk et al., 2010; Farhi et al., 2014; Padula et al., 2013a,c; Schembari et al., 2014; Yao et al., 2016) and PM2.5 (Padula et al., 2013a; Schembari et al., 2014; Vinikoor-Imler et al., 2013, 2015). Yao et al. (2016) investigated a range of birth defects with exposure to PM10 in the 3 months before conception and found no significant effects.
Three studies assessed the associations between childhood cancer and prenatal exposures to PM (Badaloni et al., 2013; Heck et al., 2013b; Lavigne et al., 2017). Badaloni et al. (2013) found no association between childhood leukemia and exposure to PM2.5 or PM10 in an Italian case-control study of 620 cases and 957 controls. Heck et al. (2013b) examined 3,590 cases of childhood cancer from the California Cancer Registry and compared them with 80,224 health children born in California and found no relationship between IQR increases in PM2.5 exposure during pregnancy and acute lymphoblastic leukemia (ALL), acute myeloid leukemia (AML), non-Hodgkin lymphoma (NHL), ependymoma, astrocytoma, intracranial and intraspinal embryonal tumors, medulloblastoma, primitive neuroectodermal tumors (PNETs), neuroblastoma, retinoblastoma, Wilms’ tumor, or hepatoblastoma in children under 6 years of age.
Another large registry-based study in a population of 2,350,898 children born in Ontario, Canada, also found no significant association between IQR increases in PM2.5 during pregnancy and ALL, AML, NHL, ependymoma, astrocytoma, medulloblastoma, PNET, neuroblastoma, retinoblastoma, Wilms’ tumor, hepatoblastoma, rhabdomyosarcoma, or germ cell tumor. However, a significant association with astrocytoma was reported for exposure in the first trimester and throughout the entire pregnancy (first-trimester hazard ratio [HR]=1.40, 95% CI 1.0– –1.86; second-trimester HR=1.26, 95% CI 0.95–1.67; third-trimester HR=0.96, 95% CI 0.71–1.28; entire pregnancy HR=1.35, 95% CI 1.0–1.85; first-year HR=1.07, 95% CI 0.79–1.28) (Lavigne et al., 2017).
In recent years, many publications have described studies of neurodevelopmental and behavioral outcomes in children exposed to various air pollutants, many of which have focused on ASDs. Therefore, the committee discussed the literature on ASD together, followed by studies that examined other neurodevelopmental outcomes, such as measures or scores to quantify the effects on intelligence quotient (IQ), cognition, motor skills, attention, memory, and behavior.
Autism Spectrum Disorders Nine studies examined ASDs in association with prenatal PM exposure (Becerra et al., 2013; Gong et al., 2014, 2017; Guxens et al., 2016; Kalkbrenner et al., 2015; Kerin et al., 2018; Raz et al., 2015; Talbott et al., 2015; Volk et al., 2013). Three of these studies considered preconception exposure (Kalkbrenner et al., 2015; Raz et al., 2015; Talbott et al., 2015).
Becerra et al. (2013) conducted a large population-based case-control study in Los Angeles County with 7,603 cases of children with a primary diagnosis of autistic disorder and 76,030 controls. No significant effects were reported in single-pollutant models, but increased risks were found for PM after controlling for NO2 exposure throughout the entire pregnancy per IQR increases (PM2.5 OR=1.07, 95% CI 1.01–1.12; PM10 OR=1.08, 95% CI 1.03–1.13). Trimester-specific estimates were inconsistent. Kerin et al. (2018) studied the severity and functioning of 327 children with ASD in California and found that PM10 and PM2.5 exposure during pregnancy was not associated with greater severity or decreased functioning.
Gong et al. (2017) conducted a large population-based case-control study (5,136 cases and 18,237 controls) in Sweden and found no association between PM10 exposure during pregnancy and ASD, either with or without intellectual disabilities. Another study by the same investigators examined ASD and attention deficit–hyperactivity disorder (ADHD) at 9 to 12 years of age among 3,426 Swedish twins and found no associations with PM10 exposure during pregnancy, the first year, or the ninth years of life (Gong et al., 2014). Guxens et al. (2016) analyzed four European cohorts totaling 8,079 children and found no association between borderline or clinical ASD and prenatal exposure to PM2.5 or PM10.
Kalkbrenner et al. (2015) examined the risk of ASD in 979 cases and 14,666 controls from North Carolina and San Francisco versus PM10 exposure starting 80 days before conception through the child’s first birthday. The only significantly elevated window of exposure was the third trimester (OR=1.36, 95% CI 1.13–1.63), and this association remained significant after further adjustment for covariates, but only among cases and controls in North Carolina, not San Francisco. No explanation for the site differences was presented. Preconception exposure was not associated with ASD (OR=0.94, 95% CI 0.82–1.08). Raz et al. (2015) conducted a nested case-control study of 245 cases and 1,522 controls using the Nurses’ Health Study II cohort to look for associations between ASD and exposures to PM2.5 and PM10–2.5. No significant association with PM10–2.5 exposure was found, but ASD was found to be related to IQR changes in PM2.5 exposure in all three trimesters (first-trimester OR=1.23, 95% CI 1.01–1.49; second-trimester OR=1.27, 95% CI 1.05–1.54; third-trimester OR=1.49, 95% CI 1.20–1.85). When
comparing exposures during the 9 months before pregnancy, during pregnancy, and after pregnancy, the authors found that exposures to PM2.5 during the preconception period or during the postnatal period were not related to ASD.
Talbott et al. (2015) reported a case-control study of 217 cases and 226 controls in Pennsylvania that assessed PM2.5 exposure from preconception through the second year of life. They found elevated, but not significant, risks for ASD per IQR changes for windows of exposure that reflected preconception through the first year of life (ORs ranged from 1.07 to 1.37), but only exposures in the second year of life were significantly associated with ASD (OR=1.45, 95% CI 1.01–2.08).
All of the studies described above measured exposure based on maternal residence at birth, except Volk et al. (2013), who collected data on the place of residence beginning 3 months before pregnancy until the time of assessment in order to account for residential mobility and better examine risk by trimester. In the study of 279 cases with full syndrome autism in California compared to 245 children with typical development, increased risks of autism were associated with elevated PM throughout pregnancy and the first year of life (PM2.5 per 8.7 μg/m3 increase: first-trimester OR=1.22. 95% CI 0.96–1.53; second-trimester OR=1.48, 95% CI 1.40–1.57; third-trimester OR=1.40, 95% CI 1.11–1.77; first-year OR=2.12, 95% CI 1.45–3.10; and PM10 per 14.6 μg/m3 increase: first-trimester OR=1.44, 95% CI 1.07–1.96; second-trimester OR=1.83, 95% CI 1.35–2.47; third-trimester OR=1.61, 95% CI 1.20–2.14; first-year OR=2.14, 95% CI 1.46–3.12). Volk et al. (2013) used two approaches to assigning exposure—regional air quality monitoring data and dispersion modeling—and found elevated risks with both approaches.
Of the three studies that considered preconception exposure, none reported an increased risk of ASD in association with PM2.5 or PM10 (Kalkbrenner et al., 2015; Raz et al., 2015; Talbott et al., 2015).
A meta-analysis of the studies summarized above reported a high degree of heterogeneity among the studies (Flores-Pajot et al., 2016). Still, the authors concluded that PM2.5 in the first trimester and in the postnatal period was significantly associated with ASD, although prenatal PM10 exposure was not. Pooled estimates for PM2.5 and PM10 for each window of exposure are reported in Table 6-3, derived from Flores-Pajot et al. (2016). A systematic review by Suades-Gonzalez et al. (2015) found an association between prenatal and postnatal PM2.5 exposure and ASD based on several high-quality studies. The significant impact of postnatal exposure on later ASD development is an important consideration when weighing the evidence on prenatal exposures.
TABLE 6-3 Pooled Estimates (95% CI) for Risk of Autism per 10 μg/m3 Increases in PM2.5 or PM10 Exposure
|Window of Exposure||OR PM2.5 (95% CI)||P-Value for Heterogeneity||OR PM10 (95% CI)||P-Value for Heterogeneity|
|Entire pregnancy||1.34 (0.83–2.17)*||0.000||1.03 (0.77–1.37)*||0.003|
|1st trimester||1.10 (1.02–1.19)||0.717||1.01 (0.86–1.20)*||0.008|
|2nd trimester||1.21 (0.88–1.66)*||0.000||1.12 (0.90–1.40)*||0.001|
|3rd trimester||1.33 (0.99–1.77)*||0.022||1.23 (0.97–1.55)*||0.000|
|Postnatal (1–4 yrs)||2.43 (1.61–3.68)||0.723||1.33 (0.86–2.05)*||0.002|
NOTE: A significant p value for heterogeneity (*p<0.05) indicates the presence of heterogeneity among the studies in the meta-analysis.
SOURCE: Flores-Pajot et al., 2016.
Other Neurodevelopmental Outcomes The committee identified nine studies that reported on associations between exposure to PM and a variety of neurodevelopmental and behavioral outcomes (Chiu et al., 2016; Eenhuizen et al., 2013; Guxens et al., 2014; Harris et al., 2016; Kim et al., 2014; Lertxundi et al., 2015; Lin et al., 2014; Stingone et al., 2016; Yorifuji et al., 2016). This section focuses on all neurodevelopmental and behavioral outcomes other than ASD. A systematic review by Suades-Gonzalez et al. (2015) evaluated 16 studies on PM and neuropsychological development and concluded that there was insufficient evidence of associations with cognitive and psychomotor development.
Lertxundi et al. (2015) studied cognitive and psychomotor development in 438 15-month-old Spanish children and found that reductions in motor scale scores were significantly associated with 1 μg/m3 increases in PM2.5 throughout pregnancy (β= –1.15, 95% CI –1.76– –0.54), but there was no association with mental-scale scores after adjustment for confounders and NO2 exposure. No adjustments were made for postnatal exposures, but results stratified by residential distance to metal processing plants showed significant relationships. Guxens et al. (2014) examined cognitive and psychomotor development among 1- to 6-year-olds in six European birth cohorts (n=9,482) and found no relationship between prenatal PM2.5, PM10, or PM10–2.5 exposure and either cognitive or psychomotor development.
Kim et al. (2014) conducted a cohort study (n=520) of cognitive and psychomotor development in South Korea. Using a general linear model, they found that a 10 μg/m3 increase in PM10 was significantly associated with decreased cognitive scores (β= –4.60, 95% CI –6.71– –2.49) and psychomotor scores (β= –7.24, 95% CI –9.79– –4.69) at 6 months, but not at 12 or 24 months. General estimating equations showed significant associations for both measures across 6 to 24 months of age (cognitive β= –2.83, 95% CI –4.71– –0.94; psychomotor β= –3.00, 95% CI –4.86– –1.14). Lin et al. (2014) conducted a cohort study of 533 Taiwanese children exposed to PM10 throughout pregnancy and up to 18 months of age and found that prenatal PM10 exposure was not associated with gross motor, fine motor, language, or social-personal scores at 6 and 18 months of age.
Yorifuji et al. (2016) examined motor and verbal developmental milestones in 2.5-year-olds and behavioral milestones in 5.5-year-olds in Japan in association with PM (<7 μm) exposure estimated for the entire pregnancy (n=46,039). Parents responded to questions about their children’s attention, self-regulation, and socially appropriate behavior as part of the national Longitudinal Survey of Babies in the 21st Century. Exposure in the 9 months before conception was adjusted for using regression models, but no adjustments were made for any exposures after birth. While no significant associations were found for milestones at age 2.5, there were several significant associations with 5.5-year milestones per IQR increases in prenatal PM (inability to listen without fidgeting OR=1.10, 95% CI 1.04–1.17; inability to focus on one task OR=1.06, 95% CI 1.00–1.13; inability to express emotions OR=1.10, 95% CI 1.05–1.16; and inability to keep promises OR=1.08, 95% CI 1.02–1.14).
Chiu et al. (2016) reported on IQ, attention, and memory in a cohort of 267 children (average age 6.5 yrs) in Boston and found differences between specific prenatal windows of exposure by sex. Children were administered the Wechsler Intelligence Scale for Children (WISC)-IV, the Conners’ Continuous Performance Test-II, and the Wide Range Assessment of Memory and Learning, 2nd Edition. While PM2.5 was not associated with these outcomes overall, among the boys significant associations were found between PM2.5 and lower IQ scores at 31–38 weeks of age, errors of omission at 20–26 weeks, slower reaction time at 32–36 weeks, and variability in response time at 22–40 weeks. Significant associations with PM2.5 among girls included decreased visual memory index at 18–26 weeks, decreased attention/concentration index at 8–18 weeks, and decreased general memory index scores at 12–20 weeks.
Harris et al. (2016) conducted a cohort study (Project Viva in Massachusetts) and reported on executive functioning and behavioral problems in 1,212 children around 8 years old in relation to PM2.5 and
black carbon exposures estimated for the third trimester, first 3 years, and first 6 years of life. Although some problems were associated with mid-childhood exposure, no executive functioning or behavioral problems were associated with prenatal exposure.
The association between exposure to PM and eczema in children was studied by Jedrychowski et al. (2011a,b) and Lu et al. (2017), while Orione et al. (2014) examined juvenile dermatomyositis, an autoimmune disease.
Jedrychowski et al. (2011a,b) used personal monitors to measure PM2.5 for 2 days in the second trimester of pregnancy. The risk of infantile eczema in a cohort of 469 Polish children was associated with a variety of risk factors (Jedrychowski et al., 2011a). While prenatal PM2.5 exposure alone was not significantly associated with eczema, elevated risks were reported among children exposed to high levels of PM2.5 and environmental tobacco smoke. In a follow-up study of 322 children through age 5, prenatal PM2.5 exposure alone was not associated with eczema, but when combined with maternal acetaminophen use during pregnancy, the risk of eczema symptoms doubled (HR=2.7, 95% CI 1.01–4.34) (Jedrychowski et al., 2011b).
Lu et al. (2017) reported on eczema in a cohort of 2,598 Chinese children ages 3 to 6 years and found no association with PM10 exposure in the 6 months before pregnancy or at any time in the prenatal period. Exposure measures for both mother and father in the 6 months before pregnancy were calculated based on air pollution monitoring results for the time period of interest, but they were not related to parental reports of eczema on their children.
Orione et al. (2014) conducted a small case-control study of 20 cases of juvenile dermatomyositis and 56 controls in Brazil to identify environmental risk factors, including those experienced during the prenatal period. There was no association between prenatal PM10 exposure and a diagnosis of juvenile dermatomyositis.
Asthma and Other Respiratory Outcomes
The Volume 11 committee considered 16 studies of PM and respiratory outcomes in children, most of which predominantly focused on asthma, but there were also explorations of wheeze, allergy, and lung function. Asthma was assessed in five cohort studies (Clark et al., 2010; Deng et al., 2015, 2016; Hsu et al., 2015; Tetreault et al., 2016; Yang et al., 2016) and one cross-sectional study (Liu et al., 2016b). Positive associations between prenatal exposure to PM2.5 and asthma were reported in three of the cohort studies (Clark et al., 2010; Hsu et al., 2015; Tetreault et al., 2016), and one of those reported an association with PM10 (Clark et al., 2010), although two did not observe an association between PM10 and asthma (Deng et al., 2015, 2016; Liu et al., 2016b). One meta-analysis examined prenatal exposure to pollutants and asthma or wheezing in children up to 14 years old and found no association between prenatal PM2.5 exposures and either outcome. However, the pooled estimate for PM10 was significant for childhood asthma (OR=1.08, 95% CI 1.05–1.12) (Hehua et al., 2017).
Hsu et al. (2015) studied a cohort of 736 children in Boston and found that prenatal PM2.5 exposure between 16 and 25 weeks of gestation was associated with a diagnosis of asthma in boys up to age 6, but not in girls. There was a significant difference between the sexes and a significant interaction between prenatal PM2.5 exposure and sex (p=0.01; ORs not reported). The sex differences were not explained.
Tetreault et al. (2016) reported a significant association between time-varying prenatal PM2.5 exposure and diagnosed asthma in children up to 12 years of age (per IQR increase HR=1.33, 95% CI
1.31–1.34). This large cohort of more than 1 million children in Quebec included 162,752 cases of diagnosed asthma. A subgroup analysis indirectly adjusted for secondhand smoke and found a much smaller effect of PM2.5 (per IQR, HR=1.04, 95% CI 1.02–1.06).
Clark et al. (2010) examined the incidence of asthma in 37,401 children ages 3–4 years in British Columbia in association with PM2.5, PM10, and black carbon. Two methods were used to estimate prenatal exposures. Regulatory air monitoring data and land use regression models for temporal variability were used to estimate individual exposures. PM2.5 was significantly associated with asthma incidence (per 1 μg/m3 increase, OR=1.02, 95% CI 1.0–1.03), as was black carbon (per 10–5/m increase in filter absorbance, OR=1.08, 95% CI 1.02–1.15); PM10 was not analyzed. Exposure, as determined by the inverse distance-weight summation of emissions from point sources, showed no association between PM2.5 and asthma incidence (per 1 μg/m3 increase, OR=0.95, 95% CI 0.91–1.0), but there was a significant association with PM10 (per 1 μg/m3 increase, OR=1.09, 95% CI 1.05–1.13); black smoke was not analyzed using this model.
Yang et al. (2016) studied 3,701 Dutch children through age 14. Those investigators found that exposure to PM2.5 estimated at the time of birth using a land use regression model was not associated with asthma or asthma symptoms in childhood.
In China, Deng et al. (2015, 2016) reported no association between prenatal PM10 exposure and asthma among 2,490 children ages 3 to 6 years. A separate Chinese cross-sectional study by Liu et al. (2016b) also found no association between asthma in 4- to 8-year-old children and prenatal PM10 exposure (n=3,358).
Yang et al. (2016) , Liu et al. (2016b), and Deng et al. (2015, 2016) also assessed respiratory allergy. Yang et al. did not find any relationship between PM2.5 concentrations at a child’s birth address and hay fever, rhinitis, or allergic sensitization in a cohort of Dutch children up to 14 years old. Liu et al. (2016b) found no association between allergic rhinitis and prenatal PM10 exposure in Chinese children who were around 5 years old. Deng et al. (2015) reported no association between prenatal PM10 and allergic rhinitis or eczema throughout pregnancy or in any specific trimester among Chinese children ages 3–6 years old (Deng et al., 2016).
Chiu et al. (2014) reported significant associations between repeated wheezing among 708 Boston children up to 2 years of age and high prenatal exposure to PM2.5 and black carbon versus low exposure to these two pollutants (PM2.5 OR=2.02, 95% CI 1.20–3.40; black carbon OR=1.84, 95% CI 1.08–3.12). These investigators also reported that psychosocial stressors such as community violence were important contributors to the occurrence of wheezing. Jedrychowski et al. (2010b) found a significant effect of prenatal PM2.5 exposure on wheezing at 2 years of age (high versus low IRR=1.377, 95% CI 1.252–1.514) but not at 4 years of age (high versus low IRR=1.063, 95% CI 0.923–1.223) among a cohort of 339 Polish children. Exposures were determined by mothers using personal monitors for 2 days in the second trimester. Rosa et al. (2017a) studied wheezing in 552 children in Mexico City in association with prenatal PM2.5 exposure and maternal stress and found a significant interaction. PM2.5 exposure in the first trimester was associated with the risk of current wheezing in children in the presence of high maternal stress (per IQR increase RR=1.35, 95% CI 1.00–1.83), but not in the case of low maternal stress. No associations were noted for other trimesters.
Respiratory system infections were examined in four studies. Liu et al. (2016b) found no association between pneumonia in 4- to 8-year-old Chinese children and prenatal PM10 exposure. MacIntyre et al. (2014) conducted a meta-analysis of 16,059 children from 10 European cohorts (ESCAPE project) and found significant associations between pneumonia in the first 2 years of life and PM10 (per 10 μg/m3 increases, OR=1.76, 95% CI 1.00–3.09) and PM2.5–10 (OR=1.24, 95% CI 1.03–1.5), but not for PM2.5. No associations between PM and croup or otitis media were found. In a study of 1,617 children ages 3–4
years in Changsha, China, prenatal PM10 exposure was found to have no association with otitis media (Deng et al., 2017). Recurrent broncho-pulmonary infections (≥5 episodes) in 214 children through age 7 were associated with prenatal PM2.5 exposure (OR=2.08, 95% CI 1.08–4.06) (Jedrychowski et al., 2013). Exposures were determined by mothers using personal monitors for 2 days in the second trimester.
Lung function was reported on by six studies. Liu et al. (2016b) found no association between prenatal PM10 exposure and lung function in 4- to 8-year-old Chinese children (n=3,358). Gehring et al. (2013) analyzed the lung function of children age 6 or 8 years from five European cohorts and found no association with PM2.5, PM10, or PM10–2.5 at the birth address (n=5,921).
Among healthy 5-week-old newborns in Switzerland (n=265), Latzin et al. (2009) found that maternal exposure to PM10 during pregnancy had affected tidal breathing (minute ventilation β=24.9, 95% CI 9.3–0.5), but not multiple breath washout or inflammatory markers; no effect was found for postnatal exposure. Jedyrchowski et al. (2010a) reported decrements in lung function in a cohort of 176 Polish children with maternal personal monitoring conducted in the second trimester. Lung function was significantly reduced at 5 years of age among children exposed prenatally to the highest quartile (>52.6 μg/m3) compared to the lowest quartile of PM2.5 (FVC β= –91.92, 95% CI –159.60– –24.24; and FEV1 β= –87.71, 95% CI –151.85– –23.57). Although there was a step-wise trend across the quartiles, the lung function of children in the second- and third-exposure quartiles were not significantly lower than in the first quartile.
Prenatal exposure to PM10 reduced respiratory function in 232 children with asthma ages 6–11 years in the San Joaquin Valley. Notably, the subgroups most affected by prenatal exposure to pollutants were children who were African American, diagnosed before age 2, and exposed to maternal smoking during pregnancy (Mortimer et al., 2008).
Other Outcomes Studied in Children
Single studies have reported on vitamin D levels, metabolic function, body weight, and pubertal development in children with prenatal exposure to PM.
In a French birth cohort of 370 pregnant women (EDEN study), Baiz et al. (2012) found that decreased vitamin D levels in cord blood were associated with maternal exposure to PM10 in the first trimester (10 μg/m3 PM10; β= –0.35, 95% CI –0.78– –0.03) and in the third trimester (β= –0.43, 95% CI –0.72– –0.14). In a Canadian cohort (MIREC) study of 1,257 mother–child pairs, researchers examined the correlation of maternal PM2.5 exposure with the levels of leptin and adiponectic—two measures of metabolic function—in cord blood (Lavigne et al., 2016). IQR increases of 3.2 μg/m3 in PM2.5 exposure throughout pregnancy were associated with increased adiponectin levels (β=11%, 95% CI 4–17%) and leptin levels (β=11%, 95% CI 1–21%); however, when birth weight was included in the adjustments, the increases in leptin levels ceased to be significant, although the increases in apiponectin levels retained significance.
Kim et al. (2016) studied changes in body weight measured from birth to age 5 in relation to prenatal PM10 exposure in 1,129 South Korean children. They did not find a relationship between prenatal PM10 exposure and changes in weight-for-age. However, a significant association was found in Bangladeshi children under 5 years old between the second, third, and fourth quartiles of mean PM2.5 exposure throughout pregnancy and stunting, wasting, and underweight (RRs ranged from 1.109 to 1.272). Exposures after birth were not significantly related to the outcomes studied (Goyal and Canning, 2017).
A few studies have reported increases in markers of obesity with PM exposure. A study of 239 children about 4 years old in the Boston area reported increased anthropometric measures associated with prenatal PM2.5 exposure estimated weekly. In boys, increased PM2.5 during gestational weeks 8–17 was
significantly associated with increased BMI z-scores, and PM2.5 in weeks 15–22 was significantly associated with fat mass. In girls, increased waist-to-hip ratio was significantly associated with increased PM2.5 in gestational weeks 10–29 (relative risks not reported) (Chiu et al., 2017). Mao et al. (2017) reported increased risks of childhood overweight or obesity among 1,446 children ages 2–9 years in the Boston area as a function of PM2.5 in each trimester (per IQR increases RR=1.1, 95% CI 1.0–1.2 for each trimester). Significant associations were also reported for exposure in the prenatal period (per IQR increase RR=1.1, 95% CI 1.0–1.2) and the first 2 years of life (per IQR increase RR=1.1, 95% CI 1.0–1.2).
Thiering et al. (2013) examined the relationship between estimated PM10 and PM2.5 concentrations at the mother’s residential address and insulin resistance in children at age 10 in two German birth cohorts (n=397). A two-standard-deviation increase in PM10 was associated with a 17.5% increase in insulin resistance (95% CI 1.9–35.6).
In a study of 4,074 children about 11 years old in Hong Kong, China, prenatal PM10 exposure was significantly related to later pubertal development in girls (Tanner stage [a measure of puberty] mean difference= –0.05, 95% CI –0.08– –0.02), but no differences were noted in boys (Huang et al., 2017).
In EPA’s 2009 review of PM, toxicological evidence was found to support an association between PM and adverse developmental outcomes. However, there were few data on biological plausibility to explain the link between PM exposure and developmental outcomes, although developmental windows of susceptibility and systemic inflammation and oxidative stress were implicated. The Volume 11 committee identified a few animal studies published since that 2009 review that examined PM exposure and developmental or generational effects.
Tang et al. (2017) found decreased lung volume parameters, compliance, and airflow on expiration and also lung damage related to oxidative stress in rat pups on postnatal day 28 after in utero exposure to PM2.5 (0.1, 0.5, 2.5, or 7.5 mg/kg once every 3 days from GD 0–18).
In rats exposed in utero to PM2.5 (0.375, 1.5, or 6.0 mg/kg on GD 10, 12, 14, and 16) researchers found various signs of cardiotoxicity, including a disordered arrangement of myocardial fibers, irregular-shaped cardiomyocytes, increased apoptotic rates in myocardium, and altered expression of mitochondrial fusion/fissure genes (Wang et al., 2017b). The exposure of pregnant rats to PM2.5 at levels corresponding to ambient levels in an urban Chinese environment for 24 hours per day throughout gestation and lactation resulted in changes related to cardiac development in the offspring. PM2.5 was associated with reduced levels of messenger ribonucleic acid (mRNA) and the proteins associated with cardiac precursor cells and cardiac development (GATA4 and Nkx2-5) and increased levels of TNF-α and IL-1β (inflammatory cytokines) in plasma, and it exacerbated the negative effects of homocysteine on cardiac development (myocardial apoptosis and structural abnormalities) (Chen et al., 2017a). Cardiac dysfunction was observed in adult rats after in utero exposure to PM2.5 (average 73.61 μg/m3 for 6 h/day throughout pregnancy), including an acute inflammatory response, chronic matrix remodeling, and an altered expression of Ca2+ handling proteins (Tanwar et al., 2017a). In mice, exposure to PM2.5 throughout gestation (average 91.78 μg/m3 for 6h/day, 5 days/week) resulted in adverse functional cardiac changes in early adolescence, including fractional shortening, impaired cardiomyocyte functioning, and reduced calcium transport amplitude and fluorescence decay rate. The authors stated that gestational exposure to PM2.5 alters calcium-handling proteins, which is responsible for the cardiac dysfunction observed in neonatal nice (Tanwar et al., 2017b).
An in vitro study found that cytotoxicity and genotoxicity differed by the size of PM (<0.49, 0.49–0.97, 0.97–3 and >3 μm), with the smallest particles having the greatest bioactivity (Velali et al.,
Epigenetic and Other Effects on DNA
Studies that focused on the genetic and epigenetic effects of prenatal PM exposure are summarized in Table 6-4. The section below provides an overview of the studies, and Table 6-4 provides additional study details, such as the tissue and cell type examined.
The epigenetic literature on PM has relied on relatively large sets of samples and studies, including ENVIRONAGE (Grevendonk et al., 2016; Janssen et al., 2012, 2013, 2015; Saenen et al., 2015, 2017; Wincklemans et al., 2017), the Children’s Health Study (Breton et al., 2016; Salam et al., 2012), PROGRESS (Programming Research in Obesity, Growth, Environment and Social Stressors) (Rosa et al., 2017b), and various consortiums and group studies (Pedersen et al., 2015; Silva da Silva et al., 2015). Three tissue types—cord blood, placenta, and peripheral blood lymphocytes—have generally been evaluated. These tissues represent a collection of different cell types, each with very different epigenomes, including different DNA methylation profiles. These studies of PM epigenetic effects have not accounted for the variability associated with heterogeneous cell types in samples. Differences in cell composition across different samples may result in what appear to be methylation differences but which actually are reflections of differences in cell purity or differences induced by immune responses (Houseman et al., 2012, 2015). Such complexities were interpreted by the Volume 11 committee to result in equivocal results.
Another challenge in interpreting results from these observational studies is that estimates of exposure are often confounded with other potential modifiers of the epigenome, as PM is frequently estimated longitudinally through source-distance models. Potential confounders, such as smoking or other environmental exposures that have been reported to be associated with altered DNA methylation (Jenkins et al., 2017; Knopik et al., 2012), are often not accounted for in these analyses, illustrating the complexity of the problem. Furthermore, the mutagenic activity of organic extracts from PM2.5 has been associated with increases in genomic DNA fragmentation (Silva da Silva et al., 2015) and higher DNA adduct levels as PM2.5 exposure increases (Pedersen et al., 2015).
Effects that change with the developmental timing of the exposure have also been reported, as different tissues are susceptible in different developmental periods. For example, during the last month of pregnancy a 16.1% decrease in the placental DNA content of cord blood (95% CI –25.2– –6.0%, p=0.003) has been reported, but no such decrease in the mitochondrial DNA content (Janssen et al., 2012). In that particular study, PM2.5 exposure over the entire pregnancy was positively associated with mitochondrial DNA methylation (MT-RNR1: +0.91%, p=0.01 and D-loop: +0.21%, p=0.05), but the exposure was inversely associated with the mitochondrial DNA content (relative change of –15.60%, p=0.001) in placental tissue (Janssen et al., 2015). Others have measured levels of oxidative stress induced by PM10 and PM2.5, as measured by 8-hydroxy′-deoxyguanosine (8-OHdG) in mitochondria. Maternal PM10 and PM2.5 exposures are both associated with increased 8-OHdG through the entire pregnancy (18.3%, 95% CI 5.6–33.4%, p=0.004; and 13.9%, 95% CI 0.4–29.4%, p=0.04, respectively), which is heightened in the third trimester (28.1%, 95% CI 8.6–51.2%, p=0.004; and 28.1%, 95% CI 3.6–58.4%, p=0.004, respectively) (Grevendonk et al., 2016). An association between 8-OHdG in cord blood and PM10 was observed only in the first and second trimesters (23.0%, 95% CI 5.9–42.8%, p=0.007; and 16.6%, 95% CI 1.8–33.5%, p=0.03, respectively). The Volume 11 committee notes that those statistically significant results do not account for the presence of multiple comparisons, nor were any sensitivity analyses performed.
TABLE 6-4 Summary of Genetic and Epigenetic Effects of Particulate Matter
|Janssen et al. (2012)||Cohort—ENVIRONAGE 178 newborns born at a Belgium hospital 2010–2011.||Modeled PM10 based on mother’s residence.||mtDNA content in placenta and cord blood.
No association between PM10 and mtDNA in cord blood. PM10 (10 μg/m3 increments) associated with decreased mtDNA in third trimester (–17.4%, 95% CI –31.8– –0.1%), last month of pregnancy (–16.1%, 95% CI –25.2– –6.0%), and last week of pregnancy (–10.1%, 95% CI –17.6– –1.9%).
|Janssen et al. (2013)||Cohort—ENVIRONAGE 240 newborns born at a Belgium hospital.||Modeled PM2.5 based on mother’s residence.
Exposure estimates accounted for mother’s change of address during pregnancy.
|Global DNA methylation in placenta.
Reduced global methylation per 5 μg/m3 increase in PM2.5 in first trimester (–2.13%, 95% CI –3.71– –0.54%) and through entire pregnancy (–2.19, 95% CI –3.65– –0.73%) associated with DNA global methylation.
Indicates that fertilization through implantation is a susceptible window.
|Janssen et al. (2015)||Cohort—ENVIRONAGE 381 newborns born at a Belgium hospital.||Modeled PM2.5 based on mother’s residence.||mtDNA methylation in placenta. IQR increases in PM2.5 associated with increased methylation at MT-RNR1 (0.91%, p=0.01) and D-loop (0.21%, p=0.05), and decreased mtDNA content (–15.6%, p=0.001).|
|Winckelmans et al. (2017)||Cohort—ENVIRONAGE 142 newborns born at a Belgium hospital.||Modeled PM2.5 based on mother’s residence. Short-term exposure = exposure in the last month of pregnancy; long-term exposure = annual average before delivery.||Whole-genome gene expression (RNA) in cord blood.
Differences in gene expression differed by duration of exposure (short- versus long-term) and infant sex.
|Grevendonk et al. (2016)||Cohort—ENVIRONAGE 224 mothers and 293 newborns born at a Belgium hospital.||Modeled PM2.5 and PM10 based on mother’s residence.||Mitochondrial 8-OHdG and mtDNA content in maternal and cord blood.
Maternal mitochondrial 8-OHdG associated with third trimester PM10 (28.1% change/IQR, 95% CI 8.6–51.2%) and PM2.5 (28.1% change/IQR, 95% CI 3.6–58.4%); first and second trimesters not significant. Cord blood mitochondrial 8-OHdG associated with first and second trimester PM10 (23.0% change/IQR, 95% CI 5.9–42.8%; and 16.6% change/IQR, 95% CI 1.8–33.5, respectively); third trimester PM10 and PM2.5 not significant.
Levels of 8-OHdG positively correlated with mtDNA content in maternal blood (β=0.38, 95% CI 0.28–0.49) and cord blood (β=0.52, 95% CI 0.43–0.61).
|Saenen et al. (2015)||Cohort—ENVIRONAGE 90 mother–newborn pairs at a hospital in Belgium 2010–2012.||Modeled PM2.5 based on mother’s residence.
Estimated exposure for preimplantation period (5 days).
|Fetal brain development measured by placental expression of BDNF, and associated pathways (% change per 5 μg/m3 PM2.5). In first trimester of pregnancy, –25.9% change BNDF (–28.7– –0.2%), specifically in the first month and around the time of implantation. Not significant for second and third trimesters. SOS2 and SYN1 also significant decreased expression in early pregnancy (data not reported).
SOS and PLCG signaling cascades reduced in first month. SOS β= –0.1, 95% CI –0.2– –0.07; PLCG β= –0.08, 95% CI –0.1– –0.03.
|Saenen et al. (2017)||Cohort—ENVIRONAGE 361 mother–newborn pairs at a hospital in Belgium 2010–2013.||Modeled PM2.5 based on mother’s residence, accounting for residential mobility during pregnancy.||LEP methylation status in the placenta was negatively associated with PM2.5 exposure during the second trimester. An interquartile range increase (7.5 μg/m3) in second-trimester PM2.5 exposure was associated with 1.4% (95% CI –2.7– –0.19%) decrease in LEP methylation.|
|Breton et al. (2016)||Cohort—Children’s Health Study
392 children ages 11–12 years with newborn blood spot tests.
|Prenatal PM exposure estimated from community monitors based on address at birth.||Newborn LINE1 and Alu methylation, and blood pressure and carotid intima-media thickness in childhood.
PM2.5 and PM10 had no association with blood pressure or intima-media thickness, LINE1 Alu methylation.
PM10 associated with LINE1 methylation in first trimester only (β= –0.66, 95% CI –1.22– –0.09).
|Rosa et al. (2017b)||Cross-sectional—PROGRESS
456 pregnant women in Mexico City 2007–2011.
|PM2.5 modeled based on maternal home address at birth.||mtDNA in cord blood leukocytes.
Significant reduction in mtDNA content per 10 μg/m3 increase in PM2.5 in weeks 35–40 only. No data reported.
|Cai et al. (2017)||Case-control
80 newborns with grown restriction and 101 healthy newborns in China 2011–2013.
|PM10 averages based on nearby air monitoring station during pregnancy.||Placental DNA methylation of LINE1, HSD11B2, and NR3C1 In cases, LINE1 methylation decreased with first trimester PM10 (10 μg/m3) (β= –1.78%, 95% CI –3.35– –0.22%); HSD11B2 methylation increased with first and second trimester PM10 exposure (β=1.03%, 95% CI 0.07–1.98%; β=2.23%, 95% CI 0.69–3.76%); No associations in health controls and NR3C1 methylation not associated with PM10 in cases or controls.|
TABLE 6-4 Continued
|Pedersen et al. (2015)||Cross-sectional—NewGeneris 511 pregnant women from 5 European countries 2006–2010.||Modeled PM2.5 based on maternal home address.||Bulky DNA adducts in cord blood.
Mean ratio of bulky DNA adduct levels per 5 μg/m3 increment of PM2.5 =1.14 (95% CI 0.99–1.32; p=0.07).
NOTE: 8-OHdG=8-hydroxy-2′-deoxyguanosine; CI=confidence interval; DNA=deoxyribonucleic acid; ENVIRONAGE=Environmental Influence ON Early AGEing; HSD11B2=11β-hydroxysteroid dehydrogenase type 2; IQR=interquartile range; LEP=leptin; mtDNA=mitochondrial DNA; PM=particulate matter; PROGRESS=Programming Research in Obesity, Growth, Environment and Social Stressors; RNA=ribonucleic acid.
In principle, DNA methylation can also vary depending on temporal and spatial measurements. For example, global changes in genome methylation in placental tissue and PM2.5 exposure during pregnancy are correlated, particularly with first-trimester exposure. A multi-lag model found significant associations between lower levels of global DNA methylation and PM2.5 exposure in the first trimester (–2.13% per 5 μg/m3 increase, 95% CI –3.71– –0.54%, p=0.009), and exposure during implantation (6–21d) showed an even stronger association (–1.08% per 5 μg/m3 increase, 95% CI –1.80– –0.36%, p=0.004); both of them led to an increased hypomethylated state (Janssen et al., 2013). During the first trimester of pregnancy, each 5 μg/m3 increase in PM2.5 was associated with a 15.9% decrease (95% CI –28.7– –3.2%, p=0.015) in the expression of placental brain-derived neurotrophic factor (BDNF) at birth and a 24.3% decrease (95% CI –42.8 – –5.8%, p=0.011) in synapsin-1 (Saenen et al., 2015), two genes involved in neurodevelopment. Using newborn blood spots, Breton et al. (2016) found that PM2.5 exposure during the first trimester was associated with a reduction in repetitive element methylation (i.e., LINE1, AluYb8). The degree of methylation was found to be genotype-dependent when 11 SNPs for DNA methyltransferase 1(DNMT1), DNA methyltransferase 3 Beta (DNMT3B), Tet methylcytosine dioxygenase 2 (TET2), and thymine DNA glycosylase (TDG) genes were assayed (Breton et al., 2016). Furthermore, differences in response have been reported with differences in the duration of exposure and infant sex. Wincklemans et al. (2017) reported differential gene expression to be associated with PM2.5 exposure—in particular, whether the exposure was short or long term (a month or a year before birth)—and also sex. For example, RNA expression reflecting DNA damage was altered in infants of both sexes with both short- and long-term exposure, whereas expression of RhoA pathways was affected in boys with short-term exposure, and defensin expression was altered in girls with long-term exposure.
The impacts of in utero PM exposure on gene expression in placenta have been studied in several populations (ENVIRONAGE, Rhode Island, and China) focusing on different genes. PM2.5 has been reported to affect the expression of LEP (leptin, an energy-regulating hormone) (Saenen et al., 2017), insulin-like growth factor-2 (IGF2), MEG3, MEST, and PLAG1 (all associated with infant birth weight); PM10 decreased the placental methylation of LINE1 (to represent global methylation) and increased the methylation of HSD11B2 (which encodes for 11β-hydroxysteroid dehydrogenase type 2, involved in glucocorticoid metabolism) (Cai et al., 2017).
These reports of changes in the methylation status of both the nuclear and the mitochondrial genomes, globally as well as at specific sites, in response to PM stress during pregnancy are intriguing and provide possible basis for the impact of PM on early development (Soto-Martinez and Sly, 2010; Teng et al., 2016). However, this body of evidence has several limitations, including small sample sizes and the heterogeneity of cell types, which affect both the reliability and the interpretation of the data. Biochemical mechanisms for how these exposures result in methylation changes have not been fully described, nor have these findings been confirmed in animals.
Although some research has shown that PM has epigenetic effects, including DNA methylation and mitochondrial DNA methylation, the impacts of those changes on the health of children are not clear, and the understanding of the mechanisms is incomplete. Research in animals has emphasized the role of gene–environment interactions, the modulation of effect as a function of development (Soto-Martinez and Sly, 2010), and exposure to multiple mixtures (Silva da Silva et al., 2015). In utero PM exposure has been reported to play a role in early cardio, orofacial, and other defects (Teng et al., 2016); the transgenerational inheritance of ovarian disease (Nilsson et al., 2012); cardiovascular disease in adulthood (Tanwar et al., 2017a); brain defects (Chao et al., 2017); and many pathways, including antioxidant response, xenobiotic metabolism, inflammatory signaling, and endothelial dysfunction (Thomson et al., 2013). While using animal models makes it possible to control exposures and permits the study of defined compositions of multiple elements and tissues from multiple organ sources, it lacks the temporal resolution and cumulative exposures that are observed in humans and their individualized responses.
Synthesis and Conclusions
This literature for PM includes data on populations from around the world, often with relatively high exposures as a result of traffic or industrial pollution. Most studies examined environmental exposures, with only a few of them assessing occupational exposures.
The relevance of any of these studies to the exposures likely experienced by veterans is unclear. Exposures described in this literature are primarily continuous environmental exposures, whereas some veterans’ exposures ceased upon their return from deployment, such as exposure to burn pits, while others may have continued, such as exposure to diesel exhaust and PM. Thus, the studies reviewed here do not necessarily address the long-term effects of concern to veterans. There is also a lack of data on the reproductive or developmental outcomes that may be expected following male or female exposures to combustion products prior to conception or, in the case of female service members, exposures in early pregnancy before they would have left the war zone. Because there are so few data on preconception and first-trimester exposures, particularly in humans, the committee’s conclusions are based on prenatal exposures in general and not specifically on preconception or first-trimester exposures.
Several methodological issues make interpretation of the studies difficult: the heterogeneity of study designs and exposure assessment methods, the complexity of pollutant mixtures, and a limited mechanistic understanding. Exposure is frequently estimated rather than measured, using methods such as the distance between maternal residences at birth and an air monitoring station or modeling based on satellite data and land use regression at varying units of area, such as zip code or census tract. The potential for exposure misclassification using such approaches is of concern, particularly as compared with studies that conduct air monitoring for each individual in a prospective manner. Studies that adjust for changes in a subject’s residence or the sources of pollution over time (such as increased traffic or the discontinuation of industrial sources) are more helpful than those that do not. Studies that measure outcomes in children, sometimes many years after exposure was estimated in pregnancy, often do not account for the expected changes in exposures that occur after birth.
Reproductive Effects in Men and Women Adverse effects on female fertility were noted for both PM2.5 and PM10 (Frutos et al., 2015; Nieuwenhuijsen et al., 2014; Slama et al., 2013). However, no effects on fertility and endometriosis were found in the Nurses’ Health Study (Mahalingaiah et al., 2014, 2016). The effects of both PM2.5 and PM10 on sperm count, concentration, motility, and morphology have been inconsistently reported (Hansen et al., 2010; Jurewicz et al., 2015; Lafuente et al., 2016; Wu et al., 2017; Zhou et al., 2014).
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between PM exposure and reproductive effects in men or women.
Adverse Pregnancy Outcomes The Volume 3 committee concluded that there was “limited/suggestive evidence of an association” between prenatal exposure to combustion products and low birth weight and preterm birth. However, they also found that the evidence was insufficient/inadequate to determine whether any association existed between reproductive outcomes and exposure during specific time periods (e.g., trimesters).
EPA (2009) concluded that the evidence was “suggestive of a causal relationship between long-term exposures to PM2.5 and reproductive . . . outcomes” but that the evidence was inadequate for PM10–2.5.
EPA justified its conclusion based on reports that PM2.5 had a greater impact on birth weight than did PM10 and on the consistent findings in epidemiologic studies of significant associations between PM2.5 and preterm birth and growth restriction. Animal studies also reported these effects but provided little information on the mechanisms or biologic plausibility.
Since 2009 the epidemiologic literature has grown, and it now supports associations between maternal exposure to PM during pregnancy and adverse pregnancy outcomes, including low birth weight, preterm birth, and small for gestational age (Shah et al., 2011; Zhu et al., 2015a). In this literature, the relationship between exposure to air pollution, especially PM2.5, and low birth weight is the most robust (Shah et al., 2011; Sun et al., 2016; Zhu et al., 2015a). Associations between exposure to PM2.5 and preterm birth and small for gestational age have been reported in systematic reviews and meta-analyses (Shah et al., 2011; Sun et al., 2015; Zhu et al., 2015a). Preterm birth, low birth weight, and small for gestational age are known to increase a child’s risk for subsequent health problems later in the lifecourse (Hack et al., 1995; Saigal and Doyle, 2008). Data on trimester-specific risks are inconsistent, but increased risks of preterm birth and decreases in birth weight associated with PM exposure in the third trimester are consistently reported. Exposure to PM during pregnancy has been associated with changes in biomarkers that are indicative of oxidative stress and inflammation processes, both of which may contribute to preterm birth and low birth weight.
Exposure to PM was also associated with miscarriage in two studies (Candela et al., 2015; Enkhmaa et al., 2014). The association between PM and stillbirth is inconsistent, with one study reporting increased risks with high levels of PM2.5 exposure in the third trimester (DeFranco et al., 2015) and another reporting increased risks with PM10 exposure in the first trimester (Hwang et al., 2011). Three other studies considered by the committee showed no such effect of PM2.5 (Faiz et al., 2012, 2013; Green et al., 2015; Pearce et al., 2010).
The risk of gestational diabetes and PM exposure was studied in three cohorts, but the results were mixed (Fleisch et al., 2014; Hu et al., 2015; Robledo et al., 2015). Preeclampsia and gestational hypertension were more extensively studied. Although PM10 was generally not associated with preeclampsia (Dadvand et al., 2013; Mobasher et al., 2013; van den Hooven et al., 2011; Wu et al., 2011), one study found increased risks of preeclampsia associated with PM10 exposure in mothers who did not smoke (Lee et al., 2013). Four studies of PM2.5 found some positive association with preeclampsia (Dadvand et al., 2013; Lee et al., 2013; Mobasher et al., 2013; Wu et al., 2011), but the timing of exposure was variable, and two studies did not find any association (Rudra et al., 2011; Savitz et al., 2015). Preconception exposure was considered by Rudra et al. (2011), who found no relationship between preconception exposure to PM2.5 and preeclampsia. Additionally, four studies reported increased blood pressure or gestational hypertension (a characteristic of preeclampsia) with PM exposure (Lee et al., 2012; van den Hooven et al., 2011; Vinikoor-Imler et al., 2012; Xu et al., 2014), and one study found no such effects (Hampel et al., 2011).
The Volume 11 committee concludes that there is sufficient evidence of an association between prenatal exposure to PM and adverse pregnancy outcomes—low birth weight and preterm birth.
The Volume 11 committee also concludes that there is limited/suggestive evidence of an association between exposure to PM during pregnancy and pregnancy-induced hypertensive disorders.
The committee found the literature on respiratory outcomes and neurodevelopmental effects to be relatively robust and convincing of associations with prenatal exposure to PM. The literature on other outcomes in children did not show any clear or consistent associations with PM exposure.
Respiratory Outcomes The Volume 3 committee concluded that there was “inadequate/insufficient evidence to determine whether an association exists” between combustion products or fuels and developmental effects in children (IOM, 2005). EPA concluded that the evidence was “suggestive of a causal relationship between long-term exposures to PM2.5 and . . . developmental outcomes” based on associations between PM exposure and infant mortality. This conclusion was based on studies of respiratory conditions in children with supporting studies on lung development (EPA, 2009).
Sixteen recent studies examined the relationship between prenatal PM exposure and asthma and respiratory outcomes in childhood. There is a robust epidemiologic literature, including several large cohort studies from various locations, that supports an association between prenatal exposure to traffic-related air pollutants (including PM2.5) and an increased risk of childhood asthma (Clark et al., 2010; Hsu et al., 2015; Tetreault et al., 2016; Voynow and Auten, 2015). Associations with other allergic diseases (atopic dermatitis and allergic rhinitis) (Deng et al., 2016; Huang et al., 2015; Liu et al., 2016b; Lu et al., 2017) and lung function growth (Eenhuizen et al., 2013; Jedrychowski et al., 2010a; Latzin et al., 2009; Morales et al., 2015; Mortimer et al., 2008) have also been reported. Interactive effects with combined prenatal exposure to community violence (Chiu et al., 2014) and traffic-related air pollution have been reported.
EPA (2009) found associations between PM2.5 and increased infant mortality (<1 year) especially due to respiratory diseases that occur during the neonatal period. EPA emphasized that the susceptibility of the developing lung from embryogenesis into adult life allows for a long period of vulnerability to stressors, including exposures to air pollutants such as PM. Because biological detoxification systems are not fully developed until after birth, the disruption of cell signaling during development can have long-term effects on lung growth.
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between prenatal exposure to PM and respiratory effects in children.
Neurodevelopmental Outcomes A variety of neurodevelopmental, cognitive, and behavioral deficits and ASD up to age 12 have been associated with prenatal PM exposure. Few of these studies, however, controlled for postnatal exposures that contribute to neurobehavioral changes.
Eight studies of ASD and PM were reviewed. Several studies found that ASD was associated with prenatal PM2.5 (Becerra et al., 2013; Raz et al., 2015; Volk et al., 2013) and prenatal PM10 exposure (Kalkbrenner et al., 2015), although trimester-specific estimates of risk were inconsistent, and several studies showed no such effects (Gong et al., 2014, 2017; Guxens et al., 2016). A meta-analysis reported increased risks of ASD with PM2.5 exposure in the first trimester and the postnatal period. None of the studies that examined preconception exposure found any association between PM2.5 or PM10 and ASD (Kalkbrenner et al., 2015; Raz et al., 2015; Talbott et al., 2015).
While two studies reported significant associations between prenatal PM exposure and cognitive or psychomotor development in children up to 2 years old (Kim et al., 2014; Lertxundi et al., 2015), these results were not corroborated in other populations of children (Guxens et al., 2014; Lin et al., 2014). In children ages 5 to 6 years, attention, IQ, and memory problems were associated with prenatal exposure to PM, but there were clear gender differences (Chiu et al., 2016; Yorifuji et al., 2016). One study found
that problems with executive functioning and behavior were associated with childhood PM exposure but not prenatal exposure (Harris et al., 2016).
Based on the positive results from several large studies and some well-conducted studies of ASD (Becerra et al., 2013; Raz et al., 2015; Volk et al., 2013) on consistent decrements in measures of neurodevelopment (although measures differed between studies), and on support from animal studies and mechanistic data that identify in utero exposure as a particularly susceptible window of development for neurodevelopmental effects associated with air pollutants and PM (Allen et al., 2017; Block et al., 2012; EPA, 2009), the committee had concerns about possible neurodevelopmental effects. The committee identified neurobehavioral effects in children of veterans as a priority area for future research, as discussed in later chapters.
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between prenatal exposure to PM and neurodevelopmental effects.
Other Developmental Effects Twenty-one studies have examined the association of birth defects with PM exposure. Most are population-based case-control studies using registries to assess a list of congenital defects and modeling exposures based on maternal residence at time of birth to represent prenatal exposure. The majority of the studies reported equivocal results, although differences between the studies (i.e., in the specificity of outcomes examined and the time frame of the exposures) makes this hard to interpret. Two meta-analyses that examined PM10 exposure and congenital anomalies did not find increased risk. Although the majority of studies that assessed the risks of congenital heart defects did not find any associations with prenatal PM2.5 or PM10, a few studies did. For PM10, associations were found for ventricular septal defects and atrial septal defects (Farhi et al., 2014) as well as multiple congenital heart defects and exposure in the first 3–8 weeks of pregnancy (Agay-Shay et al., 2013). For PM2.5, there were associations with dextro-transposition of the great arteries (Padula et al., 2013b), pulmonary valve stenosis, and tetralogy of Fallot (Warren et al., 2016). Two studies found an increased risk of cleft palate associated with PM2.5 exposure (Zhou et al., 2017; Zhu et al., 2015b), whereas 10 studies did not find increased risk of orofacial clefts with PM exposure. No increased risk of neural tube defects was reported in 7 studies. There were significantly increased risks of chromosomal defects (Farhi et al., 2014), abdominal wall defects (Schembari et al., 2014), omphalocele (Dolk et al., 2010), esophageal atresia (Padula et al., 2013c), and nervous system anomalies (Rankin et al., 2009) found in single studies.
Two studies reported analyses that included preconception exposures to explain the risk of orofacial clefts. Zhu et al. (2015b) observed a significantly increased risk of cleft palate with increased PM10 exposure through the 3 months before conception. Yao et al. (2016) found no association between PM2.5 exposure in the 3 months before conception and cleft palate or cleft lip.
Childhood cancers were not associated with PM in two studies (Badaloni et al., 2013; Heck et al., 2013b), but a third study found a significantly elevated risk of astrocytoma, although no other cancer, in children in Ontario (Lavigne et al., 2017).
Several other outcomes were reported in the literature, but none of them had enough data on which to base any conclusions.
A few studies examined immune effects in children following prenatal PM exposure, but no associations with childhood eczema or dermatomyositis were found (Jedrychowski et al., 2011a,b; Lu et al., 2017; Orione et al., 2014). Associations between prenatal exposure to air pollution and abnormal immunological parameters in neonates have been reported (Ashley-Martin et al., 2016; Baiz et al., 2011; Herr et al., 2010, 2011); however, the long-term significance of these effects is not known. One study found that higher exposures to NO2 and PM10 during pregnancy were both associated with increased
maternal and fetal cord blood levels of C-reactive protein, which is indicative of inflammation, but the significance of this association is unclear (van den Hooven et al., 2012a).
There is some evidence of an association between prenatal exposure to air pollution and cardiovascular and metabolic dysfunction later in childhood (Breton et al., 2016; Poursafa et al., 2016; Thiering et al., 2013; van Rossem et al., 2015).
Although some animal studies support the associations observed in epidemiologic studies, a lack of mechanistic understanding remains. The induction of oxidative stress, of DNA damage, and of the alteration of molecular signaling or epigenetic events have all been proposed as possible mechanisms (Saenen et al., 2015, 2016; Teng et al., 2016). Several studies support an effect of exposure to air pollutants (e.g., NO2 or PM10) and damage to fetal mitochondrial DNA from oxidative stress (Clemente et al., 2016; Grevendonk et al., 2016; Janssen et al., 2012). In a Boston birth cohort, exposure to relatively low levels of PM2.5 was associated with intrauterine inflammation (Nachman et al., 2016). Preexisting asthma and gestational diabetes in mothers may be effect modifiers of the association between air pollution and adverse birth outcomes (Erickson et al., 2016; Mendola et al., 2016).
Although some evidence has associated epigenetic changes with PM exposure, the impact of those changes on the health of children or the potential for such changes to be inherited by subsequent generations to cause generational health effects has not been demonstrated in either human or animal studies.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between prenatal exposure to PM and other developmental effects.
Polycyclic Aromatic Hydrocarbons
PAHs are chemicals formed as a result of the incomplete burning of organic substances (such as coal, oil, gasoline, diesel and jet fuel, wood, garbage, and tobacco). There are more than 100 different PAHs, and they generally are found as components of complex mixtures of chemicals. Some PAHs are used in medicines and to make dyes, plastics, and pesticides; others are in asphalt, and they are present in such substances as crude oil, coal, coal tar pitch, creosote, and roofing tar (ATSDR, 1995). Certain PAHs are carcinogenic and can cause other health effects (IOM, 2011). Benzo[a]pyrene (BaP) has been well studied in both epidemiologic studies and in the laboratory. It has been characterized as both a female reproductive toxicant and a developmental toxicant by EPA and the European Chemicals Agency in recent reviews (ECHA, 2017; EPA, 2017). BaP has been used as a sentinel of PAH exposure in some epidemiology studies (Guo et al., 2012; Herbstman et al., 2012; Perera et al., 2005a,b, 2007, 2012, 2014, 2015; Tang et al., 2006, 2012; Wu et al., 2010; Yin et al., 2017).
Gulf War personnel were exposed to PAHs as a result of the burning of over 600 oil wells in Kuwait. However, it was determined at that time that, based on the available air monitoring data, PAH exposure in soldiers deployed to Kuwait was relatively low to moderate when compared to the U.S. reference population. In addition, data from 61 soldiers indicated that the soldiers actually had lower concentrations of PAH-DNA adducts measured while in Kuwait than they had post-deployment, after spending a month at their base in Germany (Poirier et al., 1998). Post-9/11 personnel may have been exposed to PAHs as a result of burn pit use. PAH levels on military bases have been monitored as part of efforts to investigate the health hazards of burn pits (IOM, 2011). At JBB, a U.S. airbase in Iraq with a large burn pit, air sampling showed that vehicle and aircraft emissions were the primary sources of PAHs (Masiol
et al., 2016a,b). In 2007, the average total PAH concentration across JBB was 379 ng/m3, with the average concentration of BaP being 1.4 ng/m3 (average total BaP-toxic equivalent=2.9 ng/m3) (Masiol et al., 2016a). For comparison, total PAH and BaP concentrations from four of the cohorts discussed below are shown in Table 6-5.
Because PAHs were not specifically examined in any of the prior Gulf War and Health or other IOM volumes, the Volume 11 committee used the 1995 ATSDR toxicological profile on PAHs as the starting point for the literature to be considered in this report—that is, literature published after 1994. The committee identified more than 140 papers that report on reproductive, developmental, or generational effects in association with exposure to PAHs. Most studies investigated environmental exposures to PAHs. There were multiple publications for many cohorts—for example, there were 12 publications from the Columbia Center for Childrens’ Health (CCCEH) study in New York City (Bocskay et al., 2005; Choi et al., 2008; Herbstman et al., 2012; Perera et al., 2009a,b, 2012, 2014; Peterson et al., 2015; Rundle et al., 2012; Tang et al., 2012; Vishnevetsky et al., 2015). The committee reviewed three publications on occupational exposure to PAHs among Taiwanese coke oven workers (Hsu et al., 2006; Jeng et al., 2013a,b). Recent epidemiologic studies describing the reproductive and developmental effects of PAHs reviewed by the committee are summarized in Table 6-6 at the end of this section.
None of the studies reported on preconception exposures except for analyses of the National Birth Defects Prevention Study. Some studies used personal air monitoring during pregnancy (such as the CCCEH-run studies in New York and Poland). Many more estimated ambient PAH levels by modeling residential proximity to specific PAH sources (e.g., the World Trade Center disaster) or to ambient air monitoring stations. Others measured PAHs, PAH metabolites, or PAH-DNA adducts in blood, serum, urine, or placental tissue. For most studies of the effects seen in children after birth, the exposures were determined for the prenatal period and assumed to be consistent or representative of exposures through childhood. A few studies conducted follow-ups of exposure during childhood (Jedrychowski et al., 2014, 2015a,b,c; Perera et al., 2014). The only studies of ambient exposure to consider parental occupation were reported from the National Birth Defects Prevention Study (Langlois et al., 2012, 2013; Lupo et al., 2012a,b; O’Brien et al., 2016).
TABLE 6-5 Examples of PAH and BaP Exposure Levels (ng/m3) in Ambient Air in Cohort Studies
|Cohort||Mean Level (SD)||Range||Study|
|California Cancer Registry||PAH=1.48 (1.13)
|Heck et al. (2013a)|
|Columbia Center for Children’s Environmental Health Study (New York)||PAH=2.39 (2.12)||0.27–36.47||Rundle et al. (2012)|
|Choi et al. (2008)|
|Krakow, Poland||PAH=20.2 (95% CI 17.6–23.1)||Jedrychowski et al. (2015b)|
|Tongliang, China||BaP in 2002=13.14
BaP in 2005=9.14
|Tang et al. (2014)|
Reproductive Effects in Men
Several studies assessed the effects of PAHs on male reproduction. Many of these studies were conducted in China and Taiwan, with a handful conducted in populations in Europe and the United States. Many of the studies presented below measured the presence of urinary metabolites of PAHs, particularly 1-hydroxypyrene (1-OHP), a metabolite of pyrene. 1-OHP has been consistently detected in urine and has been used as a biological marker for assessing an internal dose of activated PAHs (Jeng et al., 2013a). Other urinary metabolites of PAHs include 1-hydroxynaphthalene (1-OHNa), 2-hydroxynaph-thalene (2-OHNa), 2-hydroxyfluorene (2-OHF) (Xia et al., 2009), 9-hydroxyphenanthrene (9-OHPh), 4-hydroxyphenanthrene (4-OHPh), and 9-hydroxyfluorene (9-OHFlu) (Yang et al., 2017b).
The three publications on exposure to PAHs for a cohort of 65 coke oven workers in Taiwan (Hsu et al., 2006; Jeng et al., 2013a,b) measured effects on sperm quality. Workers wore personal monitors to assess PAH exposure; 1-OHP was measured in urine after shifts (Jeng et al., 2013a). Men who smoked and had jobs that involved working close to the coke ovens, which are sources of PAH exposure, had reduced sperm quality and DNA damage (Hsu et al., 2006; Jeng et al., 2013a), but PAHs and 1-OHP did not affect sperm concentration or count or DNA fragmentation. The three papers reported conflicting results regarding the associations between ambient PAHs and sperm normal morphology, motility, and vitality and between 1-OHP and sperm motility. 1-OHP was negatively associated with normal morphology in all three publications. 1-OHP and some PAH species were significantly correlated with DNA adducts (Jeng et al., 2013b), and 1-OHP was significantly associated with measures of chromatin structure and mitochondrial membrane potential, although PAHs were not (Hsu et al., 2006).
Of the studies that examined semen parameters and sperm DNA integrity, most measured exposure based on urinary PAH metabolites (Gu et al., 2010; Han et al., 2011; Jurewicz et al., 2013b; Radwan et al., 2016; Xia et al., 2009; Yang et al., 2017b). Jurewicz et al. (2013b) and Xia et al. (2009) investigated the relationship between urinary PAH metabolites and sperm quality. Jurewicz et al. (2013b) reported that sperm neck abnormalities, volume, motility, and movement were significantly associated with increased urinary 1-OHP in 277 Polish men attending an infertility clinic; no associations were observed between 1-OHP and the DNA fragmentation index (DFI), sperm head or tail morphological abnormalities, semen concentration, the percent of atypical sperm, or measures of motility. Xia et al. (2009) compared levels of urinary PAH metabolites among 222 infertile men with abnormal semen quality, 291 infertile men with normal semen quality, and 273 fertile men in Nanjing, China, in a case-control study. Men in the highest tertile of exposure were more likely to have abnormal sperm than those in the lowest tertile (1-OHP OR=1.49, 95% CI 1.03–2.15; 2-OHF OR=1.52, 95% CI 1.06–2.17; and all metabolites OR=1.53, 95% CI 1.06–2.20). In 405 men in Wuhan, China, Yang et al. (2017b) reported no associations between urinary PAH metabolites and tail % or tail distributed moment, and only 9-OHFlu was associated with increased tail length and comet length (mean differences first versus third tertiles: 8.65%, 95% CI 2.53–15.03%, and 7.14%, 95% CI 2.33–12.19%, respectively). A significant association was observed between 9-OHPh and decreased Annexin V–/PI– spermatozoa (mean difference first versus third tertiles –5.38%, 95% CI –8.92– –1.86), but there were no other associations between metabolites and sperm apoptosis parameters. These data indicate that exposure to PAH may result in decreased sperm function.
Han et al. (2011), Gu et al. (2010), and Radwan et al. (2016) examined the association between effects on sperm DNA and urinary PAH metabolites. Han et al. (2011) noted increased DNA damage in sperm, but no effect on semen parameters in a cross-sectional study of 232 men in Chongqing, China. Total PAH metabolites were associated with the tail % (β=15.96, 95% CI 8.86–23.07), but not with the tail length or tail distributed movement. Significant associations were also reported between 2-OHNa and tail %, tail
length, and tail distributed moment (ORs=13.26, 12.25, 7.55, all p<0.05). Gu et al. (2010) reported the joint effects of PAH exposure and specific polymorphisms on sperm DNA integrity among 620 infertile couples in Nanjing, China, in a nested case-control study; they found that men with high levels of 1-OHP and certain DNA nucleotide excision repair genotypes (XPA and XPD) were at higher risk of sperm DNA damage. In a Polish cross-sectional study of 181 men with normal semen attending an infertility clinic, urinary 1-OHP levels were associated with sperm aneuploidy, specifically increased risks of total sex-chromosome disomy (β=1.46, 95% CI 1.05–2.03) and disomy 18 (β=1.22, 95% CI 1.02–1.63), but not other kinds of disomy for sex chromosomes or chromosomes 13 and 18 (Radwan et al., 2016).
Five additional studies reported on semen quality (Conti et al., 2017; Gaspari et al., 2003; Rubes et al., 2005; Song et al., 2013; Yang et al., 2017a). In a group of 182 Italian men, PAH adducts in sperm were found to be correlated with physiologic forms (r= –0.18, p=0.016; terminology undefined), abnormal head morphology (r=0.30, p=0.0001), and abnormal neck morphology (r= –0.21, p=0.009), but not with sperm count or tail morphology. PAH adduct levels were higher among those with any abnormality (p=0.04) (but not among men with asthenospermia, oligospermia, or teratospermia) and among infertile men than among fertile men (with fertility based on the pregnancy status of partners) at a 1-year followup (p=0.04), suggesting that PAH adducts might be a marker of infertility (Gaspari et al., 2003). Another study of 86 Italian men compared BaP–DNA adducts in sperm cells to semen parameters and found that BaP–DNA adducts were inversely associated with sperm motility (Conti et al., 2017). Among 53 men attending a fertility clinic in Guangzhou, China, Song et al. (2013) found that total PAHs in blood correlated with semen volume, concentration, progressive motility, and nonlinear motility (r= –0.07, 0.10, 0.07, –0.10, all p<0.05) but were not associated with motility. In the only study that reported ambient air PAHs, a study that took place in the Czech Republic, exposure in the 90 days before sample collection was not associated with sperm count or concentration, volume, motility, morphology, velocity, linearity, or aneuploidy; only sperm chromatin abnormalities (%DFI) were related to PAH exposure (β=0.19, 95% CI 0.02–0.36) among 36 men who were followed over 2 years (Rubes et al., 2005). A group of 933 men visiting an infertility clinic in Wuhan, China, showed significant inverse associations between urinary PAH metabolites and sperm parameters (Yang et al., 2017a). Among the 10 metabolites analyzed, significant decreases in sperm concentration (–22.66%, 95% CI –34.49– –8.70%), sperm count (–19.99, 95% CI –33.83– –3.34%), and percentage of sperm with normal morphology (–2.35, 95% CI –4.24– –0.46%) were found between the first and the fourth quartiles of exposure to 1-OHNa. Adverse associations were also seen for semen volume and exposure to 4-OHPh (–0.31 mL, 95% CI –0.61––0.01) and 9-OHPh (–0.43 mL, 95% CI –0.74– –0.12). Decreased motility was associated with 9-OHPh exposure (straight-line velocity –1.34 μm/sec, 95% CI –2.37– –0.31; curvilinear velocity –2.30 μm/sec, 95% CI –4.06– –0.54). No associations were noted between individual metabolites and sperm motility.
Han et al. (2010) examined the effect of PAH exposure on reproductive hormones in 562 men being evaluated for unexplained male factor infertility in Nanjing, China. An analysis by tertiles of metabolites and trisections of hormones showed both increased risks of higher hormone levels and significant trends between the following hormone–metabolite pairs: luteinizing hormone (LH) and both 1-naphthol (1-N) and 2-OHF (p-trend=0.025 and 0.019); follicle-stimulating hormone (FSH) and 2-OHF (p-trend=0.027); and estradiol and both 2-OHF and 1-OHP (p-trend=0.047 and 0.025). An analysis conducted using continuous variables showed only one significant relationship—between LH and 1-OHP (p=0.01).
Reproductive Effects in Women
Three studies examined the reproductive effects of PAHs in women. Slama et al. (2013) found that PAH exposure in the 2 months prior to the start of unprotected intercourse did not affect fecundability in
a cohort of 1,916 Czech births between 1994 and 1999. In a prospective cohort study of 51 fertile women in California, Luderer et al. (2017a) found that higher PAH levels, as assessed by urinary concentrations of nine different PAHs, were associated with changes in follicular phase length, follicular phase LH, estrone 3-glucuronide (E13G) levels, pre-ovulatory LH surge concentrations, and peri-ovulatory E13G slope and concentration. In a separate case-control study conducted in China among 80 women (50 with polycystic ovary syndrome [PCOS] and 30 without this condition), individuals with PCOS had higher concentrations of serum PAHs and their metabolites. Total serum PAHs were associated with a significant risk of PCOS with PAH exposure (aOR=4.04, 95% CI 1.16–14.0). Higher levels of certain metabolites, hydroxyl-phenanthrenes (aOR=2.34, 95% CI 0.73–7.49) and hydroxyl-biphenyls (aOR=2.01, 95% CI 0.64–6.35), were associated with a twofold increased risk of PCOS; however, these findings did not reach statistical significance. This study was limited due to its sample size and power limitations (Yang et al., 2015).
Adverse Pregnancy Outcomes
Although many studies have examined the effects of air pollution and air quality on birth outcomes, such as preterm birth and birth weight, few have specifically assessed PAHs. The committee’s literature search identified eight publications on birth size, two studies on preterm birth, one study of missed abortion (embryo died but miscarriage has not occurred in early pregnancy), and one study of reproductive hormones in umbilical cord serum.
Perera et al. (2005b) did not find any association between BaP adducts in maternal or cord blood—a proxy for prenatal PAH exposure—and birth weight, length, or head circumference among 186 women living near the World Trade Center in New York City at the time of the 9/11 disaster and delivering within 41 weeks of September 11, 2001. However, there was a significant interaction between cord blood adduct levels and environmental tobacco smoke exposure, with significantly lower birth weight and head circumference. Conversely, in the CCCEH cohort of pregnant women in New York City, total PAHs measured by personal monitor in the third trimester were significantly associated with increased risks of small for gestational age (OR=2.43, 95% CI 1.05–5.62, per 1 ln PAH unit increase), preterm births (OR=4.68, 95% CI 1.84–11.89), and reduced fetal growth ratio (<85%; OR=1.93, 95% CI 1.04–3.56) among African American mothers (n=224) but not for Dominican mothers (n=392) (Choi et al., 2008).
In a Chinese cohort of 183 births, Guo et al. (2012) did not observe any adverse effects of prenatal PAH exposure on birth height, weight, gestational age, Apgar score, or BMI, although some specific PAHs were significantly and negatively associated with height. A study of another Chinese cohort, this one consisting of 106 pregnant women in Wuhan, China, found that certain urinary metabolites measured in the third trimester were associated with decreased birth length, but none of the metabolites measured were associated with birth weight or gestational age. In a comparison of the first and third tertiles of exposure, 2-OHNa was found to be associated with a 0.80% reduction (95% CI –1.39– –0.20%) in mean birth length; neither Alu nor LINE-1 methylation (used as markers of global methylation) correlated with this association (Yang et al., 2018). Associations between higher PAH levels in the first gestational month and small for gestational age were seen in a Czech Republic study of 3,378 birth, suggesting that early exposure to PAH conferred a modest 20% increase in the odds of intrauterine growth restriction (aOR=1.22, 95% CI 1.07–1.39, per 10 ng increase in ambient PAH exposure) (Dejmek et al., 2000).
In a cohort of 344 pregnant Polish women, PAHs assessed by personal monitors in each trimester of pregnancy were associated with markers of growth restriction in the first trimester (fetal growth ratio= –3%, 95% CI –5%–0%, per 1 ln change in PAH exposure) (Choi et al., 2012) and with reduced birth weight and height for first and third trimesters (Wang and Choi, 2014). In a different cohort of 449
Polish women, urinary 1-OHP—used to represent PAH exposure—measured at 20–24 weeks of gestation was not related to birth weight, cephalization index (head circumference to body weight ratio), chest circumference, length, head circumference, or ponderal index (fetal weight to height). Only a reduction in chest circumference associated with 1-HP remained significant in the fully adjusted model, which accounted for salivary cotinine levels (Polanska et al., 2010, 2014). The authors reported a modest association between the interaction of hydroxyphenanthrene and cotinine levels on the cephalization index (p=0.04) (Polanska et al., 2014). Some studies have reported a significant interaction effect of PAH exposure and environmental tobacco smoke (Perera et al., 2005b; Polanska et al., 2014). In the same Polish birth cohort (n=455), personal monitoring for PAHs and PM was conducted in the second trimester to examine the impact of PM2.5 and PAHs together. Although PM2.5 alone was significantly associated with birth weight, birth length, and head circumferences, it was not significantly associated with these outcomes with co-pollutant models with PAH exposure. PAH exposure was strongly associated with all three measures on its own, but in the co-pollutant model PAH exposure was significantly associated only with birth weight (β= –0.20, p=0.004) and birth length (β= –0.17, p=0.025), not with head circumference (β= –0.13, p=0.086). The authors interpreted the results to indicate that PAH exposure is more biologically relevant for fetal development than PM exposure (Jedyrchowski et al., 2017). The Volume 11 committee notes that metabolic interactions between PM and PAHs were not considered in this study, but it is possible that they play some role in the observed health effects.
A Japanese study of 149 mother–child pairs found 1-OHP to be related to reductions in birth weight (β= –0.15, p=0.053), birth length (β= –0.18, p=0.022), and head circumference (β= –0.17, p=0.052), but the findings were not significant after smokers had been excluded (Suzuki et al., 2010). A cross-sectional study of 1,578 Saudi births found associations between PAH concentrations in cord blood and birth outcomes. BaP was associated with placental thickness (β= –0.071, p=0.017) and cord length (β= –0.075, p=0.012), and a summary measure of 4 PAHs was associated with decreased cord length (β= –0.077, p=0.003). Short cord length has been used as an indicator of fetal movement and developmental abnormalities. However, neither BaP nor the sum of 4 PAHs nor 1-OHP was associated with numerous other anthropometric birth outcome measures (Al-Saleh et al., 2013). This study also found an association between 1-OHP in cord blood and a marker of oxidative stress, 8-OHdG levels (β=0.303, p<0.001), which suggests that oxidative stress is a potential mechanism by which PAHs may produce adverse pregnancy outcomes. In the Generation R study, a population-based prospective cohort study of 4,680 births in the Netherlands between 2002 and 2006, investigators found inverse associations between maternal occupational exposure to PAH and fetal growth measures, with significantly lower associations between prenatal PAH exposure and relative change in fetal weight across pregnancy (β= –0.0166, SE=0.0080, p<0.05) (Snijder et al., 2012).
Langlois et al. (2014) reported on small for gestational age among births selected as controls from a series of case-control studies using the National Birth Defect Prevention Study (2,803 births). Parental occupational PAH exposure, as determined by an industrial hygienist, from the month before pregnancy through the first trimester had a significant effect on small for gestational age, although the result was based on only 17 small-for-gestational-age births with PAH exposure (OR=2.2, 95% CI 1.3–3.8).
In addition to fetal growth measures, researchers have assessed the effects of PAH exposure on preterm births. A study conducted in California of 42,904 mother–child pairs living within range of an EPA air monitoring site in Fresno, California, between 2001 and 2006, evaluated preterm births (Padula et al., 2014). This study used geocoding and modeling techniques to estimate the exposure to individual PAHs and the sum of PAHs. The model explained 81% of the between-residence variability and 18% of the within-residence variability, with the main sources of these PAHs in this area being roadways, agriculture, and residential fireplaces, stoves, and heating. In this study, PAH exposures during the last 6 weeks of pregnancy were found to be associated with a nearly threefold increase in early preterm birth (i.e.,
delivery at 20–27 weeks gestation) (OR=2.74, 95% CI 2.24–3.34) when comparing the highest quartile to the lower three quartiles. Quartiles of PAH exposure were also significantly related to preterm births, with a dose–response pattern (compared to lowest exposure: 2nd quartile OR=1.49, 95% CI 1.08–2.06; 3rd quartile OR=2.63, 95% CI 1.93–3.59; 4th quartile OR=3.94, 95% CI 3.03–5.12). The associations were similar across gradations of SES, but the associations were affected by the season of conception, with the strongest associations seen for pregnancies conceived during the summer (June–August) and inverse associations observed for pregnancies conceived in March–May (Padula et al., 2014). The authors were unable to determine the cause of these inconsistencies.
A case-control study of 29 preterm births and 31 full-term births conducted in India between 2005 and 2006 found that placental PAH levels were significantly higher in samples taken from women who had experienced preterm births than from those with full-term births, with significantly higher concentrations of fluoranthene and benzo(b)fluroanthene (both p<0.05). When considering the distribution of the PAHs in the placental tissue, the three-ringed PAHs made up more of the mixture in the preterm birth group (41%) than in the control group (31%) (Singh et al., 2008). Although this was a small study of only 60 women with limited power, it provides some suggestive evidence that PAHs may affect gestation length.
Missed abortion was studied by Wu et al. (2010) in a study of 81 cases and 81 controls in China; the researchers found no association between missed abortion and BaP–DNA adducts in fetal tissue. However, higher levels of BaP–DNA adducts in maternal blood (per 108 nucleotides) increased the risk of missed abortion (OR=1.35, 95% CI 1.11–1.64). Although adduct levels in maternal blood and fetal tissue were not correlated, maternal-blood BaP–DNA adduct levels did correlate with a survey on PAH exposure from common sources; 26% of the variance in missed abortion was explained by maternal blood BaP–DNA adduct levels. Also associated with missed abortion were other self-reported proxies for PAH exposure: residence near traffic congestion (OR=3.07, 95% CI 1.31–7.16), commute by walking (OR=3.52, 95% CI 1.44–8.57), and routinely cooking during pregnancy (OR=3.78, 95% CI 1.11–12.87).
Another study conducted in the Shendsi Islands of China cross-sectionally evaluated the association between umbilical-cord-blood PAH levels and umbilical-cord-blood hormone levels in mothers among 109 births. An inverse association was found between PAH levels and estradiol and anti-Müllerian hormone (AMH), and a positive association was found between PAH levels and FSH in women at the time of delivery (Yin et al., 2017). No significant associations were reported between testosterone, luteotropic hormone (prolactin), or estradiol and any specific PAH species or between acenaphthylene, benzo(b)fluoranthene (BbF), benzo(k)fluoranthene (BkF), or BaP and any hormone. Several significant positive associations were reported between FSH and fluorine, phenanthrene, anthracene, fluoranthene, benz(a)anthracene, and chrysene (β=0.47, 0.44, 0.51, 0.38, 0.42, 0.35, all p<0.05), and significant negative associations were reported between AMH and naphthalene, acenaphthene, fluorine, and fluoranthene (β= –0.33, –0.33, –0.44, –0.35, all p<0.05) after adjusting for confounders. Because a number of childhood and later life conditions in offspring are theorized to be associated with hormonal changes in pregnancy, this finding may provide a pathway by which higher concentrations of PAH could operate through a hormonally mediated mechanisms to affect pregnancy loss as well as pregnancy complications and fetal outcomes (Yin et al., 2017).
Developmental outcomes assessed after birth—including childhood cancers, neurobehavioral outcomes, growth, and respiratory outcomes—are discussed here. Few studies assessed the same outcome or measure in different populations, making it difficult to assess consistency across studies. None of
these studies specifically assessed preconception exposures, and many of these studies assessed outcomes in children without consideration for the continuous nature of exposures throughout childhood—for example, investigating prenatal PAH exposure with an outcome at 9 years of age without factoring in PAH exposures between birth and age 9.
As with pregnancy complications, birth defects may be the downstream effects of exposures to PAHs in utero. Several epidemiologic studies have evaluated the associations between PAHs and birth defects. These studies are mostly case-control study designs using data from national or regional registries. Most studies were conducted in China or the United States. The U.S. studies were primarily analyses based on the NBDPS, a large population-based case-control study that assessed about 30 different birth defects in 10 states over 14 years (CDC, 2017). Data on birth defects and live births were collected as part of state surveillance systems in Arkansas, California, Georgia, Iowa, Massachusetts, New Jersey, New York, and Texas, with data including live births. Controls came from the same areas as the cases and were randomly selected by birth certificate or hospital birth records. Various birth defects were considered, such as neural tube defects, cleft palate, gastroschisis, craniosyntosis, and congenital heart defects. All the U.S. studies based their exposure information on maternal occupational exposures, with most of them evaluating exposure to PAHs on the basis of retrospective data collected on direct or indirect exposures as well as on the duration and intensity of exposures for each job. Preconception exposure was not distinguished from prenatal exposure. Only inhalation was considered, and skin and ingestion exposure routes were not accounted for in these studies. On the other hand, the Chinese studies mainly looked at the relationship between PAHs—examined in a variety of tissues, including maternal blood, cord blood, and placenta—and neural tube defects.
Analyses of the NBDPS have all used a case-control design with nearly 3,000 healthy controls. Significantly elevated risks were reported for cleft lip, with or without cleft palate (n=805; OR=1.47, 95% CI 1.02–2.12), but not cleft palate alone (n=439; OR=1.24, 95% CI 0.76–2.13) (Langlois et al., 2013). The risk of gastroschisis was also elevated (n=299; OR=1.75, 95% CI 1.05–2.92), specifically among mothers 20 years of age and older (OR=2.53, 95% CI 1.27–5.04) (Lupo et al., 2012a). No significant associations with congenital heart defects were observed (n=1,907), including for conotruncal defects, left ventricular outflow tract defects, right ventricular outflow tract defects, or septal defects (Lupo et al., 2012b). However, the risk of craniosynostosis was significantly elevated (OR=1.75, 95% CI 1.01–3.05), although only 16 of the 3,016 cases had reported PAH exposure (O’Brien et al., 2016). No significant risks of neural tube defects were observed (n=520) (Langlois et al., 2012).
Two separate case-control studies in Shanxi Province, China, reported on the relationship between neural tube defects and PAH exposure (Yi et al., 2015; Yuan et al., 2013). Among 80 cases and 50 controls born in 2005–2007, investigators found a significant inverse relationship between PAH adducts in placental tissue and neural tube defects, with the controls having significantly higher concentrations of PAHs than the cases. No significant differences were found for spina bifida or anencephaly subtypes. When stratified by PAHs and PAH–DNA adducts, significantly elevated risks remained only among those with high PAH exposure and low PAH adducts (OR=8.2, 95% CI 2.1–32.7) (Yuan et al., 2013). Among 60 cases and 60 controls born in 2010–2012, Yi et al. (2015) reported a significant dose–response relationship for neural tube defects, spina bifida, and anencephaly with tertiles of PAH–DNA adducts in umbilical cord tissue and blood (p-trend <0.05 for all). Risks at the highest tertiles of exposure were significantly increased for neural tube defects (OR=2.96, 95% 1.16–7.58), spina bifida (OR=2.97, 95% CI 1.18–7.49), and anencephaly (OR=3.12, 95% CI 1.14–8.55).
Several case-control studies (Heck et al., 2013a, 2014; Shrestha et al., 2014; von Ehrenstein et al., 2016) have used the California Cancer Registry to examine the risks of cancer among children younger than 6 years of age; data from birth certificates were used to identify population-based controls to match with the cases of children with cancer. Heck et al. (2013a) estimated exposure to PAHs for each trimester of pregnancy at the mother’s residence using data collected from nearby community-based air pollution monitors. There was no significant association of PAH levels with an increased risk of neuroblastoma for children within 5 km of a monitor (75 cases and 14,602 controls); however, those living within 2.5 km showed an increased risk of neuroblastoma associated with total PAHs (OR=1.39, 95% CI 1.01–1.91). Among children of mothers living within 5 miles of a monitor, there was no increase in the risk of primary neuroectodermal tumors (43 cases) or astrocytoma (106 cases), but the risk of medulloblastoma (in 34 cases versus 30,569 controls) was correlated with total PAH exposure (OR=1.44, 95% CI 1.15–1.80), particularly for benzo(k)fluoranthene, benzo(a)pyrene, benzo(b)fluoranthene, indeno(1,2,3-cd)pyrene, dibenzo(a,h)anthracene, and benzo(g,h,i)perylene (von Ehrenstein et al., 2016). Risk of Wilms’ tumor, a type of kidney cancer of embryonic origin, among 337 cases and 96,514 controls living within 15 miles of a monitor was associated with PAH exposures throughout pregnancy (OR=1.15, 95% CI 1.02–1.31), but the estimates were not significant for individual trimesters (Shrestha et al., 2014). For children living within 6 km of an air monitor, there was an increased, but not significant, risk of ALL (69 cases and 2,994 controls) and AML (46 cases and 19,209 controls) with exposure to PAHs throughout pregnancy. Third-trimester exposure to PAHs was associated with a modest increase in ALL (OR=1.16, 95% CI 1.04–1.29), including significantly increased associations with individual PAH species in the third trimester, with ORs ranging from 1.06 to 1.27. Several other exposures, including arsenic, 1,3-butadiene, and lead, also showed significant associations (Heck et al., 2014).
Using data from children diagnosed with retinoblastoma from a private medical practice in Philadelphia or at a U.S. or Canadian institution of the Children’s Oncology Group (more than 200 medical centers) between June 2006 and 2012, investigators examined 187 unilateral and 95 bilateral cases of retinoblastoma (Omidakhsh et al., 2018). Controls (n=155) were identified through cases by identifying friends or a relative of the case child who was less than 15 years of age. Exposure data were assessed on the basis of paternal occupational exposure to PAHs in the 10 years before conception. Maternal occupational exposures were assessed 1 month before as well as during pregnancy. Increased, but not significant, associations were noted for unilateral (OR=1.34, 95% CI 0.48–3.72) and bilateral (OR=1.51, 95% CI 0.36–6.38) neuroblastoma, but the analyses were limited by the relatively few PAH-exposed cases (n=22) and controls (n=18). (Omidakhsh et al., 2018).
In a subset of 60 infants born to African American and Dominican women participating in the CCCEH cohort study, investigators measured chromosomal aberration frequencies in cord blood as markers of cancer risk. Prenatal exposures to PAHs measured in air were positively associated with stable chromosomal aberrations (r=0.35, p<0.01, and β=0.14, p=0.006), but not with PAH–DNA adducts. This finding suggests that PAHs have the potential to cause cytogenic damage related to an increased risk of cancer (Bocskay et al., 2005).
The associations between prenatal exposures to PAHs and childhood neurological conditions have been the subject of multiple investigations. Most studies come from two countries—the United States and Poland. A total of 11 epidemiological studies have been conducted, with 4 of the studies from Poland and 7 from the United States. Most of the U.S. studies are of the CCCEH cohort conducted in
the northern Manhattan section of New York City among Dominican and African American mother–child pairs. A range of neurological outcomes were assessed, including IQ and cognitive dysfunction as assessed by the WISC, behavioral development evaluated by a 118-item Child Behavior Checklist for children ages 6–9 years, and cognitive and psychomotor development evaluated by Bayley-II Scales of Infant Development (BSID-II). Other neurological conditions were assessed, including ADHD, as well as white matter function, cognition, and behavior in later childhood.
Three reviews sought to summarize the literature base regarding the impacts of exposure to environmental toxicants, including PAHs, on children’s neurodevelopment and behavior. Polanska et al. (2012) studied risk factors for ADHD in the literature published between 2000 and 2012 and cited one study on the risk of ADHD with PAH. Polanska et al. (2012) concluded that the ADHD risk associated with tobacco smoke, a source of PAH exposure, was consistent but that the evidence regarding environmental exposures, including PAHs, was insufficient to reach any conclusions about ADHD and environmental exposures, specifically PAHs. Regarding cognitive development and behavioral problems, Jurewicz et al. (2013a) cited eight studies and noted the possible adverse impact of PAHs on children’s neurodevelopment. Suades-Gonzalez et al. (2015) summarized the literature on air pollutants and neuropsychological development and found seven publications that reported on prenatal PAH exposure. The authors concluded that there was sufficient evidence to show an association between pre- or postnatal PAH exposure and decreased global IQ scores, but that there was inadequate evidence to support associations with academic skills or ADHD.
Neurobehavioral outcomes in children have been well studied in the CCCEH cohort study. This study measured environmental PAH exposure in mothers who wore personal monitors in the third trimester; determined PAH–DNA adducts in maternal and cord blood and in children’s urine at ages 3 and 5 years; and assessed neurodevelopmental indicators at age 7 (brain development, behavior, and IQ) (Perera et al., 2009b, 2012, 2014, 2015; Peterson et al., 2015; Vishnevetsky et al., 2015).
Two publications based on the CCCEH cohort examined measures of brain physiology and development. Brain development, as assessed by brain-derived neurotrophic factor in cord blood and the BSID-II Mental Development Index (MDI) scores at age 2 years, was negatively correlated with PAH–DNA adducts in cord blood (β= –0.11, p=0.02) and with MDI scores (β= –2.07, p=0.04). These associations were not observed in a model that included all three variables, at age 3, or when using BaP–DNA adducts to represent exposure (n=505; Perera et al., 2015). Prenatal PAH exposure was inversely correlated with white matter surface measures, especially in the left hemisphere, in 40 children assessed between ages 7 to 9, and high exposure was associated with reduced processing speed during intelligence testing (mediated by white matter disturbances). Such an effect could help explain the reductions in verbal IQ and information processing and the increases in attention and behavioral problems reported in other studies (Peterson et al., 2015).
Publications from the CCCEH study also reported differences in behavior. In children 6 or 7 years of age, ambient prenatal PAH exposures were associated with anxiety and depression scores (continuous exposure OR=1.2, 95% CI 1.06–1.37), attention problems (high/low OR=3.79, 95% CI 1.14–12.66), anxiety disorder (high/low OR=4.59, 95% CI 1.46–14.27), and ADHD (high/low OR=2.3, 95% CI 0.79–6.7), but none of the behaviors were significantly associated with maternal or cord blood PAH–DNA adduct measurements taken at the time of delivery (n=253) (Perera et al., 2012). DNA adducts in maternal blood were correlated with ADHD at age 9 in 250 children, after adjusting for prior childhood exposure (using urinary PAH metabolites at ages 3 and 5), with significantly increased risk for the DSM-IV inattentive subtype (OR=5.06, 95% CI 1.43–17.93). Adducts in cord blood were not related to ADHD (Perera et al., 2014).
Multiple publications from the CCCEH study reported on the relationship between PAH exposure and IQ measures performed using the Wechsler Preschool and Primary Scale of Intelligence-Revised.
Perera et al. (2009b) studied a group of 249 children and reported that high levels of prenatal PAHs were significantly associated at age 5 with IQ decrements on verbal performance (β= –3.53 points, p=0.002) and full-scale IQ scores (β= –3.0 points, p=0.009). Children exposed to high PAHs had IQ scores about 4 points lower than children with lower exposure (above versus below the median). Vishnevetsky et al. (2015) measured IQ at age 7 using the WISC and noted a significant interaction between material hardship (unmet basic needs in the areas of food, housing, and clothing) and PAH–DNA adducts on working memory (β= –8.07 points, 95% CI –14.48– –1.66), such that children with high prenatal hardship and high levels of PAH-DNA adducts had lower working memory scores (β= –6.67 points, 95% CI –11.38– –1.95). Children who experienced high prenatal hardship and had high PAH-DNA adducts also had significant decrements in full-scale and perceptual reasoning scores (β= –5.81 points, 95% CI –10.35– –1.26 and β= –5.44 points, 95% CI –10.27– –0.61, respectively), but there was no significant effect of the interaction term on full-scale IQ, verbal comprehension, processing speed, and perceptual reasoning scores in 276 children with low prenatal material hardship. In another analysis, PAH–DNA adducts measured in cord blood were found to be associated with depressed verbal IQ scores among 170 children at 7 years of age (RR=3.00, 95% CI 1.32–6.79) but not with other IQ scores. However, breastfeeding exclusively for at least 6 months had a protective effect on verbal IQ scores attributed to PAH exposure (RR=0.31, 95% CI 0.11–0.89) (Jedyrchowski et al., 2015a).
IQ and neurodevelopment were also studied in cohorts of children in New York, China, and Poland. Edwards et al. (2010) observed a cohort of 214 Polish children whose mothers were monitored for PAH exposure. Decrements in neurodevelopment at age 5 were associated with second-trimester exposure when comparing high to low exposure (above or below the median of 17.96 ng/m3) (β= –1.36 points, 95% CI –2.48– –0.23, p=0.02) and continuous exposure (β= –0.56 points, 95% CI –1.00– –0.11, p=0.02). Perera et al. (2007) assessed a cohort of 98 children born in New York City to mothers who resided within 1 kilometer of the World Trade Center and had exposure to PAHs during pregnancy as a result of the towers burning. In a multivariate analysis, BaP–DNA adducts in cord blood at the time of delivery were not associated with the BSID psychomotor development index or MDI scores at age 3 years (Perera et al., 2007).
Other studies of neurobehavioral development include those of cohorts of children from China and of two Polish cohorts. In a study of Chinese children, Tang et al. (2014a) found a significantly increased percentage of children who had developmental delays among 2-year-old children with higher PAH exposure (n=150; expressed as plasma BDNF=1,266.56 pg/mL) when compared with those with lower PAH exposure (n=158; expressed as plasma BDNF=752.87 pg/mL), such that a unit increase in cord blood PAH–DNA adducts was negatively associated with average (β=212.11, p=0.01), motor (β=210.70, p=0.05), and adaptive (β=216.47, p=0.02) Gesell Developmental Schedule developmental quotient scores. No differences in language delays were noted. In another study of 604 Polish children 1 or 2 years of age, Polanska et al. (2013) found that prenatal PAH exposure, determined by urinary 1-OHP collected between 20 and 24 weeks of pregnancy, was not associated with cognitive, language, or motor development, although exposure to environmental tobacco smoke had a significant effect on cognitive and motor development. In a 2015 study evaluating behavioral development in 151 children ages 6–9 years old who were born to mothers living in Krakow, Poland, data on PAH exposure were collected from personal air monitors worn for 48-hours during the second trimester of pregnancy. Associations varied based on the micronutrients (retinol, alpha (α)-tocopherol, gamma (γ)-tocopherol, and a composite variable for carotenoids) in maternal and cord blood. Higher PAH concentrations were associated with higher scores in anxious/depressed domains as well as with being withdrawn/depressed and having social problems, attention problems, and aggressive behaviors on the basis of scores from the Child Behavior Checklist. The strongest associations were seen for low retinol intake and higher PAH
levels with withdrawn/depressed behavior, with a 0.41 increase in this score for every unit increase in PAH levels (p=0.0008) (Genkinger et al., 2015).
Neurodevelopmental effects were significantly affected by factors such as breast-feeding (Jedyrchowski et al., 2015a), maternal hardship (Vishnevetsky et al., 2015), nutrition/antioxidant levels (Grenkinger et al., 2015), and environmental tobacco smoke (Polanska et al., 2013).
Given that a number of studies have suggested that prenatal PAH exposure relative to growth restriction and other measures of newborn and early infancy growth measures, it is plausible that prenatal exposure could affect later childhood measures of growth. The impacts of prenatal PAH exposure on growth in childhood were reported in three studies. A cohort of 150 children born in 2002 and a second cohort of 158 unexposed children born in 2005, both living near a power plant that ceased operation in 2004 in China, were studied for effects on growth at birth, 18, 24, and 30 months (Tang et al., 2006, 2014b). Exposure was estimated based on environmental sampling and PAH–DNA adducts in maternal and cord blood. Growth rates did not differ between the two cohorts, although the 2005 cohort had significantly lower BaP exposure (according to both environmental sampling and measures of DNA adducts, p<0.001). In the 2002 cohort, PAH–DNA adducts were significantly related to a smaller head circumference at birth and lower weight at 18, 24, and 30 months (Tang et al., 2014b).
Jedrychowski et al. (2015b) assessed the height through age 9 of 379 Polish children and compared that with prenatal PAH exposures (measured by personal monitors in the second trimester) and indoor and outdoor PAH exposure measured at age 3. Researchers found that a higher prenatal exposure to PAHs measured in the second trimester was associated with modest but significantly lower height gain between the ages of 3 and 9 years than in children who were not exposed (β= –1.07 cm, 95% CI –2.01––0.05); however, the decrements in height gain were mediated by shorter birth lengths that were also associated with prenatal PAH exposure.
Among the children of women residing in New York City who were assessed for PAH exposure during their third trimesters by personal ambient air monitors over a 48-hour period, prenatal PAH exposure was found to be associated with BMI at both 5 years (n=453) and 7 years (n=371) of age. At both ages, children in the second and third tertiles of exposure had increased risks of obesity compared with children in the first tertile of exposure (at age 5, PAH 2nd tertile [1.73-3.07 ng/m3] RR= 1.79, 95% CI 1.08–2.98 and PAH 3rd tertile [≥3.08 ng/m3] RR=1.79, 95% CI 1.09–2.96; at age 7, PAH 2nd tertile RR=2.25, 95% CI 1.27–4.01 and PAH 3rd tertile RR= 2.26, 95% CI 1.28–4.00). Furthermore, higher ambient prenatal PAH exposure was associated with higher fat mass (Rundle et al., 2012).
A series of publications on a cohort of Polish mothers and children examined the respiratory effects of prenatal PAH exposure measured by personal monitor in the second trimester. In the two studies that distinguished between prenatal and postnatal exposure, significant effects of postnatal exposure were reported (Jedrychowski et al., 2014, 2015c).
Prenatal PAH exposure was significantly associated with the number of respiratory symptom episodes in the first year of life, including ear infection (RR=1.82, 95% CI 1.03–3.23), cough (RR=1.27, 95% CI 1.07–1.52), cough without cold (RR=1.72, 95% CI 1.02–2.92), barking cough (RR=4.80, 95% CI 2.73–8.44), wheezing without cold (RR=3.83, 95% CI 1.18–12.43), and sore throat (RR=1.96, 95% CI 1.38–2.78). The durations of respiratory episodes were similar to the number of occurrences. In addition,
a significant interaction between prenatal PAH exposure and postnatal environmental tobacco smoke exposure smoking was observed for these associations (n=333) (Jedyrchowski et al., 2005). At age 4, the severity of wheezing was associated with prenatal PAH exposure (IRR=1.12, 95% CI 1.07–1.18) and also with PAH exposure measured at age 3 (IRR=1.08, 95% CI 1.02–1.14). Recurrent wheezing was associated with postnatal PAH exposure (OR=1.61, 95% CI 1.16–2.24) but not with prenatal exposure (OR=1.40, 95% CI 0.97–2.03) (n=257) (Jedyrchowski et al., 2014). Lung function (FEV05, FEV1, FEF25–75) measured between the ages of 5 and 9 years was significantly reduced among children who had been in the highest tertiles of prenatal PAH exposure. Postnatal indoor and outdoor PAH exposure measured at age 3 was also related to decreased lung function between the ages of 5 and 9 (n=195) (Jedrychowski, et al., 2015c). Based on PAH–DNA adducts measured in cord blood in this cohort, prenatal PAH exposure was significantly related to fractional exhaled nitric oxide (FeNO) (β=0.32, p=0.006) at 7 years old in a subset of 82 children. Atopy and maternal allergy were also significantly related to FeNO, but PAH exposure explained 11% of the variance in FeNO levels (all three variables together explained 21% of the variance). FeNO was used as an indicator of airway inflammation (Jedrychowski et al., 2012). One meta-analysis of six studies reporting on childhood wheezing and prenatal PAH exposure reported a nonsignificant increased association (OR=1.04, 95% CI 0.94–1.15) (Hehua et al., 2017).
One exploratory study evaluated the mechanisms involved in prenatal PAH exposure and the DNA methylation of the acyl-CoA synthetase long-chain family member (ACSL3) as it related to parental report of asthma symptoms. The ACSL3 gene is associated with fatty acid metabolism and is expressed in the lung (Perera et al., 2009a). As such, this exploratory study may provide an epigenetic mechanism by which PAH exposure could affect lung function.
There is a large literature that specifically examines BaP with rodent and in vitro studies. EPA’s 2017 Toxicologic Review of BaP stated: “Animal studies demonstrate that exposure to benzo[a]pyrene is associated with developmental (including developmental neurotoxicity), reproductive, and immunological effects” (EPA, 2017).
The EPA review found that animal models indicated that BaP exposure prior to mating and during pregnancy can decrease the ability of females to maintain pregnancy (including embryo/fetal resorptions, decreased litter size, and decreases in ovulation rate); one study found decreased litter size after a single exposure prior to mating (Mattison et al., 1980). Multigenerational studies of BaP exposure have shown effects on fertility and the development of reproductive organs (decreased ovary and testes weight, decreased sperm quality, the destruction of primordial follicles, decreased corpora lutea, the stimulation of oocyte apoptosis, decreased sensitivity to FSH-stimulated follicle growth) in offspring exposed in utero.
The Volume 11 committee identified several animal studies on BaP exposure that were not cited in EPA’s review. These studies are briefly summarized below.
In one study, alterations in male reproductive parameters were evaluated after repeated exposure to BaP (Jeng et al., 2015). Young, adult male Hsd: ICR (CD1) mice (n=8) were administered BaP by oral gavage at doses of 0, 1, 10, 50 or 100 mg/kg/day for 30 or 60 days. Body weight was not affected by the treatment. In mice treated for 60 days with 100 mg/kg/day, the weights of the testes, seminal vesicle, and epididymis were significantly less (p<0.05) than those of the controls. With 60 days of treatment, sperm motility and vitality were significantly decreased at 100 mg/kg/day, while sperm morphology was significantly altered at both 50 and 100 mg/kg/day. DNA strand breaks in spermatozoa were significantly increased at doses of ≥50 mg/kg/day for 30 days and ≥10 mg/kg/day for 60 days. Microscopic lesions of the testes and a loss of integrity of the seminiferous tubule and epithelium were observed at ≥50 mg/kg/
day. Another study of the effects of BaP exposure (25, 50, 100 mg/kg/day orally for 28 days) on male germ cells 42 days after the end of exposure found that there were twice as many mutations in dividing spermatagonia at the 50 and 100 mg/kg doses than in control mice (Rowan-Carrol et al., 2017).
Other PAHs have also been known to affect male reproductive function. For example, in mice exposed in utero, low-dose exposure to BbF was associated with altered sperm function, with more apoptotic cell death at higher concentrations (Kim et al., 2011). Other studies have shown associations between certain PAHs and effects on sperm parameters, including decreased motility, abnormal head morphology, vitality, and concentration of mature spermatozoa as well as alterations in DNA repair genes in animal models (Jeng et al., 2015).
Li et al. (2017) reported that BaP impaired endometrial decidualization and decidual angiogenesis (which is related to embryo implantation and maintenance of pregnancy) in pregnant mice. BaP (0, 0.2, 2, and 20 mg/kg/day by gavage on GD 1–8) was associated with several adverse reproductive effects, including smaller and more irregular uterine implantation sites; impaired response to artificial decidualization stimulus; decreased estradiol and progesterone steroid levels in serum and a reduced expression of estradiol and progesterone hormone receptors in the uteri; reduced levels of key molecules related to decidualization; and the disruption of progesterone and vascular endothelial growth factor signaling mechanisms.
One in vitro study found that BaP (50, 100 or 250 μM) was associated with meiotic failure and the impaired fertilization of oocytes from pigs by increasing reactive oxygen species and apoptosis (Miao et al., 2018).
A study of the ovarian effects of prenatal exposure to BaP (2 mg/kg/day orally on GD 6.5–15.5) on female offspring found that exposed female F1 mice reached puberty earlier than controls (5 days) and showed a depletion of ovarian follicles. Maternal glutamase cysteine ligase modifier subunit (Glcm) deficiency did not modify the ovarian effects of prenatal BaP exposure. Neither BaP exposure nor Glcm genotype were related to Kras mutations in F1 mice (Luderer et al., 2017b).
Increased mutation rates and mosaicism in the brain, bone marrow, liver, and sperm of mice exposed to BaP in utero (10, 20, or 40 mg/kg/day by gavage on GD 7–16) were reported by Meier et al. (2017). F1 mice also showed decreased sperm count and motility compared to controls.
Learning deficits following in utero exposure to BaP were assessed by McCallister et al. (2016). Pregnant Long Evans Hooded rats were administered 0, 150, 300, 600, or 1200 μg/kg/day on GD 14–17. Metabolites were measured in blood and cortex samples from pups on PNDs 0–15. Beginning on PND 40, selected offspring were tested in a spatial discrimination and reversal task using operant chambers. Treatment with BaP had no effect on maternal and offspring body weight or on reproductive parameters, and no clinical signs of toxicity were observed. Total metabolites showed a dose-dependent increase and a time-dependent decrease in offspring blood and brain tissue during lactation; the most common metabolites were -diol and -epoxide derivatives. Offspring from treated and control dams showed no differences in the total number of correct or incorrect lever presses or omissions throughout behavioral testing, exhibiting a similar ability to acquire original discrimination and reach criterion. However, on the first day of reversal, a dose-related decrease in correct responses and increase in incorrect responses was observed in offspring exposed to 600 and 1200 μg/kg/day. Response acquisition for the second reversal was not affected by treatment and was similar to the original discrimination for the treated groups.
In vitro studies have demonstrated the mechanistic plausibility of early embryo loss due to BaP. The migration and invasion of HTR-8/SVneo trophoblast cells were inhibited at BaP concentrations ≥1 μM (Liu et al., 2016a), suggesting an adverse effect on implantation. BaP caused cell cycle arrest in the JEG-3 human trophoblast cancer cell line but not in ex vivo human trophoblasts from term placentas (Wakx et al., 2016). An increase in reactive oxygen species and in the percentage of apoptotic cells was observed in mouse preimplantation embryos exposed to ≥5 nM BaP (Zhan et al., 2015). Another
in vitro study showed that benzo[a]pyrene-7,8-diol-9,10-epoxide, a metabolite of BaP, can adversely affect embryo implantation by the induction of apoptosis of trophoblast cells, which create the outer layer of the embryo and develop into part of the placenta in human cell lines (Wang et al., 2017d, 2018).
In an in vitro study using rat cells, researchers showed the importance of the timing of exposure on the sensitivity of neurodifferentiation, presumably resulting in neurodevelopmental effects and behavioral effects. The differentiation of neurotypic PC12 cells (already set to become neural cells) was more sensitive to BaP, but embryonic stem cells (not yet determined to become neural cells) were more sensitive to the PAH mixture (Slotkin et al., 2017).
Although the routes of exposure studied in animal studies are not always identical to the major routes of exposure relevant to human populations, it is important to note that these findings align with the findings from epidemiological studies.
Three CCCEH studies assessed the epigenetic effects of PAHs by measuring DNA methylation in cord blood (Herbstman et al., 2012; Perera et al., 2009a; Tang et al., 2012). Herbstman et al. (2012) analyzed umbilical cord white blood cells from 164 newborns and reported that decreased global methylation was associated with prenatal PAH exposure (β= –0.11, 95% CI –0.21–0.0), there were no associations between methylation and urinary metabolites, and increased global methylation was associated with BaP adducts in cord blood (OR=2.35, 95% CI 1.35–4.09). In an in vitro study, Tang et al. (2012) found that prenatal PAH exposure was associated with an increased methylation of the interferon gamma (IFNγ) promoter region in umbilical cord blood white blood cells from 53 CCCEH participants (low PAH 88.7% methylation versus high PAH 97.1% methylation, p<0.01). There was no association with IL4 promoter methylation. The IFNγ gene is related to asthma risk. An analysis by Perera et al. (2009a) examined the methylation of ACSL3 5′CGI because hypermethylation at this location and the degree of gene expression were strongly correlated with PAH exposure. Among 56 children in the CCCEH cohort, prenatal PAH exposure (above or below 2.41ng/m3) was associated with the methylation of ACSL3 5′CGI (OR=13.8, 95% CI 3.8–50.2) and methylation was associated with asthma diagnosed before the age of 5 years (OR=3.9, 95% CI 1.1–14.3).
In a cohort of 106 pregnant Chinese women, urinary PAH metabolites were found to be significantly associated with birth outcomes, although global methylation (as represented by LINE-1 and Alu repetitive elements) in cord DNA was not found to mediate the association between urinary PAHs and birth outcomes. Specifically, two PAH metabolites (2-OHNa and 1-OHPh) out of the 10 that were measured were associated with lower Alu methylation, with no significant relationships between PAH metabolites and LINE-1 methylation (Yang et al., 2018).
In a small case control study of 20 children with neural tube defects and 20 controls, PAH content in fetal neural tissue and maternal serum was positively correlated with methylation at two CpG sites (CTNNA1 and MYH2), with the correlation confirmed in a mouse model using in utero exposure to BaP. The results led researchers to conclude that the hypermethylation of those two genes is associated with an increased risk of neural tube defects (Wang et al., 2017c).
Godschalk et al. (2015) investigated BaP mutations in the germ lines of male mice that were deficient for nucleotide excision repair gene Xpc (Xpc–/–) as well as in the germ lines of their wild-type littermates. Males were bred 6 weeks after exposure ceased (13 mg/kg BaP, 3 days/week for 6 weeks by gavage). Although DNA adducts in the testes of F0 mice were unaffected, the litters fathered by treated Xpc–/– males were significantly smaller (p=0.014) than the controls, but the litter size of exposed wild-type males was only slightly smaller than that of controls, and the difference was not significant. Significant testicular
DNA hypomethylation was found in the wild-type treated males but not in the Xpc–/– treated males, compared with their respective controls. The global methylation of retrotransposons in liver samples of offspring was not affected by paternal treatment. The mutation rates of paternal origin were increased only in offspring from repair-deficient sires treated with BaP (0.056±0.020 versus 0.017±0.013 for controls; p=0.034), and there was no effect of exposure on mutation rates in wild-type males.
An animal study found changes in adult weight gain and increased adult adipocyte size in males and females in the F1 and F2 generations as well as increased expression of peroxisome proliferator-activated receptor γ, Cox2, and C/EPBα in white and brown adipose tissue. Sex and tissue differences were also present for certain adipogenic genes (i.e., Adipoq and Fas expression) (Yan et al., 2014).
Taken together, the human and animal studies suggest that prenatal PAH exposure can elicit changes in the methylation status of coding and noncoding segments of genes and the repetitive elements associated with prenatal PAH exposure.
Synthesis and Conclusions
Reproductive Effects in Men and Women In general, higher PAH levels, assessed using a number of different markers, were found to be associated with significantly higher sperm DNA damage and genetic alterations (Gaspari et al., 2003; Gu et al., 2010; Han et al., 2010; Hsu et al., 2006; Jeng et al., 2013a,b; Jurewicz et al., 2013b; Radwan et al., 2016; Rubes et al., 2005; Song et al., 2013; Xia et al., 2009), including polymorphisms in the xeroderma pigmentosum group A gene, which have been shown to have deleterious effects on spermatogenesis in mice (Gu et al., 2010). Higher concentrations of PAH biomarkers were associated with polymorphisms in the nucleotide excision repair pathway linked to sperm DNA damage, which could provide critical information about the mechanisms involved (Gu et al., 2010).
Other associations with PAH exposure include altered hormone levels, including abnormal and “super” normal LH levels and abnormal levels of FSH and estradiol (Han et al., 2010). Other studies show associations between urinary 1-OHP concentrations and sperm neck abnormalities and decreased semen volume and sperm motility (Jeng et al., 2013a, 2015; Jurewicz et al., 2013b). Associations were stronger in smokers and had some seasonal variation. In addition, associations were seen for total sex chromosome disomy and sperm aneuploidy (Radwan et al., 2016). Occupational studies also show associations between higher PAH levels and male reproductive outcomes, including sperm dysfunction (Hsu et al., 2006; Jeng et al., 2013b, 2015).
The committee noted that research based in several populations shows sperm dysfunction and effects on DNA integrity in epidemiologic studies from a variety of populations and that these effects are supported by results from animal studies. However, because most of the studies on sperm effects assessed exposure using PAH metabolites sampled at the same time as sperm was collected, and PAHs are quickly metabolized, it is unclear if these results reflect long-term effects.
The known effects of PAH exposure on female reproduction are few, with only one study each having been done on fecundability, hormone changes, and PCOS (Luderer et al., 2017a; Slama et al., 2013; Yang et al., 2015). While the evidence of female preconception reproductive outcomes is based on a small number of limited-power epidemiological studies, toxicology and animal-based studies have found mechanistic associations between PAHs and decreased fertility, ovarian effects, decreased estradiol levels, and uterine inflammation (EPA, 2017).
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between exposure to PAHs and reproductive effects in men.
The Volume 11 committee also concludes that there is inadequate/insufficient evidence to determine whether an association exists between exposure to PAHs and reproductive effects in women.
Adverse Pregnancy Outcomes Birth weight and prenatal PAH exposure were assessed in five cohorts, with significant associations reported in two (Choi et al., 2008, 2012; Guo et al., 2012; Langlois et al., 2014; Perera et al., 2005b; Polanska et al., 2010, 2014; Wang and Choi, 2014). Two studies found associations between prenatal PAH exposure and preterm birth (Padula et al., 2014a; Singh et al., 2008). Some measures of BaP and PAH exposure were associated with missed abortion (Wu et al., 2010), and one study reported the effects of certain species of PAHs on the levels of FSH and AMH in umbilical cord serum (Yin et al., 2017). Additionally, the ability of BaP to cause reproductive and developmental effects, including reduced birth weight, had been previously described by EPA (2009). Animal models also suggest that BaP exposure in the second trimester has a dose-dependent effect on fetal survival in pregnant rats and mice. Furthermore, animal studies have reported accumulations of PAHs in the placenta, umbilical cord, and endothelium as well as in neonatal white blood cells (Bocskay et al., 2005).
Together, these epidemiologic and animal studies provide limited/suggestive evidence that PAHs may have an association with adverse birth outcomes.
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between prenatal PAH exposure and adverse birth outcomes—specifically, low birth weight and preterm birth.
Associations between prenatal PAH exposure and birth defects were investigated in the NBDPS in the United States and in birth cohorts in China. In the United States, increased risks were found for cleft lip, gastroschisis, and craniosynostosis but not for congenital heart defects or neural tube defects (Langlois et al., 2012, 2013; Lupo et al., 2012a,b; O’Brien et al., 2016). These analyses relied on an industrial hygienist’s assessment of the maternal self-reported exposures, including occupational exposures, with no assessment of paternal exposures. No associations were seen between maternal occupations that have higher levels of PAH exposure and congenital heart defects (Lupo et al., 2012b). Studies conducted in one region of China at different times showed an inverse association between PAH exposure in 2005–2007 and neural tube defects, and by 2010–2012 there was a significant dose–response relationship between risk of neural tube defects and PAH–DNA adducts (Yi et al., 2015; Yuan et al., 2013). However, the Chinese studies failed to control for confounding factors, including cigarette smoking. The associations with neural tube defects were relatively consistent across study populations and exposure assessment methods.
Prenatal PAH exposure was associated with some childhood cancers, including ALL, neuroblastoma, medulloblastoma, and Wilms’ tumor. However, studies used various distances to air monitoring stations (ranging from 2.5 km to 15 miles) to represent PAH exposure, which may have introduced exposure misclassification, and the associations with trimesters of exposure and specific species of PAHs were inconsistent.
Given that some PAHs are known carcinogens, and considering the recent data showing that prenatal exposure to PAHs is associated with an increased risk of certain tumors, the evidence suggests that these chemicals could affect cancer risk in offspring.
A variety of neurodevelopmental, cognitive, and behavioral deficits have been associated with prenatal PAH exposure up to age 9 years. Few of these studies, however, controlled for postnatal exposures that might contribute to neurobehavioral changes.
Studies of neurobehavioral effects and PAH exposure have used several proxies for exposure, including DNA adducts in cord blood, maternal urinary metabolites, personal sampling, and residential location in polluted areas. Systematic reviews have indicated that prenatal PAH exposure have an impact on neurodevelopment and IQ scores but have also noted the difficulties involved in distinguishing the effects of PAHs from other pollutants (Perera et al., 2011; Polanska et al., 2012; Suades-Gonzalez et al., 2015). The studies reviewed by the committee reported decrements for many different outcomes and measures in association with prenatal PAH exposure, including brain development (BDNF, white matter surface disturbances), MDI scores, mental health scores, and IQ scores; delays in developing motor skills and adaptive and social capabilities (Edwards et al., 2010; Perera et al., 2009b, 2014, 2015; Peterson et al., 2015; Tang et al., 2014b; Vishnevetsky et al., 2015) were also reported in assocation with prenatal PAH exposure. Two studies found no effects on psychomotor or mental development (Perera et al., 2007; Polanska et al., 2013). However, those studies of neurodevelopment were limited to maternal exposures assessed during second and third trimesters of pregnancy, and no studies examined the effects of preconception or first-trimester exposure, or paternal exposure. Neurodevelopmental effects were significantly influenced by factors such as breast-feeding (Jedrychowski et al., 2015a), maternal hardship (Vishnevetsky et al., 2015), nutrition/antioxidant levels (Genkinger et al., 2015), and environmental tobacco smoke (Polanska et al., 2013). None of these studies adjusted for the effects of PAH exposure between birth and the time of the assessment. EPA’s review of BaP did describe supporting evidence for the neurodevelopmental toxicity that has been observed in animal studies (EPA, 2017).
Respiratory effects associated with prenatal PAH exposure were studied in one cohort that assessed prenatal PAH exposure with personal monitors in the second trimester as well as by measuring adducts in maternal and cord blood and maternal urinary metabolites. Significant associations were found between those measures and respiratory symptoms in the first year of life (Jedrychowski et al., 2005). Both prenatal exposure and exposure at 3 years were associated with lung function between the ages of 5 and 9 (Jedrychowski et al., 2015c). In analyses that considered postnatal exposure (measured at age 3), PAH exposure in childhood was found to have a significant effect on respiratory conditions (Jedrychowski et al., 2014). Exposure to cigarette smoking and maternal atopy were significant factors associated with respiratory symptoms (Jedyrchowski et al., 2005, 2012). Additionally, some evidence suggests that respiratory symptoms caused by maternal exposure are mediated by epigenetic changes (Perera et al., 2009a).
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between prenatal exposure to PAHs and developmental effects—specifically, birth defects, childhood cancer, neurodevelopmental effects, and respiratory outcomes—in childhood.
Some studies examined the effect of prenatal PAH exposure on child growth. The growth outcomes associated with that exposure include decreased head circumference, decreased rate of growth (height), and obesity (Jedrychowski et al., 2015b; Rundle et al., 2012; Tang et al., 2006, 2014b). The evidence from this limited number of studies suggests that PAHs could affect growth and metabolic health, but the possibility that the observed associations are related to unknown confounders limits interpretation.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between prenatal PAH exposure and other developmental effects.
TABLE 6-6 Summary of Reproductive and Developmental Effects of Polycyclic Aromatic Hydrocarbons
|Reproductive Effects in Men|
|Gaspari et al. (2003)||Cross-sectional 182 men recruited from an infertility clinic (75 followed up for fertility status 1 year later) in Milan, Italy.
|PAH adducts measured in semen, self-reported occupational and lifestyle exposures.||Sperm abnormalities and fertility.
PAH adducts correlated with physiologic forms (r= –0.18, p=0.016), abnormal head morphology (r=0.30, p=0.0001), and abnormal neck morphology (r= –0.21, p=0.009), but not sperm count or tail morphology.
PAH adducts greater among those with any abnormality (p=0.04) but not among those with asthenospermia, oligospermia, or teratospermia.
PAH adducts were greater in infertile men compared with fertile men (based on pregnancy status) after 1 year of followup (p=0.04).
Associations were observed between alcohol and PAH adducts but not smoking and PAH adducts.
|Gu et al. (2010)||Nested case control 620 infertile men and 273 fertile controls seeking care at a fertility clinic in Nanjing, China.
|Urinary 1-OHP and four functional genotypes.||Sperm DNA integrity.
No significant associations between 1-OHP and genotypes for ERCC1 8092C/A or XPF Ser835Ser.
Significant associations between XPA-4 G/A genotypes and XPD Lys751Gln genotypes and DNA damage among men with high 1-OHP (p<0.05) but not low 1-OHP. Interaction effects: 1-OHP × XPA-4 G/A, p=0.02; 1-OHP × XPD Lys751Gln, p=0.08.
Shows joint effects of PAH exposure and DNA repair gene polymorphisms on male reproductive capacity.
|Han et al. (2010)||Cohort
562 men with unexplained male factor infertility, in Nanjing, China.
|Urinary PAH metabolites 1-N, 2-N, 2-OHF, and 1-OHP.||Reproductive hormone levels in blood (FSH, LH, E2, T, PRL).
Analysis as continuous variables: significant negative correlation between 1-OHP and LH (p=0.01), no other associations found.
Analysis by tertiles of metabolites and trisections of hormones: Increased ORs and significant trends between LH and 1-N (p-trend=0.025); LH and 2-OF (p-trend = 0.019); FSH and 2-OF (p-trend=0.027); E2 and 2-OF (p-trend=0.047); E2 and 1-OP (p-trend= 0.025).
|Han et al. (2011)||Cross-sectional
232 men ages 20–40 years residing in Chongqing, China.
|Urinary hydroxylated PAH metabolites (2-OHNa, 9-OHPh, 2-OHF, 1-OHP).||Sperm DNA damage.
No association between metabolites and semen parameters. Total PAH metabolites associated with PI+ cells (β=12.59, 95% CI 4.15–21.02) and Annexin V-/PI- (β= –13.02, 95% CI –21.55– –4.5), but not Annexin V+/PI- (β=0.03, 95% CI –0.11–0.17).
Significant associations between 2-OHNa and Tail%, tail length, and TDM (ORs=13.26, 12.25, 7.55, all p<0.05).
Total PAH metabolites associated with Tail% (β=15.96, 95% CI 8.86–23.07); but not tail length (β=16.56, 95% CI –0.39–33.52) or TDM (β=6.29, 95% CI –2.36–14.95).
Adjusted for age, abstinence, smoking status.
|Hsu et al. (2006)||Cross-sectional
48 coke oven workers in Taiwan.
|PAHs in personal breathing zone on first and third days after 2 days off; and urinary 1-OHP in urine collected post-shift on days when air monitoring was conducted.||Semen quality and DNA damage.
Among smokers, topside-oven workers had significantly more individuals with oligospermia and asthenospermia, a greater percent of abnormal morphology, and less motility (p<0.05). No significant differences between nonsmoking side-oven and topside-oven workers. No association between air PAH and sperm quality.
Significantly greater DNA denaturation (COMP αT, αT) among smoking topside-oven workers than nonsmoking and smoking side-oven workers (p<0.05).
Urinary 1-OHP significantly related to abnormal morphology (β=0.107, p=0.012), chromatin structure integrity (COMP αT and αT, β=0.235 and 0.078, both p<0.01).
|Jeng et al. (2013b)||Cross-sectional
36 coke oven workers and 15 administrative and security staff as controls, in Taiwan, ages 25–50.
|PAHs in personal breathing zone air sampling on first and sixth days after a 2-day break; and urinary 1-OHP in urine collected before and after shifts on the first and sixth workdays and at the time of semen collection.||Semen quality and DNA integrity.
No differences between exposed workers and controls for semen concentration, oligospermia, volume, pH, motility (progressive, nonlinear, non-progressive), morphology (total head defects, total coiled tail), vitality, and DNA fragmentation.
Significant increases in immotile sperm, asthenospermia, and decreases in normal morphology, and bulky DNA adducts among exposed workers, especially those highly exposed.
TABLE 6-6 Continued
|Jeng et al. (2013a)||Cohort
65 coke oven workers ages 25–60, in Taiwan.
|Occupational exposure measured as urinary 1-OHP levels (pyrene metabolite); and personal breathing zone sampling for 16 PAHs.||Semen quality and DNA integrity.
1-OHP level not correlated with semen concentration, motility, vitality or DNA fragmentation; significantly correlated with morphology (β= –0.28, p<0.05), 8-oxodGuo (β=0.37, p<0.05), and bulky DNA adducts (β=0.45, p<0.05).
Some PAH species were significantly correlated with sperm morphology (14 of 16), motility (4 of 16), 8-oxodGuo (2 of 16), and bulky DNA adducts (2 of 16).
|Jurewicz et al. (2013b)||Cross-sectional
277 men attending an infertility clinic with normal semen concentration or slight oligozoospermia (15–300 mLn/mL), in Poland.
|Urinary 1-OHP, collected at the same time as sperm and questionnaire.||Semen quality parameters: Sperm neck abnormalities (%) β=2.12, p=0.001; Volume (ml) β= –0.06, p=0.014; Motility (%) β= –8.33, p=0.0001; Static (%) β= 0.16, p=0.018.
No significant association between 1-OHP and DFI, sperm head or tail abnormalities, semen concentration, atypical %, or measures of motility. Adjusted for age, smoking, past diseases, season, sexual abstinence.
|Radwan et al. (2016)||Cross-sectional
181 men attending an infertility clinic with normal semen concentration or slight oligozoospermia (20–300mLn/mL), in Poland. 2008–2011
|Urinary 1-OHP, collected at the same time as sperm and questionnaire.||Sperm aneuploidy.
Total sex-chromosome disomy: β=1.46, 95% CI 1.05–2.03; Disomy 18: β=1.22, 95% CI 1.02–1.63.
Nonsignificant results for other kinds of disomy (X-Y-18, X-X-18, Y-Y-18, 13-13, 18-18, 21-21, or total chromosome disomy).
Adjusted for abstinence, age, smoking, season of year, past diseases.
|Rubes et al. 2005||Cohort
36 men born 1974–1976, sampled 7 times between 1995–1997, in the Czech Republic.
|Average ambient PAHs in the 90 days before sample collection based on a monitoring site.||Sperm DNA damage.
No association between PAH/air pollution and sperm count or concentration, volume, motility, morphology, velocity, linearity, or aneuploidy.
Only sperm chromatin abnormalities (%DFI) were related to PAHs/air pollution (β=0.19, 95% CI 0.02–0.36, adjusted for smoking and briefs).
|Song et al. (2013)||Cross-sectional
53 men attending a fertility clinic, in Guangzhou, China.
|PAHs in blood.||Semen quality.
Total PAHs correlated with semen volume, concentration, progressive motility, and nonlinear motility (r= –0.07, 0.10, 0.07, –0.10, all p<0.05).
Total PAHs not associated with motility (expβ=0.96, 95% CI 0.85–1.81). Adjusted for age and BMI.
|Xia et al. (2009)||Case-control
513 infertile men (291 with normal semen quality, 222 with abnormal semen quality) and 273 fertile men, in Nanjing, China.
|Urinary PAH metabolites (1-N, 2-N, 1-OHP, 2-OHF, sum PAHs).||Idiopathic infertility defined by semen quality.
Mean PAH metabolite levels were lowest in controls and highest in cases with abnormal sperm (except for 1-N). Dose–response relationship by tertile of metabolites present for all 5 measures of PAHs, significant for 1-OHP, 2-OHF, and sum PAHs (p-value for trend=0.034, 0.022, and 0.022).
Sum PAHs 2nd tertile OR=1.2, 95% CI 0.84–1.71; 3rd tertile OR=1.53, 95% CI 1.06–2.20; compared to 1st tertile, adjusted for age and abstinence time.
|Yang et al. (2017b)||Cross-sectional
405 men visiting an infertility clinic in Wuhan, China.
(See Yang et al., 2017a)
|PAH metabolites in urine.||Sperm DNA damage.
9-OHF, highest versus lowest tertile, was associated with increased tail length and comet length (mean increases of 8.65%, 95% CI 2.53%–15.03%, p-value for trend=0.05; and 7.14%, 95% CI 2.33%–12.19%, p-value for trend=0.01, respectively).
9-OHPh, highest versus lowest tertile, was associated with decreased percentage of Annexin V−/PI−spermatozoa (mean decrease 5.38%, 95% CI –8.92– –1.86, p-value for trend=0.04).
No significant associations for other metabolites or measures of DNA damage.
|Yang et al. (2017a)||Cross-sectional
933 men visiting an infertility clinic in Wuhan, China.
2013 (See Yang et al., 2017b)
|PAH metabolites in urine.||Semen quality.
1-OHNa, highest versus lowest quartiles, associated with decreased sperm count (–19.99%, 95% CI –33.83%− –3.34%), sperm concentration (–22.66% (95% CI –34.49%– –8.70%), and percentage of normal morphology (–2.35%, 95% CI –4.24%− –0.46%) (all p-value for trends <0.05).
9-OHPh associated with decreased semen volume (0.43 mL, 95% CI –0.74− –0.12) and sperm straight-line velocity (−1.34 μm/sec, 95% CI –2.37− –0.31) (both p for trends <0.05).
4-OHPh associated with decreased semen volume (0.31 mL, 95% CI –0.61− –0.01).
No associations between the urinary PAH metabolites and the total sperm motility to progressive sperm motility. Suggests that naphthalene and phenanthrene are related to decreased semen quality.
TABLE 6-6 Continued
|Conti et al. (2017)||Cross-sectional
86 men in Italy.
|BaP adducts in sperm DNA, and BaP–metabolite adducts [anti--BPDE hydrolyzed DNA adducts or Tetrol I-1 (r-7,c-10t-8,t-9-tetrahydroxy-7,8,9, 10-tetrahydrobenzo(a)pyrene) and syn-BPDE hydrolyzed DNA adducts or Tetrol II-2 (8r-7,t-9,t--10,t-8-tetrahydroxy-7,8,9,10-te-trahydrobenzo(a)pyrene)]||Semen quality.
Tetrol I-1 and Tetrol II-2 adducts were associated with decreased progressive motility (r= –0.26, p=0.014; r= –0.22, p=0.040), but not sperm concentration, normal forms, or count.
|Reproductive Effects in Women|
|Luderer et al. (2017a)||Cohort—California, Orange County, National Children’s Study.
51 fertile women not intending to become pregnant/not using hormonal contraception, ages 18–44.
|Urinary hydroxylated PAH metabolites collected on the 10th day after menses onset for two or three cycles.||Endocrine endpoints: cycle length, follicular phase length, ovulatory status, LH, and estrone 3-glucuronide levels.
Positive associations of fluorene metabolites and negative associations of naphthalene metabolites with two measures of LH surge amplitude. Positive associations of 2-OHNa and 1-OHP and negative association of phenanthrene metabolites with average follicular LH. Follicular phase length was positively associated with 3-OHF and 11-OHP.
|Slama et al. (2013)||Cohort
1,916 births in the Czech Republic.
|Average monthly concentrations of total carcinogenic PAHs at an air-monitoring site.||Self-reported time to pregnancy. Probability of pregnancy during first month of unprotected intercourse not related to average total PAH concentration in fully adjusted model. Excluded unplanned pregnancies, 12 months of infertility, and fertility treatments.|
|Yang et al. (2015)||Case-control
50 cases of PCOS and 30 controls recruited from a hospital in Northern China.
|EDCs in blood (PCBs, organochlorine pesticides, and PAHs).||Polycystic ovary syndrome.
dl-PCBs OR=5.52, 95% CI 1.51–20.2 and all PCBs OR=3.34, 95% CI 0.9–11.6.
Organochlorine pesticides OR=14.8, 95% CI 3.2–68.8; PAHs OR=4.04, 95% CI 1.16–14.0; hydroxyl-PAHs OR=1.12, 95% CI 0.37–3.41.
Octylphenol, nonylphenol, and BPA all nonsignificant. Adjusted for triglycerides, cholesterol, age, education, occupation.
|Adverse Pregnancy Outcomes|
|Choi et al. (2008)||Prospective cohort—CCCEH.
616 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2005.
|Personal monitoring in third trimester for PAHs. Total PAHs expressed as ln.||Gestational age:
African American mothers β= –0.35, 95% CI –0.71–0.006 Dominican mothers β= –0.006, 95% CI –0.19–0.18 Preterm delivery:
African American mothers OR=4.68, 95% CI 1.84–11.89 Dominican mothers OR=0.52, 95% CI 0.18–1.50 Adjusted for BMI, sex, parity, delivery season, ETS exposure. SGA:
African American mothers OR=2.43, 95% CI 1.05–5.62 Dominican mothers OR=0.87, 95% CI 0.51–1.47
Adjusted for BMI, ETS exposure, parity, winter delivery.
Fetal growth ratio <85%:
African American mothers OR=1.93, 95% CI 1.04–3.56 Dominican mothers OR=0.82, 95% CI 0.51–1.33 Adjusted for BMI, gestational weight gain, ETS exposure, parity.
|Choi et al. (2012)||Prospective cohort
344 pregnant women in Krakow, Poland, recruited 2000–2003.
(See Wang and Choi, 2014)
|Personal monitoring of PAHs for 48 hours in each trimester.||Birth size.
PAH exposure in the first trimester is associated with a reduction of markers for growth restriction.
|Guo et al. (2012)||Cohort:
183 pregnant women from Guiya, where there was open combustion of e-waste, and 80 pregnant women from Chaonan, China.
|PAH concentrations in cord blood.||Adverse birth outcomes.
Total PAHs not related to birth height, weight, gestational age, APGAR score, or BMI.
Height negatively associated with BaA (β= –0.23, p=0.006); gestational age negatively associated with chrysene (β= –0.20, p=0.013) and BaP (β= –0.17, p=0.042).
TABLE 6-6 Continued
|Langlois et al. (2014)||Cross-sectional—NBDPS.
2803 births without birth defects (221 were SGA).
|Any self-reported maternal occupational PAH exposure in the month before conception through third month of pregnancy, assigned by an industrial hygienist.||SGA.
OR=2.2 (95% CI 1.3–3.8) (adjusted for maternal age).
17 SGA births had PAH exposure.
|Padula et al. (2014)||Cohort—SAGE study, Fresno, CA.
42,904 births to women living within 20 km of a monitoring site.
|Prenatal exposure is the modeled sum of semi-volatile PAHs based on maternal residence on birth certificate.||PTB on birth certificates.
No association between gestational ages 34–36 weeks, 32–33 weeks, 28–31 weeks and PAHs in any trimester.
PTB 20–27 weeks gestation associated with PAHs in the 6 weeks before birth OR=2.74 (95% CI 2.24–3.34). Dose response for quartiles of exposure, compared to lowest exposure: 2nd quartile OR=1.49 (95% CI 1.08–2.06); 3rd quartile OR=2.63 (95% CI 1.93–3.59); 4th quartile OR=3.94 (95% CI 3.03–5.12).
Stratification by month conceived: greatest association between PTBs and PAHs in the last 6 weeks of pregnancy for births conceived in June–August (20–27 weeks gestation OR=10.2, 95% CI 7.8–13.4).
All adjusted for maternal age, education, race/ethnicity, first trimester prenatal care, payment of birth expenses.
|Perera et al. (2005b)||Cohort—WTC.
186 women pregnant as of 9/11/01 who delivered at hospitals near the WTC, NY in December 2001–June 2002.
|BaP–DNA adducts in maternal and/or cord blood samples at delivery; residential and work addresses in relationship to WTC.||Birth outcomes in medical records.
No significant relationship between BaP adducts and birth weight, length, or head circumference.
Significant effect of BaP adduct x ETS interaction on head circumference.
|Polanska et al. (2010)||Prospective cohort—REPRO_PR.
449 mother–child pairs recruited at 8–12 weeks of pregnancy, in Poland.
|PAH exposure determined by urinary 1-OHP collected at 20–24 weeks pregnancy.||Size at birth as reported by gynecologist or neonatologist.
Birthweight (g) β= –105.40, p=0.10;
Length (cm) β= –0.40, p=0.30; Head circumference (cm) β= –0.30, p=0.20;
Chest circumference (cm) β= –0.40, p=0.07;
Ponderal index (g/cm3) β= –0.003, p=0.40; and
Cephalization index (cm/g) β=3.0, p=0.09.
Adjusted for gestational age, gender, marital status, education level, season of last menstruation period, prepregnancy BMI, weight gain in pregnancy, saliva cotinine level.
|Polanska et al. (2014)||Prospective cohort—REPRO_PR.
210 mother–child pairs recruited at 8–12 weeks of pregnancy, in Poland.
|PAH exposure determined by urinary metabolites collected at 20–24 weeks pregnancy.||Size at birth as reported by gynecologist/neonatologist.
None of the metabolites were significantly associated with birth weight, cephalization index, chest circumference, length, head circumference, or ponderal index, adjusted for gestational age and gender.
Interaction of OH-PHE and environmental tobacco smoke associated with cephalization index (p=0.04).
|Singh et al. (2008)||Case-control
29 preterm deliveries versus 31 full-term deliveries, in India.
placental concentration of 10 PAHs.
Concentrations for almost all PAHs higher among PTBs; fluoranthene (mean: 208.6 versus 325.91) and benzo(b) fluoranthrene (mean: 23.84 versus 61.91) were significantly higher (p<0.05).
|Wang and Choi (2014)||Prospective cohort
344 nonasthmatic children of pregnant women in Krakow, Poland, recruited 2000–2003.
(See Choi et al., 2012)
|Personal monitoring of PAHs for 48 hours in the second trimester.||Birth weight, height, and head circumference.
Reduced head circumference associated with PAHs across the pregnancy but not significant.
Reduced birth weight and height associated with PAHs, but only significant in the first and third trimesters.
Adjusted for season of birth, gestational age at birth, gender, parity, c-section, prepregnancy weight, and maternal height.
TABLE 6-6 Continued
|Yang et al. (2018)||Cohort
106 pregnant women in China.
|PAH metabolites in urine of pregnant women during late pregnancy.||Birth outcomes and DNA methylation of LINE1 and Alu repetitive elements to represent global DNA methylation in cord blood.
Decreased birth length associated with 2-OHNa and ΣOH-PAHs, first versus third tertiles (-0.80%, 95% CI –1.39%––0.20%; and –0.70%, 95% CI –1.39%– –0.10%). No associations between PAH metabolites and birth weight or gestational age.
Decreased Alu methylation associated with 1-OHPh and 2-OHNa (2.57%, 95% CI –4.30%– –0.80%; and 1.88%, 95% CI –3.73%– –0.10%).
No significant associations between LINE-1 methylation and PAH metabolites.
Alu and LINE-1 methylation were not mediators of the associations between PAH metabolites and birth outcomes.
|Dejmek et al. (2000)||Cohort
4854 births (446 IUGR births) in Teplice and Prachatice, Czech Republic.
|Modeled PAHs based on air monitors.||IUGR; birthweight <10th percentile.
For births in Teplice, levels of PAH exposure in the first gestation month was associated with IUGR: medium levels OR=1.60 (95% CI 1.06–2.15) and high levels OR=2.15 (95% CI 1.27–3.63), per 10 ng increase of PAH, OR=1.22 (95% CI 1.07–1.39). Adjusted for parity, maternal age and height, prepregnancy weight, education, marital status, month-specific maternal smoking, season, rhythm, and year of the study.
Similar results for births from Prachatice. No clear associations for exposure during other months of gestation.
|Jedrychowski et al. (2017)||Cohort
455 infants in Krakow, Poland.
|PAHs measured by personal monitor for 48 hours in the second trimester.||Birth outcomes (weight, length, and head circumference).
Analyses also considered PM exposure; PAH exposure was more strongly related to birth outcomes than PM.
Birth weight: β= –0.20, p=0.004
Birth length: β= –0.17, p=0.025
Head circumference: β= –0.13, p=0.086
Adjusted for maternal education (years), child sex, parity, maternal prepregnancy weight (kg), weight gain in pregnancy (kg), gestational age (weeks), exposure to environmental tobacco smoke (yes versus no), and birth season.
|Suzuki. et al (2010)||Cohort
149 pregnant women in Tokyo, Japan.
|Metabolite of pyrene (1-OHP) in urine collected between gestational weeks 9 and 40.||Birth outcomes (weight, height, head circumference, gestational age).
Statistically significant negative associations were observed between 1-OHP and birth weight (β= –0.15, p=0.05), birth length (β= –0.18, p=0.02) and head circumstances (β= –0.17, p=0.05). Analyses of nonsmokers only found no associations.
|Al-Saleh et al. (2013)||Cross-sectional
1578 births in Saudi Arabia.
|PAHs in placenta, maternal blood, and cord blood (BaP, and sum of benzo(a)anthracene, chrysene, benzo(b)fluoranthene, dibenzo(a,h)anthracene; and 1-OHP in maternal urine.||Birth outcomes (birth size, APGAR scores, placental weight and thickness, ponderal index, cephalization index) and markers of oxidative stress (malondialdehyde in cord and maternal serum and 8-hydroxy-2-deoxyquanosine in maternal urine).
None of the tested PAHs was found in maternal or cord blood. All five PAH compounds were detected in placentas. Placental BaP associated with placental thickness (β= –0.071 cm, p=0.018) and cord length (β= –0.074 cm, p=0.013); sum of other placental PAHs associated with cord length (β= –0.074 cm, p=0.013); and urinary 1-OHP was not associated with any birth outcomes.
|Snijder et al. (2012)||Cohort—Generation R Study.
4,680 births in the Netherlands.
|Maternal occupational PAH exposure assessed by questionnaire and industrial hygiene assessment.||Fetal growth and birth outcomes.
Placental weight: β= –7.64, 95% CI –52.03–36.76;
Fetal weight: β= –0.017, SE=0.0080, p<0.05;
Head circumference: β= –0.011, SE=0.011, p>0.05;
Fetal length: β= –0.0031, SE=0.010, p>0.05.
|Wu et al. (2010)||Case-control
81 cases of missed abortion and 81 controls (elective abortions) matched on hospital, maternal age, gravidity, gestational age, in Tianjin, China.
|BaP adducts in fetal tissue and maternal blood.
Proxy measures of PAH exposure based on self-report.
|Missed abortion (embryo has died but miscarriage has not occurred, in early pregnancy). No association between BaP adducts in fetal tissue and missed abortion. BaP adducts in maternal blood (per 108 nucleotides) increased risk of missed abortion—OR=1.35, 95% CI 1.11–1.64. Also associated with residence near traffic congestion (OR=3.07, 95% CI 1.31–7.16), commute by walking (OR=3.52, 95% CI 1.44–8.57), and routinely cooking during preganancy (OR=3.78, 95% CI 1.11–12.87).
Adjusted for maternal education and household income.
TABLE 6-6 Continued
|Yin et al. (2017)||109 pregnant women, in the Shendsi Islands, China.
|16 PAHs in umbilical cord serum.||Reproductive hormones in umbilical cord serum. No significant associations between testosterone, luteotropic hormone, or estradiol and any PAHs; or between acenaphthylene, benzo(b)fluoranthene, benzo(k)fluoranthene, or BaP and any hormone. Significant positive associations between FSH and fluorine, phenanthrene, anthracene, fluoranthene, BaA, chrysene (β=0.47, 0.44, 0.51, 0.38, 0.42, 0.35, all p<0.05). Significant negative associations between anti-Müllerian hormones and naphthalene, acenaphthene, fluorine, and fluoranthene (β= –0.33, –0.33, –0.44, –0.35, all p<0.05).
Adjusted for neonate gender, maternal age, prepregnancy BMI, gestational age, parity, number of abortions, education, pregnancy weight gain.
|Yi et al. (2015)||Case-control
60 cases of NTD and 60 controls matched on birth hospital, sex, mother’s county of residence, mother’s date of last menstrual period; born in northern Shanxi province, China.
|PAH adducts in umbilical cord tissue and blood.||Significant dose response for NTD, spina bifida, and anencephaly with tertiles of PAH adducts in cord tissue (p-value for trend<0.05 for each).
NTDs at highest tertile of PAH adducts in tissue OR=2.96, 95% 1.16–7.58;
Spina bifida at highest tertile OR=2.97, 95% CI 1.18–7.49; Anencephaly OR=3.12, 95% CI 1.14–8.55.
Adjusted for maternal age, education, occupation, parity, hyperthermia during pregnancy, obesity, periconceptional folate supplement.
|Yuan et al. (2013)||Case-control
80 NTD-affected pregnancies and 50 uncomplicated pregnancies in Shanxi Province, China,
|PAHs and PAH adducts in placental tissue.||Controls had greater median concentration of PAH adducts than cases (9.92 versus 8.12 adducts/108 nucleotides, p<0.05). Dose response by PAH adduct tertiles showed inverse relationship: 1st tertile OR=1.48, 95% CI 0.59–3.75 and 2nd tertile OR=2.67, 95% CI 1.07–6.64, compared to the 3rd (highest) tertile (p trend=0.029). Trend also noted for spina bifida and anencephaly subtypes (p trend=0.054 and 0.081).
Stratified by PAH and PAH adducts, only significantly elevated risks remained among those with high PAH exposure and low PAH adducts (OR=8.2, 95% CI 2.1–32.7).
Adjusted for maternal age and hyperthermia.
|Langlois et al. (2012)||Case-control—NBDPS
520 cases of NTD versus 2,989 controls.
|Any self-reported maternal occupational PAH exposure in the month before conception through third month of pregnancy, assigned by an industrial hygienist.||NTDs. OR=1.01 (95% CI 0.61–1.66) (adjusted for BMI, second-hand smoke, study center).
All adjusted ORs for subanalyses were close to 1.0 and not significant.
|Langlois et al. (2013)||Case-control—NBDPS 805 cases of cleft lip with or without cleft palate, 439 cases of cleft palate versus 2,989 controls.
|Any self-reported maternal occupational PAH exposure in the month before conception through third month of pregnancy, assigned by an industrial hygienist.||Cleft lip, with or without cleft palate.
OR=1.47 (95% CI 1.02–2.12) (adjusted for maternal education).
OR=1.24 (95% CI 0.76–2.03) (adjusted for maternal secondhand smoke exposure).
|Lupo et al. (2012a)||Case-control—NBDPS
299 cases of gastroschisis versus 2,993 controls.
OR=1.75, (95% CI 1.05–2.92) (adjusted for maternal age, BMI, education, gestational diabetes, maternal smoking, study center).
Mothers <20 years old: OR=1.14 (95% CI 0.55–2.33);
Mothers ≥20 years old: OR=2.53 (95% CI 1.27–5.04).
|Lupo et al. (2012b)||Case-control—NBDPS
1,907 cases of congenital heart defects versus 2,853 controls.
|Congenital heart defects.
Conotruncal defects: OR=0.98 (95% CI 0.58–1.67);
Left ventricular outflow tract defects: OR=1.31 (95% CI 0.74–2.30);
Right ventricular outflow tract defects: OR=0.54 (95% CI 0.23–1.24); and
Septal defects: OR=1.28 (95% CI 0.86–1.90).
All adjusted for maternal age, race or ethnicity, education, smoking, folic acid supplement, study center.
TABLE 6-6 Continued
|O’Brien et al. (2016)||Case-control—NBDPS
316 cases of craniosynstosis versus 2,993 controls.
16 cases exposed.
OR=1.75 (95% CI 1.01–3.05) (adjusted for maternal age and education).
|Bocskay et al. (2005)||Nested cohort—CCCEH
60 newborns of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited, randomly selected.
|Personal monitoring in third trimester for PAHs; PAH adducts in cord blood.||Chromosomal aberrations as markers of cancer risk.
No associations between PAH air exposure or adducts and unstable aberrations.
Stable aberrations associated with PAH air exposure, r=0.35, p<0.0, and β=0.14, p=0.006.
|Heck et al. (2013a)||Case-control—California Cancer Registry, APCC
75 cases of childhood neuroblastoma younger than 6 years old versus 14,602 matched, randomly selected controls, within 5 km of an air monitor.
|Modeled PAH exposure based on home address listed on birth certificate for each trimester of pregnancy.||Neuroblastoma. All nonsignificant at 5 km. At 2.5 km:
Total PAHs OR=1.39 (95% CI 1.01–1.91);
Benzo(k)fluoranthene OR=1.18 (95% CI 0.91–1.51);
Benzo(a)pyrene OR=1.15 (95% CI0.91–1.46);
Benzo(b)fluoranthene OR=1.19 (95% CI 0.91–1.58);
Indeno(1,2,3-cd)pyrene OR=1.39 (95% CI 1.05–1.84);
Dibenzo(a,h)anthracene OR=1.11 (95% CI 1.03–1.20);
and Benzo(g,h,i)perylene OR=1.62 (95% CI 0.84–3.13).
Adjusted for birth year, mother’s age, mother’s race/ethnicity, method of payment for prenatal care.
|Omidakhsh et al. (2018)||Case-control
282 cases of retinoblastoma and 155 friend-controls recruited from multiple medical centers.
|Paternal occupational PAH exposure in 10 years before conception, and maternal occupational PAH exposure in the month before conception ascertained by questionnaire.||Retinoblastoma in children up to 15 years old.
Paternal PAH exposure 10 years before conception, unilateral retinoblastoma OR=1.34, 95% CI 0.48–3.72; bilateral retinoblastoma OR=1.51, 95% CI 0.36–6.38. Analyses of paternal PAH exposure 6 months before conception showed no association. Associations were stronger among fathers >30 years of age, fathers who had exposure to any hazardous agent, and a higher income bracket.
|Heck et al. (2014)||Case-control—California Cancer Registry, APCC Cases of childhood cancer younger than 6 years old matched with randomly selected controls based on data from birth certificates.
69 cases ALL versus 2,994 controls, within 2 km of an air monitor; 46 cases of AML versus 19,209 controls, within 6 km of an air monitor.
|Modeled PAH exposure based on home address listed on birth certificate for each trimester of pregnancy.||ALL:
Total PAHs through entire pregnancy OR=1.17 (95% CI 0.94–1.45);
Total PAHs in 1st trimester OR=0.95 (95% CI 0.82–1.11);
Total PAHs in 2nd trimester OR=1.04 (95% CI 0.94–1.14); and
Total PAHs in 3rd trimester OR=1.16 (95% CI 1.04–1.29).
Exposure to each individual PAH in 3rd trimester statistically significant, ORs range 1.06–1.27.
Total PAHs through entire pregnancy OR=1.07 (95% CI 0.69–1.65);
Total PAHs in 1st trimester OR=0.88 (95% CI 0.63–1.21);
Total PAHs in 2nd trimester OR=0.96 (95% CI 0.74–1.24); and Total PAHs in 3rd trimester OR=1.13 (95% CI 0.93–1.36).
Adjusted for maternal race/ethnicity, birth year, parity, maternal birth place, neighborhood SES estimated by census block.
|Shrestha et al. (2014)||Case-control—California Cancer Registry, APCC
337 cases of childhood Wilms’ tumor younger than 6 years old versus 96,514 matched, randomly selected controls based on data from birth certificates, within 15 mi of an air monitor.
|Modeled PAH exposure based on home address listed on birth certificate for each trimester of pregnancy.||Wilms’ tumor.
1st trimester OR=0.99 (95% CI 0.88–1.12)
2nd trimester OR=1.06 (95% CI 0.98–1.15)
3rd trimester OR=1.10 (95% CI 0.99–1.22)
Entire pregnancy OR=1.15 (95% CI 1.02–1.31)
Adjusted for birth year, mother’s age, mother’s race/ethnicity, parity, SES estimated by census block.
TABLE 6-6 Continued
|von Ehrenstein et al. (2016)||Case-control—California Cancer Registry, APCC Cases of childhood cancer younger than 6 years old matched, randomly selected controls. 43 cases of PNET, 34 cases of medulloblastoma, 106 cases of astrocytoma versus 30,569 controls, within 5 mi of an air monitor.
Total PAHs OR=1.06 (95% CI 0.73–1.55); and
All individual PAHs nonsignificant, ORs range 0.81–1.58.
Total PAHs OR=1.44 (95% CI 1.15–1.80);
Benzo(k)fluoranthene OR=1.44 (95% CI 1.07–1.52);
Benzo(a)pyrene OR=1.25 (95% CI 1.06–1.46);
Benzo(b)fluoranthene OR=1.33 (95% CI 1.10–1.61);
Indeno(1,2,3-cd)pyrene OR=1.38 (95% CI 1.13–1.69);
Dibenzo(a,h)anthracene OR=1.07 (95% CI 1.01–1.14); and
Benzo(g,h,i)perylene OR=1.94 (95% CI 1.20–3.14).
Total PAHs OR=1.06 (95% CI 0.85–1.33).
All individual PAHs nonsignificant, ORs range 0.96–1.09.
Adjusted for maternal race/ethnicity, age, education, birth place, and birth year.
|Edwards et al. (2010)||Prospective cohort 214 children of pregnant women in Krakow, Poland.
|Personal monitoring of PAHs for 48 hours in the second or third trimester.||Neurodevelopment (Raven Coloured Progressive Matrices score) at age 5.
PAHs associated with significantly reduced neurodevelopmental scores.
PAH (high versus low) β= –1.36, p=0.02 ln(PAH) β= –0.56, p=0.02
Both models adjusted for environmental tobacco smoke, sex of child, maternal education.
|Jedrychowski et al. (2015a)||Prospective cohort
170 children of pregnant women in Krakow, Poland, recruited 2000–2003.
|PAH adducts in cord blood at birth; and household air monitoring for 48 hours at age 3.||IQ at age 7.
17 children had a depressed verbal IQ score.
No significant IQ or covariate differences between detectable and non-detectable levels of PAH–DNA adducts in cord blood.
To predict depressed verbal IQ scores ≥22 points below: Cord blood adducts RR=3.0 (95% CI 1.32–6.79);
Postnatal PAH RR=1.63 (95% CI 1.03–2.53).
Exclusive breast feeding (6+ months) RR=0.31 (95% CI 0.11–089).
Adjusted for birth season, birth season × PAH–DNA adducts interaction, maternal education, gender, parity.
|Perera et al. (2007)||Prospective cohort
98 children of women pregnant as of 9/11/01 who delivered at hospitals near the WTC, NY in December 2001–June 2002.
|BaP–DNA adducts in maternal and/or cord blood samples at delivery; residential and work addresses in relationship to WTC.||Development at age 3 documented in medical records.
No significant effect of BaP adducts on psychomotor development index or MDI. Significant interaction of BaP adducts × ETS on mental development index (β= –12.25, p=0.02).
|Perera et al. (2009b)||Prospective cohort—CCCEH
249 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2003.
|Personal monitoring in third trimester for PAHs.
Does not distinguish between prenatal and postnatal exposures.
|IQ (Wechsler Preschool and Primary Scale of Intelligence-Revised) at age 5.
IQ versus high or low PAH exposure (< or >2.26ng/m3):
Full-scale IQ β= –4.31, p=0.007;
Verbal IQ β= –4.67, p=0.003;
Performance IQ β= –2.37, p=0.17.
Adjusted for ETS exposure, gender, maternal education, ethnicity, HOME score, Test of Maternal Nonverbal Intelligence, Third Edition score.
IQ versus ln(PAH):
Full-scale IQ β= –3.0, p=0.009;
Verbal IQ β= –3.53, p=0.002; and
Performance IQ β= –1.47, p=0.24.
TABLE 6-6 Continued
|Perera et al. (2012)||Prospective cohort—CCCEH
253 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2003.
|Personal monitoring in third trimester for PAHs; BaP adducts in maternal and cord blood at delivery.
BaP used as proxy for PAHs.
Does not distinguish between prenatal and postnatal exposures.
|Behavior at ages 6–7 (CBCL).
Environmental PAH (high/low) OR=8.89, 95% CI 1.7–46.51 and PAH (continuous) OR=1.2, 95% CI 1.06–1.37);
Maternal and cord adducts ORs=1.42 and 2.56, both nonsignificant.
Attention problems scale:
Environmental PAH (high/low) OR=3.79, 95% CI 1.14–12.66;
Maternal and cord adducts ORs=2.24 and 4.06, both nonsignificant.
Anxiety disorder DSM-IV:
Environmental PAH (high/low) OR=4.59, 95% CI 1.46–14.27 and PAH (continuous) OR=2.03, 95% CI 0.97–4.26;
Maternal and cord adducts ORs=2.19 and 2.53, both nonsignificant.
Environmental PAH (high/low) OR=2.3, 95% CI 0.79–6.7 and
PAH (continuous) OR=1.54, 95% CI 0.7–3.4;
Maternal and cord adducts, ORs=1.84 and 2.64, both nonsignificant.
Adjusted for prenatal ETS, sex of child, gestational age, maternal IQ, HOME inventory, maternal education, ethnicity, prenatal demoralization, age at assessment, season.
|Perera et al. (2014)||Prospective cohort—CCCEH
250 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2006.
|BaP adducts in maternal and cord blood at delivery; urinary PAH metabolites at ages 3 or 5 years.||ADHD and behavior at age 9 (CBCL and CPRS).
Cord blood adducts not related to any ADHA outcome.
CBCL ADHD not related to PAH exposure.
Maternal BaP adducts related to CPRS scales:
ADHD index β=0.14, 95% CI 0.03–0.25, OR=1.83, 95% CI 0.61–5.54;
DSM-IV hyperactive-impulsive β=0.16, 95% CI 0.03–0.29, OR=1.58, 95% CI 0.55–4.52;
DSM-IV inattentive β=0.17, 95% CI 0.04–0.31, OR=5.06, 95% CI 1.43–17.93.
Adjusted for postnatal PAH exposure by urinary metabolites at age 3 or 5, prenatal ETS, child sex, maternal education, ethnicity, gestational age, maternal demoralization, season, HOME caretaking environment, maternal intelligence, age at assessment, maternal ADHD, child anxiety/depression at age 9.
|Perera et al. (2015)||Prospective cohort—CCCEH
505 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2006.
|32P post-labeling assay for PAH adducts in cord blood and BaP adducts in maternal and cord blood at delivery.
Does not distinguish between prenatal and postnatal exposures.
|BNDF in cord blood; MDI (using the Bayley Scales of Infant Development-Revised) at age 2.
BaP-adducts not related to BNDF or MDI.
Effect of 32P adducts on MDI: β= –2.07, p=0.04;
Effect of 32P adducts on BDNF: β= –0.11, p=0.02;
Effect of BNDF on MDI: β=2.69, p=0.003;
Effect of 32P adducts (and BNDF) on MDI: β= –1.92, p=0.07.
Conclusion: PAHs affect development by reducing BNDF;
BNDF plays a role in early brain development.
|Peterson et al. (2015)||Prospective cohort—CCCEH
40 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2006.
|Personal monitoring in third trimester for PAHs.||Brain white matter, cognition, and behavior (WISC-IV) at 7–9 years old.
Prenatal PAH inversely correlated with white matter surface measures, especially the left hemisphere.
Reduced white matter associated with externalizing problems and externalizing symptoms.
High prenatal PAH associated with reduced processing speed during intelligence testing (mediated by white matter disturbances).
Prenatal PAH not correlated with PAHs at age 5.
TABLE 6-6 Continued
|Tang et al. (2014a)||Prospective cohort
150 newborns born March–June 2002 and 158 newborns born March–May 2005, residing within 2.5 km of operational Tongliang power plant in China (power plant shut down in 2004).
(see Tang et al., 2014b)
|Environmental air sampling at 3 sites in 2002–2003 and 2005–2006; BaP–DNA adducts (as proxy for PAH–DNA adducts) in maternal and cord blood at birth.||Development at age 2 (Gesell Developmental Schedule) and BDNF, a protein involved in neuronal growth, in umbilical cord blood.
Lower mean levels of PAH–DNA adducts (0.204 versus 0.324 adducts/108 nucleotides), higher concentrations of BDNF (1266.57 versus 752.87 μg/dL) and higher DQ scores (100.3 versus 99.42) were in the 2005 cohort enrolled after closure of the power plant (all p<0.05).
Among both groups combined, PAH–DNA adducts were correlated with BDNF (r= –0.23, p<0.01) as well as the change in motor scores (β= –10.70, p = 0.05), adaptive scores (β= –16.48, p=0.022), and average score (β= –12.11, p=0.014), after adjusting for cord blood lead, cord mercury, ETS, mother’s education, mother’s age, gestational age, and gender
BDNF levels (log BNDF) were associated with changes in motor (β=2.12, p=0.018), social (β=3.22, p=0.001), and average score (β=2.50, p=0.017), adjusted for cord PAH–DNA adducts, prenatal ETS, cord lead, cord mercury, gender, gestational age, mother’s education, mother’s age, and income.
|Polanska et al. (2013)||Prospective cohort—REPRO_PR.
411 mother–child pairs examined at 1 year of age in Poland, 198 mother–child pairs examined at 2 years of age, recruited at 0–12 weeks of pregnancy in 2007.
|PAH exposure determined by urinary 1-OHP collected at 20–24 weeks pregnancy.||Neurodevelopment at 1 or 2 years.
1-OHP levels not associated with psychomotor scores at 1 or 2 years.
|Vishnevetsky et al. (2015)||Prospective cohort—CCCEH
276 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2006.
|PAH adducts in maternal and cord blood at delivery, PAH metabolites in urine at age 5.||IQ at age 7 using the WISC-IV.
Effect of prenatal material hardship × adducts interaction on working memory β= –8.07, 95% CI –14.48– –1.66.
Nonsignificant for full scale, verbal comprehension, processing speed, and perceptual reasoning.
No associations between IQ components and PAH-adducts among low prenatal material hardship subjects.
Among high prenatal material hardship subjects, PAH adducts associated with the full scale β= –5.81, 95% CI –10.35––1.26;
verbal comprehension β= –3.36, 95% CI –7.61–0.90;
processing speed β= –4.17, 95% CI –9.75–1.41;
perceptual reasoning β= –5.44, 95% CI –10.27– –0.61; and
working memory β= –6.67, 95% CI –11.38– –1.95.
Adjusted for ETS, sex, maternal education and intelligence, ethnicity, HOME caretaking.
|Genkinger et al. (2015)||Cohort—CCCEH
151 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, recruited 1998–2006.
|PAH exposure measured by personal air monitoring for 48 hours in the third trimester.||Behavior between 6 and 9 years of age modified by micronutrients measured in cord blood.
Interactions were observed between the effects of prenatal micronutrients and PAH exposure on childhood behavior/symptoms.
α-tocopheroland PAH: withdrawn/depressed β=0.25, p=0.02; aggressive behavior β=0.15, p=0.02; internalizing problems β=0.11, p=0.03; externalizing problems β=0.11, p=0.04.
γ-tocopheroland PAH: internalizing problems β=0.14, p=0.007.
Carotenoids and PAH: somatic complaints β= –0.23, p=0.05; aggressive behavior β=0.14, p=0.02; externalizing problems β=0.12, p=0.02.
Retinol and PAH: no significant interaction effects. Adjusted for sex of the newborn, maternal gestational age, maternal years of education, maternal psychological distress, the age of the child at the time of the CBCL assessment, heating season, lead exposure, prenatal and postnatal environmental tobacco smoke, and dietary PAH.
TABLE 6-6 Continued
|Rundle et al. (2012)||Prospective cohort—CCCEH
702 children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, NY, in 1998–2006 (453 at age 5; 371 at age 7).
|Personal monitoring in third trimester for PAHs.
Does not distinguish between prenatal and postnatal exposures.
|Obesity at ages 5 and 7.
At age 5, compared to lowest tertile of PAH exposure (>1.73ng/m3):
PAH 2nd tertile (1.73–3.07 ng/m3) RR= 1.79, 95% CI 1.08–2.98.
PAH 3rd tertile (≥3.08 ng/m3) RR=1.79, 95% CI 1.09–2.96.
At age 7: PAH 2nd tertile RR=2.25, 95% CI 1.27–4.01.
PAH 3rd tertile RR=2.26, 95% CI 1.28–4.00.
Adjusted for birth weight, ethnicity, child’s sex, maternal prepregnancy obesity, child age at measurement.
|Tang et al. (2006)||Prospective cohort
150 newborns born March–June 2002 residing within 2.5 km of Tongliang power plant in China.
|Environmental exposure estimated from number of months of gestation during power plant operation; PAH adducts in maternal and cord blood at birth.||Growth at birth, 18, 24, and 30 months.
Maternal adducts not related to growth.
High cord blood adducts (0.36 adducts/10–8 nucleotides) significantly associated with lower weight (p=0.023), smaller head circumference (p=0.056), but not height.
Duration of exposure associated with height (p<0.001) but not weight or head circumference. Adjusted for ETS, sex, maternal height and weight.
|Tang et al. (2014b)||Prospective cohort
150 newborns born March–June 2002 and 158 newborns born March–May 2005, residing within 2.5 km of operational Tongliang power plant in China (power plant shut down in 2004).
|Environmental air sampling at 3 sites in 2002–2003 and 2005–2006; BaP adducts in maternal and cord blood at birth.||Growth at birth, 18, 24, and 30 months and IQ (Gesnel Developmental Scale) at age 2.
The 2005 cohort had significantly lower environmental BaP and lower BaP adducts (p<0.001).
Growth rates in the 2005 cohort were not greater than the 2002 cohort.
Cord adducts significantly negatively associated with motor developmental quotient in the 2002 cohort, β= –16.01, p=0.043 (but not for adaptive, language, or social measures or any measures in the 2005 cohort).
|Jedrychowski et al. (2015b)||Prospective cohort
379 children of pregnant women in Krakow, Poland, recruited 2000–2004.
|Personal monitoring of PAHs for 48 hours in the second trimester; and household air monitoring for 48 hours at age 3 years.||Height gain at age 9. Prenatal PAH exposure correlated with birth length (r= –0.092, p<0.05), but not height at older ages.
Postnatal household PAHs not related to height.
Height decreased 1cm for children exposed to PAHs >34.7 ng/m3 (p=0.04), adjusted for cord blood Hg, cord blood Pb, prepregnancy maternal weight, maternal height, maternal weight gain in pregnancy, maternal education, sex of child, age of child, parity. Birth length and maternal height were the strongest predictors of child’s height.
|Jedrychowski et al. (2005)||Prospective cohort
333 pregnant women in Krakow, Poland, recruited 2000–2002.
|Personal monitoring of PAHs for 48 hours in the second or third trimesters.
Does not distinguish effects of prenatal from postnatal exposure.
|Respiratory symptoms in first year of life—runny or stuffy nose, ear infection, cough, cough without cold, barking cough, difficult breathing, wheezing, wheezing without cold, sore throat.
Statistically significant positive associations between either number of episodes and/or the duration for all the symptoms.
Highest OR=4.8 (95% CI 2.73–8.44) for number of episodes of barking cough.
Adjusted for child gender, birth weight, season of birth, postnatal ETS, mother’s allergy, mother’s education, and molds at home.
|Jedrychowski et al. (2012)||Prospective cohort
82 nonasthmatic children of pregnant women in Krakow, Poland, recruited 2000–2002.
|PAH adducts in cord blood at birth. Does not distinguish effects of prenatal from postnatal exposure.||Fractional exhaled nitric oxide at age 7.
Regression model predicting fractional exhaled nitric oxide:
Prenatal PAH exposure β=0.32, p=0.006; maternal atopy β=0.25, p=0.03; atopy β=0.23, p=0.4.
|Jedrychowski et al. (2014)||Prospective cohort
257 children to pregnant women in Krakow, Poland, recruited 2000–2002.
|Personal monitoring of PAHs for 48 hours in the second trimester; and survey of home and outdoor air pollution at age 3 years.||Wheezing (20 or more days in the follow-up period) up to age 4.
Prenatal PAH OR=1.40 (95% CI 0.97–2.03);
Postnatal PAH OR=1.61 (95% CI 1.16–2.24).
Prenatal exposure modifies effect of postnatal exposure on wheezing.
Adjusted for asthma and number of older siblings.
TABLE 6-6 Continued
|Jedrychowski et al. (2015c)||Prospective cohort
195 nonasthmatic children of pregnant women in Krakow, Poland.
|Personal monitoring of PAHs for 48 hours in the second trimester; and household air monitoring for 48 hours at age 3 years.||Lung function at ages 5–9; and atopy (skin prick) at ages 5–8.
Highest tertiles of exposure for prenatal PAHs, indoor PAHs at age 3, and outdoor PAHs at age 3 significantly related to reduced FEV05, FEV1, FEF25–75 (only outdoor PAHs at age 3 related to reduced FVC). Adjusted for gestational age, gender, height, parity, atopic status, ETS, and season of residential air pollution survey.
|Perera et al. (2009a)||Sub-cohort—CCCEH Children of African American and Dominican women living in Washington Heights, Harlem, or South Bronx, New York, recruited 1998–2006. 20 children sampled for DNAm and PAH analyses; 56 children sampled for analyses of DNAm and asthma.||Personal air monitoring for 48 hours in the third trimester.||Asthma and DNAm in cord blood and placental tissue. More than 30 DNA sequences were identified whose methylation status was associated with maternal PAH exposure; 6 sequences were related to genes with ≥1 5’CpG islands; and acyl-CoA synthetase long-chain family member 3 (ACSL3) had the highest concordance between methylation of 5’-CpG islands in umbilical cord white blood cells and the level of gene expression in matched fetal placental tissue in the initial 20 cohort children (tau= –0.45, 95% CI –0.76––0.15).
ACSL3 is related to asthma and was then investigated in the sample of 56 children. ACSL3 59-CpG island methylation was significantly associated with high maternal airborne PAH exposure (>0.41 ng/m3; OR=13.8; p<0.001; sensitivity=75%; specificity=82%) and with a parental report of asthma symptoms in children prior to age 5 (OR=3.9, 95% CI 1.1–14.3, p<0.05).
NOTE: ADHD=attention deficit–hyperactivity disorder; ALL=acute lymphoblastic leukemia; AML=acute myeloid leukemia; APCC=Air Pollution and Childhood Cancer Study; APGAR=Appearance, Pulse, Grimace, Activity, and Respiration; BaA=benzo[a]anthracene; BaP=benzo[a]pyrene; BMI=body mass index; BNDF= brain-derived neurotrophic factor; BPA=bisphenol; BPDE=benzo[a]pyrene diol-epoxide-DNA adducts; BSID=Bayley Scales of Infant Development II; CBCL=Child Behavioral Checklist; CCCEH=Columbia Center for Children’s Environmental Health study; CI=confidence interval; COMP αT=cells falling outside of the main population; CTRS=Conner’s parent rating scale-revised; DFI=DNA fragmentation index; DNA=deoxyribonucleic acid; DNAm=DNA methylation; DQ=development quotient; DSM-IV=Diagnostic and Statistical Manual of Mental Disorders, Fourth Edition; E2=estradiol; EDC=endocrine disrupting chemical; ETS=environmental tobacco smoke; FEF25–75=forced expiratory flow 25–75%; FEV0.5=forced expiratory volume in 0.5 s; FEV1=forced expiratory volume in 1 s; FSH=follicle-stimulating hormone; FVC=forced volume capacity; HOME=Home Observation for Measurement of the Environment; IQ=intelligence quotient; IUGR=intrauterine growth restriction; LH=luteinizing hormone; MDI=mental development index; 1-N=1-naphthol; 2-N=2-naphthol; NBDPS=National Birth Defects Prevention Study; NTD=neural tube defect; 2-OF=2-hydroxyfluorene; OHNa=hydroxynaphtha¬lene; 1-OHP=1 hydroxypyrene; OHPh=hydroxyphenanthrene; OR=odds ratio; PAH=polycyclic aromatic hydrocarbon; PCB=polycyclic biphenyl; PCOS=polycystic ovary syndrome; PM=particulate matter; PNET=primitive neuroectodermal tumors; PRL=prolactin; PTB=preterm birth; RR=relative risk; SE=standard error; SES=socioeconomic status; SGA=small for gestational age; T=testosterone; TDM=tail distributed moment; WISC-IV=Wechsler Intelligence Scale for Children-IV; WTC=World Trade Center.
Polychlorinated Dibenzodioxins and Polychlorinated Dibenzofurans
PCDD/PCDFs are products of combustion that have been measured in the emissions from burn pits on military bases (IOM, 2011; Masiol et al., 2016a,b). Air monitoring at JBB in 2007 and 2009 showed that at some sampling locations PCDD/PCDFs concentrations were relatively high compared to urban and industrial areas, such as Beijing, with average concentrations reaching 1,309 fg I-TEQ/m3 (IOM, 2011).3 A pilot study showed that 400 Post-9/11 military personnel who had been deployed to bases operating burn pits or other regions had higher serum concentrations of PCDD/PCDFs congeners than the general U.S. population. For example, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) was detected in 8 out of 800 serum samples, with a mean of 0.2 pg/mL (200 ppt4), compared to a geometric mean of <0.0038 pg/mL (3.8 ppt) in the general population, as reported in the 2003–2004 National Health and Nutrition Examination Survey. Pre- versus post-deployment samples indicated that exposures during deployment did not lead to increases in congener levels (Xia et al., 2016). Estimates of Vietnam veterans’ exposures to PCDD/PCDFs as a result of the TCDD contamination of Agent Orange are also problematic because of the lack of data on individual behaviors and locations, of an understanding of the chemicals’ behavior in the environment, and of contemporaneous chemical measurements. However, TCDD exposure estimates based on serum samples indicate that Vietnam veterans who sprayed Agent Orange (Ranch Hands, the most highly exposed group of veterans) were exposed to levels in the ppt range (about 10–20 ppt), which although higher than for ground troops, were an order of magnitude lower than exposure levels for occupational cohorts (20–300 ppt) and three orders of magnitude lower than exposure levels associated with the Seveso, Italy, industrial explosion (4,540–16,600 ppt) (NASEM, 2016; Pirkle et al., 1995).
Like PAHs, PCDD/PCDFs exist as part of a variable mixture of many species, the toxicity of which is usually presented based on the levels of the most toxic species (TCDD). Even though there are many species of PCDD/PCDFs, TCDD has been the most extensively studied species because of its high toxicity and its common occurrence in herbicides, including Agent Orange, during the Vietnam War. TCDD is a persistent compound with a biological half-life that in human studies has ranged from 3 months to more than 10 years depending on BMI, age, sex, and TCDD concentration (NASEM, 2016).
There is extensive information available about the health effects of TCDD from authoritative bodies such as EPA, ATSDR, and the National Academies, all of which have described studies on the health of children born to exposed parents. Specifically, in 2012 EPA released an assessment of TCDD and ATSDR released an assessment of the larger class of chemicals—chlorinated dibenzodioxins—both of which included literature published through 2011 (ATSDR, 2012; EPA, 2012). The most recent review, conducted by the National Academies in 2016, is the 10th update in the Veterans and Agent Orange (VAO) series on the health effects of Agent Orange in Vietnam veterans (NASEM, 2016). The Gulf War and Health Volume 11 committee relied on the 2016 VAO report and any new data published since that review’s cutoff in 2014.
The 10th VAO report reviewed the literature on TCDD and other chemical constituents present in the herbicide mixtures used in Vietnam, through 2014, and found that
3TEQs are a weighted quantity measure based on the toxicity of each member of the dioxin and dioxin-like compounds category relative to the most toxic members of the category. TEQs are used to account for how dioxin and dioxin-like compounds vary in toxicity and to better understand the cumulative toxicity of mixtures of dioxin and dioxin-like compounds (EPA, 2016).
4 1 pg/mL=1,000 ppt; 1 ppt=0.001 pg/mL.
- The studies linking dioxin exposure with endometriosis are few and inconsistent. Although animal studies support the biologic plausibility of an association, contemporary human exposures may be too low to show an association consistently.
- Regarding fertility, although there was relative consistency in reports of reduced testosterone with the highest exposure and increased gonadotropins, levels were still within the normal ranges and clinical consequences were not expected. Only one study examined, and observed, the relationship between dioxin exposure in women and time to pregnancy and infertility, and no data were available regarding effects on menstrual cycle function in humans, thus data were insufficient.
- A single study concerning dioxin and spontaneous abortion, stillbirth, neonatal death, or infant death has been published since the previous volume, but it did not provide supporting evidence of an association between dioxin and these outcomes, nor do toxicologic studies provide clear evidence for the biologic plausibility of an association (p. 741).
- Although the animal literature does support an effect of TCDD exposure at high doses on birth weight, the epidemiologic literature is insufficiently robust to allow a final determination (p. 735).
Based on these findings, the VAO committee concluded that
there is limited or suggestive evidence that paternal exposure to TCDD is not associated with risk of spontaneous abortion and that insufficient information is available to determine whether there is an association between maternal exposure to TCDD or either maternal or paternal exposure to 2,4-D, 2,4,5-T, picloram, or cacodylic acid and the risk of spontaneous abortion. The committee concludes that there is inadequate or insufficient evidence to determine whether there is an association between exposure to the COIs [chemicals of interest] and stillbirth, neonatal death, or infant death.
That same committee also gave special consideration to the possibility of adverse health outcomes at any time during the lives of all progeny of Vietnam veterans. Regarding the potential health effects in the children of Vietnam veterans, that committee found that
- It is clear that the fetal rodent is more sensitive to the adverse effects of TCDD than the adult rodent. Human data are generally lacking, however, and the sensitivity to developmental disruption in humans is less apparent, in part because contemporary studies of environmental dioxin exposure and birth defects have involved extremely low exposures. Human population-based studies have provided mixed results in attempts to link TCDD exposures to birth defects.
- No associations were observed in the two case-control studies that considered childhood ALL [acute lymphoblastic leukemia] and exposure to phenoxy herbicides and to 2,4-D in particular. Furthermore, evidence is sparse that exposure to dioxin increases the risk for childhood cancers.
- The animal literature contains evidence that environmental agents mediated by maternal exposure affect later generations through fetal and germ line modifications, but in the case of adult male exposures before conception of the next generation, there is insufficient evidence of transgenerational effects. Although it may be physiologically possible for paternal exposure to cause changes in offspring that are manifested later in life, none of the published epidemiologic studies assessed such potential. Thus, the observation of any changes reported in studies should be applicable only to children born to female Vietnam veterans during or after their deployment in Vietnam. Thus, no transgenerational studies have been reported to date.
The committee concluded that no adverse outcomes in future generations had sufficient evidence of an association with dioxin and that there was inadequate or insufficient evidence to determine whether there is an association between parental exposure to the chemicals of interest and birth defects, childhood cancers, or disease in the children of Vietnam veterans as they mature or in later generations (NASEM, 2016).
An 11th update in the VAO series is scheduled to be released in fall 2018. Like the previous volumes, it reviews any reproductive and developmental outcomes in Vietnam veterans and their descendants.
Because of the extensive volume of research conducted on dioxins, the Volume 11 committee limited its review to studies on reproductive, developmental, and generational effects published since the most recent NASEM review (see Table 6-7). The results include 26 human studies and 21 animal studies published since 2014.
Reproductive Effects in Men and Women
In a case-control study in Belgium, subfertility was found to be not related to PCDD/PCDF exposure (OR=1.59, 95% CI 0.96–2.65) among 40 infertile men and 80 controls (Den Hond et al., 2015). A Russian study, however, identified levels of PCDD/PCDFs in the semen of 105 infertile men (both with and without abnormal sperm) that were more than twice as high as those in fertile men (n=63), but the results were not adjusted for any potential confounders (Galimova et al., 2015). Among 251 men living near a municipal waste facility in France, semen parameters were largely unaffected by PCDD/PCDF exposure except for a reduced multiple anomalies index (β= –0.013, 95% CI –0.023– –0.004) and fewer abnormal mid-pieces (β= –0.010, 95% CI –0.018– –0.002) by year as emissions decreased (Faure et al., 2014).
In a small case-control study of 30 women with deep infiltrating endometriosis and 30 controls, adipose samples from the cases contained higher median TEQs for 3 out of the 17 PCDD/PCDFs that were quantified. However, any differences disappeared after the results were adjusted for BMI, age, and smoking (Martinez-Zamora et al., 2015).
In an in vitro study, estradiol-17β production was decreased in human luteinizing granulosa cells that had been cultured in the presence of TCDD (Baldridge et al., 2015), but clear effects on ovulation and fertility have not been seen in humans (NASEM, 2016).
Adverse Pregnancy Outcomes
The sex ratios of children born to workers at a New Zealand herbicide production plant were compared with back-calculated TCDD serum concentrations at the time of birth. Although there was no relationship between the sex ratios and TCDD levels among exposed mothers, the probability of having a male child was significantly reduced for fathers with serum levels ≥20 pg TCDD/g lipid (sex ratio=0.47, OR=0.49, 95% CI 0.30–0.79) (Mannetje et al., 2016).
A systematic review and meta-analysis that included 15 studies on dioxin exposure and birth outcomes found associations between dioxins in food and LBW and between maternal exposure to dioxin in solid contaminants (residing near hazardous waste sites or a solid waste incinerator) and birth defects, but dioxin exposure was not associated with other outcomes, including spontaneous abortion, SGA, and stillbirth (Pan et al., 2015). Researchers have examined birth outcomes from 1996 through 2009 among women exposed to TCDD as a result of the 1976 chemical explosion in Seveso, Italy. In that study, no associations were found between exposure and spontaneous abortion, fetal growth, gestational length, or birth weight (Wesselink et al., 2014). However, specific genotypes can also affect the impact of PCDD/
PCDF exposure on birth outcomes. For example, a strong association between PCDD/PCDF exposure and reduced birth weight was reported for women with glutathione S-transferase mu 1(GSTM1) null genotype (β= –345, 95% CI –584– –105), but not women with aromatic hydrocarbon receptor (AHR), cytochrome P450 1A1 (CYP1A1), or GSTM1 non-null genotypes (Kobayashi et al., 2017).
An ecological study in Guam of infant deaths in 1970–1989 due to congenital anomalies found that deaths were more likely among mothers who lived in areas that had been sprayed with Agent Orange (β=2.02, 95% CI 0.08–3.96) (Noel et al., 2015).
In the latest VAO review (NASEM, 2016), there was no available epidemiologic evidence that convincingly demonstrated a relationship between paternal exposure to dioxins before conception and outcomes in offspring. The committee found that although some evidence showed effects in offspring related to maternal exposure, there was no evidence to support the transmission of effects from father to offspring.
Most new epidemiologic studies published since 2014 of children exposed to PCDD/PCDFs prenatally or perinatally, as represented by concentrations in breast milk, have assessed growth and neurocognitive development. The environmental exposures that have been studied represent continuous exposures, which may have limited applicability to veteran exposures. One study examined the children of Vietnam veterans (Grufferman et al., 2014), and several studies examined Vietnamese children born after the war. One study relied on dioxin in maternal serum measured soon after a chemical plant explosion in Seveso, Italy, to study atopy and asthma (Ye et al., 2018).
A case-control study of rhabdomyosarcoma in children of Vietnam veterans (295 cases of rhabdomyosarcoma diagnosed between 1982 and 1988, and 295 controls) found that although increased risks were observed with maternal military service (OR=2.75, 95% CI 0.71–10.62) and paternal exposure to Agent Orange (OR=1.72, 95% CI 0.55–5.41), paternal military service was not associated with an increased risk (OR=0.85, 95% CI 0.58–1.25); none of the associations were statistically significant (Grufferman et al., 2014).
A pooled analysis of three European cohorts (n=367) found that prenatal/perinatal exposure to dioxins as measured in maternal cord blood and breast milk was related to the growth of the children. Changes in z-scores for weight-for-age measurements in children 0–24 months of age were non-significantly associated with dioxin exposure (β=0.07, 95% CI –0.01– 0.14) with no sex differences. Among children assessed for BMI at age 7 years, dioxin exposure was significantly associated with increased BMI among girls (β=0.49, 95% CI 0.07–0.91), but not among boys (β= –0.03, 95% CI –0.55–0.49). The interaction between dioxin exposure and sex was significant (p=0.044). No associations between growth and dioxin exposure were found in any single cohort, and all three cohorts differed substantially (p<0.05) (Iszatt et al., 2016).
Immune function as represented by allergies and infections at ages 3.5 and 7 years was investigated among the children of 514 Japanese women. Prenatal/perinatal exposure to dioxin-like chemicals (DLCs) was measured in maternal blood and compared to IgE in cord blood as well as to reports of food allergies, eczema, wheezing, otitis media, and respiratory infections in the children. Maternal DLC concentration was related to decreased cord blood IgE in boys (β= –0.87, 95% CI –1.68– –0.06) but not in girls (β=0.27, 95% CI –0.38–0.91). Cord blood IgE levels were not associated with an increased risk of allergies or infections at either age. At 3.5 years, girls had a nonsignificantly increased risk of wheezing with DLC exposure (OR=2.52, 95% CI 0.10–60.34), but not boys (OR=0.03, 95% CI 0.00–0.94). At 7 years, wheezing was significantly increased among all DLC-exposed children (OR=7.81, 95% CI 1.42–42.94), but not for boys alone (OR=12.05, 95% CI 0.99–146.37) or girls alone (OR=4.49, 95% CI
0.33–60.44). No patterns were noted for allergy, food allergy, eczema, or infections (Miyashita et al., 2018). These findings are further supported by the lack of positive associations found between maternal dioxin exposure from the Seveso, Italy, chemical plant explosion in 1976 and asthma, hay fever, or eczema in 676 children born since 1976 and assessed in 2014 (Ye et al., 2018).
Among a cohort of 161 11-year-old children born in Hong Kong, researchers found no associations between dioxins in breast milk (representing prenatal dioxin exposure) and measures of neurocognitive and intellectual function (IQ, fine motor coordination, verbal and nonverbal reasoning, learning, and attention) (Hui et al., 2016).
Some deficits in attention in German children ages 7–11 years were associated with prenatal PCDD/PCDF exposure as measured in maternal blood and breast milk (n=117). Distractibility (measured by reaction time) was significantly associated with the levels of PCDD/PCDFs in breast milk (geometric mean ratio=1.05, 95% CI 1.0–1.11), and divided attention (measured by number of omissions) was significantly associated with the levels of PCDD/PCDFs in maternal blood (geometric mean ratio=1.47, 95% CI 1.08–2.0). Other measures of distractability and divided attention were not significantly associated with PCDD/PCDF exposure, nor were measures of performance speed or performance quality. ADHD scores were negatively associated with PCDD/PCDFs in blood, although not significantly (Neugebauer et al., 2015).
In the same German birth cohort (n=136), maternal blood levels of PCDD/PCDFs assessed during pregnancy were found to be inversely associated with sex-typical behaviors at age 9, significantly among girls (social responsiveness scale β= –10.98, 95% CI –19.43– –2.53), but not among boys. Measures of empathy or systemizing were also inversely associated with PCDD/PCDF levels. The results indicated a significant interaction effect of sex and exposure for all three measures of behavior. Maternal blood levels of PCDD/PCDFs were not above background levels (Nowack et al., 2015).
A series of publications have examined the growth and development of 241 Vietnamese children living in a high TCDD/PCDD/PCDF exposure area near Da Nang Air Base5 (Boda et al., 2018; Nishijo et al., 2014, 2015; Pham et al., 2015; Tai et al., 2016; Tran et al., 2016; Van Tung et al., 2016). Exposure was measured in cord blood at birth and in breast milk collected 28–33 days after birth. Assessments on the children were conducted at 1 month, 4 months, 1 year, and 3 years of age. Using cord blood samples available for 162 infants, investigators found decreases in estradiol and testosterone that varied by sex. Among female infants, dioxin levels in breast milk were inversely associated with testosterone levels in cord blood for certain measures (1,2,3,6,7,8-HxCDD β= –0.32, 95% CI –0.54– –0.10; TEQ-PCDD β= –0.23, 95% CI –0.45– –0.01), although this was not seen with male infants. No associations with estradiol were found (Boda et al., 2018). At 1 year of age, TCDD and total PCDD/PCDF exposure were not associated with cognitive, language, or motor function, but social-emotional scores were significantly reduced among those with high PCDD/PCDF and TCDD exposure (Pham et al., 2015). Growth and development through 3 years of age was associated with TCDD and PCDD/PCDFs in a sex-specific manner. Among boys, high PCDD/PCDF exposure was significantly associated with decreased expressive communication scores, and neither TCDD nor PCDD/PCDF exposure was associated with reduced body size. No association between PCDD/PCDFs and neurodevelopment was observed for girls, but girls with high exposure had greater head and abdominal circumferences (Tai et al., 2016).
5 Dioxin concentrations in the environment and in humans living there remained elevated in areas sprayed with herbicides, including Agent Orange, during the Vietnam War (1961–1971). Several decades after the Vietnam war, total PCDD/PCDF TEQ remained high around Da Nang Airbase, with measurements up to 365,000 pg/g. A survey of 520 nursing mothers in 2008–2009 showed that primiparae women living in hot spots (14.10 pg/g lipid) and sprayed areas (10.89 pg/g lipid) had higher total TEQ of 2,3,7,8-subsitituted dioxin concentrations in breast milk than women residing in unsprayed areas (4.09 pg/g lipid) (Tai et al., 2011).
In this Vietnamese cohort, greater rates of autistic traits were associated with high TCDD exposure in both boys and girls, but not with total PCDD/PCDFs. High exposure to total PCDD/PCDFs, but not to TCDD, was associated with significantly lower neurodevelopmental scores in boys, but not in girls (Nishijo et al., 2014). A nested case-control study within this cohort investigated changes in urinary amino acid profiles to identify alterations in amino acids that function as neurotransmitters in the developing brain. The authors noted that urinary levels of the amino acids histidine and tryptophan were significantly lower among children highly exposed to total PCDD/PCDFs, and the histidine-to-glycine ratio was significantly lower among children with high exposure to both total PCDD/PCDFs and TCDD. Histidine and the histidine-to-glycine ratio were also correlated with neurodevelopmental scores for language and fine motor skills (Nishijo et al., 2015). A follow-up assessment at 5 years of age showed decrements in movement and neurodevelopment among boys but not girls. Compared with boys with lower mean PCDD/PCDF exposure, boys with higher exposure had lower mean scores for total movement (7.7 versus 9.7, p<0.01), balance (7.3 versus 9.3, p<0.01), nonverbal index (78.9 versus 90.1, p=0.01), and pattern reasoning (5.3 versus 7.1, p=0.041) (Tran et al., 2016). The investigators suggested that TCDD acts differently than other PCDD/PCDFs and that there are sex differences in responses that last into childhood.
Nearly 45 years after the Vietnam War, Van Tung et al. (2016) assessed the birth weight of 58 babies born to mothers living in another Agent Orange–contaminated area of Vietnam compared with 62 control babies born in another community that was not sprayed with herbicides during the war. The PCDD/PCDF levels in breast milk were three to four times higher in exposed children (means 12.31 versus 3.51 pg/g lipid total TEQ PCDD/PCDFs). Two congeners were significantly correlated with birth weight (2,3,7,8-TeCDD r= –0.18; 2,3,4,7,8-PeCDF r= –0.21, both p<0.05), but no congeners were correlated with infant size at 8–9 or 12–14 weeks of age. Maternal salivary cortisol levels collected at the same time as breast milk, as a marker of an adverse intrauterine environment, were not related to PCDD/PCDF exposure.
The Volume 11 committee identified more than 20 animal studies published since 2014 on the reproductive, developmental, and generational effects of PCDD/PCDFs, all of which reported on TCDD specifically.
Two studies provided additional support for the adverse effects of TCDD on male reproduction. Rats treated with TCDD (100 ng/kg bw/day) for 15 days had reduced testicular steroidogenesis and increased oxidative stress (Dhanabalan et al., 2015). A day after exposure to TCDD (4 μg/kg bw), significant changes in gene expression and in the levels of nine genes related to AhR expression in testicular or epididymal tissue (caput, corpus, and cauda) were reported. AhR expression was greater in the epididymis than in the testes, and AhR protein levels were affected in epididymal epithelial cells but not Sertoli or Leydig cells. The authors interpreted these results as showing that TCDD can cause male infertility by affecting AhR expression and proper sperm development (Wajda et al., 2017).
Effects of prenatal exposure to TCDD were assessed in 15 studies. Wu et al. (2014) found that an oral dose of TCDD (1.6 or 8.0 μg/kg bw) on GD 15 was associated with suppressed placental vascular remodeling at GD 20. Although there was no relationship with fetal capillary density, changes in the expression of angiogenic growth factors were also observed. An in vitro study showed that TCDD decreased the viability of human umbilical cord vein and artery cells (Li et al., 2015).
Although in utero TCDD is widely used to induce cleft palate in animal models in order to study mechanisms, humans are much less sensitive to this endpoint than rodents and other species (Abbott
et al., 1999; EPA, 2012; NASEM, 2016). Therefore, although there continues to be work done with this model endpoint, this committee did not delve further into the data on TCDD-induced cleft palate in animal models. One study showed a possible role of TCDD in the development of other congenital defects. Sholts et al. (2015) reported that rats exposed to TCDD at 1,000 ng/kg on GD 11 had reduced cranium size and fluctuating asymmetry at 1 year of age.
Immune effects were observed by Ahrenhoerster et al. (2014) in an in vitro study, which showed that relatively low levels of TCDD (3 μg/kg) on GDs 0.5 and 7.5 inhibited lymphocyte differentiation by altering hematopoietic stem and progenitor cell homeostasis in such a way that B- and T-lymphocyte differentiation decreased 2.5-fold at GD 14.5. Van Esterik et al. (2015) reported that prenatal exposure to TCDD (10–10,000 pg/kg) was associated with altered immune homeostasis in mouse offspring; dams were exposed via their feed for up to 10 weeks prior to mating and through PND 14. At 1 year of age decreased fat pad and spleen weights and an increase in IL-4 production by splenic immune cells were observed in the male offspring, while increased fat pad weights and production of IFNγ were seen in the female offspring. Laiosa et al. (2016) showed that TCDD’s effects on the developing hematopoietic system are mediated through fetal activation of the AhR.
A variety of endocrine effects have been studied in association with prenatal TCDD exposure. Takeda et al. (2014b) examined the impact of prenatal TCDD exposure on hormones and sexual maturity. TCDD (0.2–20 μg/kg) administered on GD 12 reduced the pituitary expression of mRNAs for LH, FSH, and glycoprotein hormone α-subunit (αGSU) and of growth hormone measured on GD 18, PND 2, and PND 4 in mice. TCDD also reduced the gonadal expression of steroidogenic proteins and had adverse effects on litter size and pup body weight. Mouse strains that express high or low AhR affinity for dioxins respond to TCDD exposure differently, which emphasizes the role of AhR activation in causing these effects (Takeda et al., 2014b). In another study by Takeda et al. (2014a), 1 μg TCDD/kg on GD 15 affected the levels of hypothalamic gonadotropin-releasing hormone (GnRH) and sexual behaviors in rats. Deficits in GnRH levels and sexual behaviors were remedied by equine chorionic gonadotropin at GD 15, and an intracerebroventricular infusion of GnRH after sexual maturity restored sexual behaviors (Takeda et al., 2014a). The effect of the TCDD was subsequently shown to be due, in part, to an altered energy status (decreased ATP) in the fetal whole brain and hypothalamus (Takeda et al., 2017).
Schneider et al. (2014) studied the effect of prenatal TCDD (5 μg/kg, po, on embryonic day 15.5) on mouse prostate development and found that ventral prostatic budding in the embryonic urogenital sinus had been prevented as a result of β-catenin signaling at embryonic day 16.5.
Sugai et al. (2014) studied the effects of prenatal TCDD exposure in mice (3 μg/kg on GD 12.5) and then assigned the offspring to either regular or high-calorie diets. Unexposed mice eating high-calorie diets were obese and showed signs of lipid dysregulation at 26 weeks, whereas TCDD-exposed mice eating the same high-calorie diet did not, suggesting that TCDD attenuated the effects of a high-calorie diet on obesity and related adult diseases.
Brain development was the subject of three recent studies. Kobayashi et al. (2015) showed that an oral dose of TCDD (20–5,000 ng/kg) at 8.5 days post coitus had significant impacts on brain development at embryonic day 18.5. Mice exposed to the highest dose showed reduced glial fibrillary acidic protein (GFAP) immunoreactivity in the dentate gyrus and hippocampal fimbria and reduced immunoreactive fimbria; all doses were associated with decreased proliferating cell nuclear antigen-positive cells. Rats administered TCDD and TBDD (2,3,7,8-tetrabromodibenzo-p-dioxin) at 0, 200, or 800 ng/kg on GD 15 exhibited impaired paired-associated learning in adulthood for both compounds at 200 ng/kg, but not at 800 ng/kg. Similarly, rats exposed to 200 ng/kg of TCDD or TBDD showed greater anxiety-like behaviors than controls or highly exposed animals (Kakeyama et al., 2014). Safe and Luebke (2016) considered auditory neuropathy among three strains of mice exposed prenatally to TCDD (500 ng/kg) on
embryonic day 12 by oral gavage. At 1.5 months of age, the cochlear threshold sensitivity was significantly elevated in mice with certain AhR alleles. The authors concluded that prenatal TCDD exposure may be a risk factor for auditory neuropathy.
Muenyi et al. (2014) found that prenatal TCDD (10 μg/kg on GD 12) accelerated the development of abnormal epidermal permeability barrier with acanthosis and hyperkeratosis and also epidermal tight junction function as measured through PND 1 in mice. Such changes have been linked to atopic diseases.
Adult mice (up to 4 months old) exposed to a single dose of TCDD in utero (13 μg/kg on day 13 post-coitus) and slow release implants of testosterone and 17 β-estradiol at 8 weeks old had increased bladder, kidney, prostate, and seminal vesicle weights. They also had more hydronephrosis, prostate epithelial cell proliferation, thickened prostate periductal smooth muscle, and altered collagen fiber distribution in their prostate and bladder tissue. Mice exposed to in utero TCDD but not exogenous hormones had few changes from controls. The authors interpreted these data as showing that early life exposure to TCDD alters susceptibility to adult urinary tract dysfunction (Ricke et al., 2016).
Studies have also reported effects in subsequent generations. ATSDR (2012) reported examples of intergenerational effects in hamsters and mice in which reproductive effects were observed in F2 or F3 offspring after the gestational exposure of the F0 female, and transgenerational effects have been the subject of a recent systematic review. That review of the evidence of inherited transgenerational effects found that although there is no human evidence concerning the transgenerational effects of dioxin, there have been eight studies in animals, from five groups of investigators, that examined the transgenerational effects of gestational dioxin exposure on offspring, including its effects on growth and development, its reproductive effects in offspring, and its hepatic, immune, and renal effects (Walker et al., 2018). Four studies reported transgenerational effects on male reproductive endpoints (Bruner-Tran et al., 2014; Manikkam et al., 2012a,b; Sanabria et al., 2016). Although dioxin exposure had no reported effects on the traditional measures of male reproductive endpoints (sperm concentration and motility, apoptosis of testes germ cells, prostate or testes disease, LH concentration, reproductive organ weights, or puberty), one study found an increase in abnormal sperm morphology (Bruner-Tran et al., 2014), and two studies from the same researchers found age-dependent changes in testosterone concentrations (decreased at 3–4 months of age and increased at 12 months of age) (Manikkam et al., 2012a,b). Five studies reported on female reproductive outcomes (Bruner-Tran and Osteen, 2011; Bruner-Tran et al., 2016; Manikkam et al., 2012a,b; Nilsson et al., 2012). One group of researchers reported decreased primordial follicles and increased ovarian cysts (Manikkam et al., 2012a,b; Nilsson et al., 2012) and early onset of puberty (Manikkam et al., 2012a,b) associated with dioxin exposure. One study investigated adenomyosis (similar to endometriosis) and found increased uterine microvessel density in the offspring of exposed rats (Bruner-Tran et al., 2016). Other endpoints that have been studied but were not affected by exposure include fertility, sex ratio, tumor development, and reproductive organ weight. Two studies reported growth and developmental outcomes (Manikkam et al., 2012a,b) and found that dioxin was associated with an increased body weight among males at weaning and at 12 months of age, but no effect was observed in females (Manikkam et al., 2012a,b). Additionally, kidney disease (Manikkam et al., 2012b), effects on liver function (Ma et al., 2015), and increased immune response (Bruner-Tran et al., 2014) were all observed in males. Inconsistencies across studies (heterogeneity in endpoints selected and in the measurement of endpoints), a lack of information about the age of animals when observations were made, and a lack of analysis using the litter as the unit of analysis limit confidence in these findings.
Since 2014, three groups of researchers have reported on transgenerational effects observed in rodents, all of them cited in the systematic review described above (Bruner-Tran et al., 2014, 2016; Ma et al., 2015; Sanabria et al., 2016; Walker et al., 2018).
Ma et al. (2015) studied TCDD’s epigenetic effect on the IGF2 gene in rats through three generations. Only the F0 dams were treated, and the F2 generation was used for production of the F3 but was not further evaluated. Prenatal exposure (200 or 800 ng/kg TCDD on GD 2-14) was associated with increased IGF2 expression, hypermethylation of the IGF2 imprinted control region, hypomethylation of the differentially methylated region 2, and altered expression of DNA methyltransferases in F1 and F3 mice compared with control mice.
In studies on mice, Bruner-Tran and colleagues (2017) observed an endometriosis-like phenotype and other reproductive effects in the female offspring (F1–F4) after in utero exposure to TCDD, and they reported reduced endometrial progesterone receptor (PR) protein expression in uterine tissue in F1–F3 mice. They also reported partial methylation of the PR in 60% of tissues from F1 females and 40% of tissues of F3 females, leading them to conclude that there was transgenerational inheritance of epigenetic silencing of the PR.
In reporting results from a mouse model of TCDD-induced palatogenesis, researchers suggested that TCDD caused the upregulation of DNA methyltransferase 3a in fetal palate tissue and increased global DNA methylation and caused palate malformation. However, the association was only observed in fetal mice at GD 13.5, but not GD 14.5 or 15.5 (Wang et al., 2017a).
Synthesis and Conclusions
The effects of PCDD/PCDFs have been studied extensively, with much attention given to the effects of exposures in Vietnam veterans. Although some studies have shown reproductive or developmental effects at very high doses, such as those experienced in the Seveso, Italy, accident, those associations have not been shown—or, in some cases, even studied—in other populations. Therefore, associations at lower levels of exposure cannot be assessed. For example, studies in Seveso found associations between women with TCDD levels >100 ppt and time to pregnancy (Eskenazi et al., 2010), but other studies of similar endpoints have not been conducted, so it is not possible to assess the consistency of this association (NASEM, 2016). Similarly, the developmental effects of early life exposure in the Seveso cohort included reduced semen quality among adult men who had been breast-fed by exposed mothers, but because of questions about the reliability of the data and covariates, these results require further validation or support (IOM, 2014; Mocarelli et al., 2011).
Since the last VAO review of the literature on the health effects of dioxins (NASEM, 2016), numerous studies have reported on dioxin’s effects on male fertility (Den Hond et al., 2015; Galimova et al., 2015), endometriosis (Martinez-Zamora et al., 2015), estradiol production (Baldridge et al., 2015), the sex ratios of children (Mannetje et al., 2016), adverse birth and pregnancy outcomes (Pan et al., 2015; Wesselink et al., 2014), and birth defects (Noel et al., 2015), but the Volume 11 committee found no consistent patterns for reproductive outcomes when considering the entirety of the evidence base, including reviews by the National Academies (2016), EPA (2012), and ATSDR (2012).
Studies of adverse pregnancy outcomes show some evidence of effects of prenatal exposure, but there is much variability among the studies. Kobayashi et al. (2017) provides a mechanistic rationale for differences among exposed individuals with certain genotypes.
Developmental outcomes that have been studied include rhabdomyosarcoma in the children of Vietnam veterans (Grufferman et al., 2014), infant and child growth (Iszatt et al., 2016; Tai et al., 2016; Van Tung et al., 2016), allergies and infections in childhood (Miyashita et al., 2018), language development (Caspersen et al., 2016), neurocognitive development and intellectual function (Hui et al.,
2016), attention measures (Neugebauer et al., 2015), sex-typical behaviors (Nowack et al., 2015), and neurodevelopment (Nishijo et al., 2014, 2015; Pham et al., 2015; Tran et al., 2016). Some studies support associations between dioxin exposure and effects that are mediated by genotype (Kobayashi et al., 2017) and sex (Miyashita et al., 2018; Nishijo et al., 2014, 2015; Nowack et al., 2015; Tai et al., 2016; Tran et al., 2016). The results of these studies are inconsistent, and few outcomes were studied in more than one population.
The Volume 11 committee identified more than 20 animal studies on the reproductive and developmental effects of prenatal TCDD exposure. The studies generally support the conclusions reached in recent reviews by EPA (2012), ATSDR (2012), and the National Academies (2016) that TCDD may result in adverse male reproductive effects and a variety of developmental effects, including endocrine, immune, and neurodevelopmental changes. Animal models point to oxidative stress and AhR-activation as the mechanisms by which TCDD causes these effects (Dhanabalan et al., 2015; Laiosa et al., 2016). However, the doses used in the animal studies are much higher than expected human exposures. Additionally, there is great variability in species sensitivity to PCDD/PCDFs, depending on AhR binding ability, and humans are much less sensitive than laboratory rodents (ATSDR, 2012; EPA, 2012; NASEM, 2016). Thus, the relevance to human health outcomes of the large evidence base derived from animal studies is unclear.
Although there is some evidence from animal studies that maternal exposure to dioxins may result in epigenetic effects and transgenerational health effects (ATSDR, 2012; Bruner-Tran et al., 2014, 2016; NASEM, 2016; Walker et al., 2018), there is no evidence of such effects in humans.
There is a large evidence base describing human health effects in populations exposed occupationally, environmentally, and as a result of the Vietnam war and a similarly robust evidence base in animal models (many thousands of publications have been reviewed by the VAO series), but there are few consistencies across cohorts, and the ability to use animal studies to support associations noted in humans studies is limited by the ability to extrapolate outcomes in animals to humans.
The Volume 11 committee discussed the association between dioxins and reproductive effects at length. The committee was challenged by the large volume of literature describing a wide range of effects in particular populations with very high levels of exposure (e.g., Seveso). The studies included a number of effects, but few of them examined the same effects or found the same results. The committee’s discussion differed in focus from those of other committees, such as VAO and prior Gulf War and Health committees. Those committees focused on diagnosed clinical outcomes such as fertility and endometriosis. Although those committees considered effects such as abnormal semen parameters, these effects alone did not justify a conclusion of adverse clinical outcomes. By contrast, the Volume 11 committee, given its task of assessing generation effects in veterans and their descendants and also assessing animal models, considered the broader range of biological effects, such as changes in semen parameters, to be indicative of potential adverse reproductive outcomes regardless of the level of exposure. The committee finds that, taken as a whole, the literature it reviewed above for both humans and animals demonstrates that biological effects on the reproductive system may be possible at high levels of exposure and represents a residual level of concern. Therefore, the committee identified reproductive effects of dioxins as a major priority for future monitoring and research as discussed in the following chapters.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between exposure to PCDDs/PCDFs and reproductive or developmental effects.
TABLE 6-7 Summary of Reproductive and Developmental Effects of Polychlorinated Dibenzodioxins
|Den Hond et al. (2015)||Case-control
40 cases of male infertility and 80 controls, recruited from fertility clinics in Belgium.
|PCDD/PCDFs measured in serum.||Semen parameters and sex hormones in serum.
No association between PCDD/PCDFs and subfertility (OR=1.59, 95% CI 0.96–2.65) or hormone levels.
|Faure et al. (2014)||Ecological
251 men living near a municipal waste incinerator in France where exhaust controls were implemented in 1998, semen samples collected in 2001–2007.
|Modeled exposure to PCDD/PCDFs from incinerator emissions by year as surrogate for exposure.||Semen parameters by year.
Abnormal mid-piece: β= –0.010, 95% CI –0.018– –0.002; multiple anomalies index: Symbol = –0.013, 95% CI –0.023––0.004.
Other parameters not significant.
|Galimova et al. (2015)||Case-control
105 men with pathospermia and 63 men with normal sperm seen at a fertility clinic, and 49 fertile controls, living near a chemical plant in Russia.
|PCDD/PCDFs in semen.||Semen parameters.
Higher PCDD/PCDF content TEQ among infertile normospermia and pathospermia men than fertile men (total TEQ pg/g lipids: 466.6, 482.1, 212.5, respectively).
No statistical testing conducted.
|Martinez-Zamora et al. (2015)||Case-control
30 cases of deep infiltrating endometriosis and 30 controls, in Catalonia, Spain.
|17 PCDD/PCDFs measured in adipose tissue.||Significantly higher median TEQ among endometriosis cases than controls for
TCDD: OR=1.41, 95% CI 1.12–2.10;
1,2,3,7,8-PeCDD: OR=1.82, 95% CI 1.36–7.14;
2,3,4,7,8-PeCDF: OR=2.13, 95% CI 1.97–6.42.
Also, 4 PCB congeners were significantly associated with endometriosis.
All other congeners not significant.
|Adverse Pregnancy Outcomes|
|Mannetje et al. (2016)||Cohort
355 children born to parents (127 fathers and 21 mothers) employed at a chemical plant in New Zealand.
|TCDD in parental serum back-calculated to estimate exposure at the time of birth.||Sex ratio.
Any maternal exposure not significant. Paternal exposure ≤4pg/g: SR=0.60 (ref); 4–20 pg/g SR=0.62, OR=1.0, 95% CI 0.50–2.02; 20–100 pg/g SR=0.47, OR=0.52, 95% CI 0.29–0.92; ≥100 pg/g SR=0.46, OR=0.45, 95% CI 0.23–0.89; p trend=0.007.
TABLE 6-7 Continued
|Wesselink et al. (2014)||Cohort
1,211 pregnancies between 1976 (chemical plant explosion) and 2009, in Seveso, Italy.
|TCDD in serum collected in 1976 or 1996.||Adverse birth outcomes.
TCDD measured in 1976 or estimated at birth was not related to spontaneous abortion, fetal growth, gestational length, or birth weight.
|Kobayashi et al. (2017)||Cohort—Hokkaido Study
421 pregnant women in Japan.
|17 PCDD/PCDFs measured in maternal blood in third trimester or within a week of birth; AHR, CYP1A1, and GSTM1 genotypes.||Birth size.
No association between AHR and CYP1A1, and GSTM1 non-null genotypes and PCDD/Fs on birth weight, height, or head circumference. Among women with the GSTM1 null genotype, PCDD/Fs associated with reduced birth weight (β= –345, 95% CI –584– –105).
|Noel et al. (2015)||Ecological
All live births in Guam 1970–1989 (N=49,841); 121 infant deaths due to congenital anomalies.
|Maternal residence in areas sprayed with Agent Orange (high risk) versus areas not sprayed (low risk).||Infant mortality due to congenital anomalies.
Deaths associated with maternal residence in a high-risk area: β=2.02, 95% CI 0.08–3.96.
|Gufferman et al. (2014)||Case-control
295 cases of rhabdomyosarcoma and 295 matched random controls, from U.S. births 1982–1988.
|Parental military service and reported Agent Orange exposure.||Rhabdomyosarcoma.
Maternal military service: OR=2.75, 95% CI 0.71–10.62.
Paternal military service: OR=0.85, 95% CI 0.58–1.25.
Paternal Agent Orange exposure: OR=1.72, 95% CI 0.55–5.41.
Not significant for paternal deployment periods, exposure to radiation, radar or microwaves, and nuclear, chemical, or biological weapons.
Adjustment for drug use, but no assessment for alcohol or smoking.
|Iszatt et al. (2016)||3 Cohorts
96 Belgian children, 207 Slovak children, and 64 Norwegian children.
|PCDD/PCDFs in cord blood and breast milk.||Infant growth 0–24 months and BMI at 7 years. Change in infant weight-for-age z-scores β=0.07, 95% CI –0.01– 0.2.
BMI at 7 years old for girls β=0.49, 95% CI 0.07–0.91; not significant for boys. Significant interaction dioxin × sex (p=0.044).
Significant differences between cohorts for most demographic and outcome measures.
|Miyashita et al. (2018)||Cohort-Hokkaido Study
514 pregnant women at a hospital in Japan.
|Dioxin-like compounds in maternal blood and in cord blood||Cord blood IgE and allergies at ages 3.5 and 7 years.
Maternal dioxin-like compounds associated with decreased cord blood IgE in boys (β= –0.87, 95% CI –1.68– –0.06); not significant for girls.
At 3.5 years, reduced risk of wheezing associated with dioxin-like compounds exposure among boys (OR=0.03, 95% CI 0.0–0.94); not significant for girls.
At 7 years, risk of wheezing increased for all children (OR=7.81, 95% CI 1.42–42.94); not significant for boys and girls individually.
No association btw cord blood IgE and allergies at 3.5 or 7 years of age. No associations with food allergy, eczema, or infections.
|Ye et al. (2018)||Cohort-Seveso Women’s Health Study
677 children born to TCCD exposed mothers 1976–2016.
|TCDD in maternal serum measured following the explosion of a chemical plant in 1976.||Atopic conditions (eczema, hay fever, asthma) in children ascertained by mothers’ self-report.
Eczema: RR=0.63, 95% CI 0.40–0.99; Hay fever: RR=0.99, 95% CI 0.76–1.27; and Asthma: RR=0.93, 95% CI 0.61–1.40.
Adjustments for maternal age at explosion, atopic history, age at pregnancy, BMI closest to birth year, smoking during pregnancy, birth order, child sex, child age, and primary wage earner education.
|Hui et al. (2016)||Cohort 161 mother–child pairs in Hong Kong.
|Dioxins in breast milk measured 2–6 weeks postpartum.||Neurocognitive development at 11 years old. No associations found with full-scale IQ, verbal comprehension, perceptual reasoning, processing speed, working memory, learning, intrusion error, preservation error scores, or tests of attention. No sex × breast-feeding interactions note.|
TABLE 6-7 Continued
|Neugebauer et al. (2015)||Cohort-Duisburg Birth Cohort
117 mother–child pairs in Germany.
|PCDD/PCDFs in maternal blood at 32nd week of pregnancy and breast milk collected within 2 weeks of birth.||Attention performance (using KITAP, a computerized test battery for attentional performance for children) and attention behavior (FBB-ADHD, a parent rating scale for ADHD) in school-age children.
Distractability: slower reaction time associated with total PCDD/PCDFs TEQ in breast milk (geometric mean ratio=1.05, 95% CI 1.0–1.11), but not omissions or false alarm rates.
Divided attention: increased number of omissions associated with PCDD/PCDFs TEQ in maternal blood (geometric mean ratio=1.47, 95% CI 1.08–2.0), but not false alarm rates or reaction time.
Negative but not significant associations between maternal blood or breast milk PCDD/PCDFs with performance speed or quality or ADHD.
|Nowack et al. (2015)||Cohort-Duisburg Birth Cohort
136 mother–child pairs in Germany.
|PCDD/PCDFs in maternal serum during pregnancy.||Empathizing, systemizing, and autistic traits at 9–10 years.
Empathizing and systemizing quotients not related to PCDD/PCDF exposure in boys or girls. Social responsiveness scale for autistic traits associated with PCDD/F exposure among girls (β= –10.98, 95% CI –19.43––2.53) but not boys (β= –2.37, 95% CI –9.18–4.45).
|Van Tung et al. (2016)||Case-control
58 mother–infant pairs from an Agent Orange contaminated region in Vietnam and 62 pairs from a region with no contamination.
|PCDD/PCDFs and TCDD in breast milk and maternal cortisol levels. Infants were 7–16 weeks old at time of sampling.||Birth weight.
Average PCDD/PCDFs higher in breast milk of exposed mothers than controls (12.3 versus 3.5 pg/g lipid, p<0.001).
Two congeners correlated with birth weight (TCDD r= –0.18; 2,3,4,7,8-PeCDF r= –0.21; both p<0.05).
Mean cortisol was higher in mothers of low birth weight babies than normal weight (6.52 versus 3.2, p<0.05), not related to total PCDD/PCDFs or specific congeners.
No difference in body size at 8–9 weeks or 12–14 weeks of age.
|Boda et al. (2018)||Cohort
162 mother–infant pairs in Da Nang, Vietnam.
(See Pham et al., 2015)
|PCDD/PCDFs in breast milk collected 1 month after birth.||Estradiol and testosterone in cord blood.
No association between estradiol and PCDD/PCDFs in breast milk.
Decreased testosterone associated with PCDD/PCDFs in breast milk among girls (1,2,3,6,7,8-HxCDD β= –0.32, 95% CI –0.54– –0.10; and TEQ-PCDD β= –0.23, 95% CI –0.45––0.01, other congeners not significant).
No dose–response relationship found.
No association among boys.
|Pham et al. (2015)||Cohort
214 mother–infant pairs in Da Nang, Vietnam.
|PCDD/PCDFs in breast milk collected 28–33 days after birth (low: ≤7.4 pg/g lipid TEQ; mild: 7.4–11.5; moderate: 11.5–17.6; high: ≥17.6).
Daily dioxin intake (low: ≤49.8 pg TEQ/kg/day; mild: 49.8–77.8; moderate: 77.8–118.2; high: ≥118.2).
TCDD (low: ≤ 0.9 pg/g lipid; mild: 0.9–1.8; moderate: 1.8–3.5; high: ≥3.5).
|Neurodevelopment at 1 year (Bayley-III scores).
Mean cognitive, language, and motor scores not related to PCDD/PCDFs or TCDD exposure. Cognitive scores were significantly lower in the low daily dioxin intake exposure group than mild, moderate, or high groups (mean scores 97.4 versus 102.8, 102.1, and 102.4, p<0.042 for all three).
Social emotional scores significantly lower among those with high PCDD/PCDFs exposure (89.9 versus 93.4, p=0.049) and high TCDD exposure (88.4 versus 90.8, p=0.019) than low exposure. No association for adaptive behavior scores.
|Tai et al. (2016)||Cohort
217 mother–infant pairs in Da Nang, Vietnam.
(See Pham et al., 2015)
|PCDD/PCDFs in breast milk collected 28–33 days after birth.||Neurodevelopment at 3 years (Bayley-III scores and growth expressed by z-scores to show marginal means).
Boys with high PCDD/PCDF exposure had low mean expressive communication scores (8.4 versus 9.0, p=0.03); boys with high TCDD exposure had low mean composite motor scores (94.9 versus 100.5, p=0.049) and gross motor scores (9.1 versus 10.4, p=0.036).
Boys with high PCDD/PCDF exposure had reduced mean weight growth (–0.38 versus –0.08, p=0.01) and BMI (–0.43 versus –0.14, p=0.02).
No neurodevelopmental differences among girls. Girls with high TCDD exposure had greater mean head circumference growth (0.18 versus –0.28, p=0.00), and high PCDD/PCDFs exposures had greater chest circumference growth (0.17 versus –0.12, p=0.02).
TABLE 6-7 Continued
|Nishijo et al. (2014)||Cohort
153 mother–infant pairs in Da Nang, Vietnam.
(See Pham et al., 2015)
|PCDD/PCDFs in breast milk collected 28–33 days after birth (high versus low exposure: ≥ or <3.5 pg/g fat). TCDD exposure (high versus low: ≥ or <17.9 pg/g fat).||Autistic traits using the Autism Spectrum Rating Scale and neurodevelopment (Bailey-III scales) at 3 years.
Children with high TCDD exposure had higher mean autism score than low exposure (boys: 61.5 versus 55.0, p=0.042; girls: 58.2 versus 54.3, p=0.012) but no difference in neurocognitive scores.
Children with high PCDD/PCDF exposure had similar autism scores as low exposure children, but highly exposed boys had lower cognitive scores (93.0 versus 96.8, p=0.023).
|Nishijo et al. (2015)||Nested case-control
26 children in Da Nang, Vietnam.
(See Pham et al., 2015)
|PCDD/PCDFs in breast milk collected 28–33 days after birth in 4 groups: (1) low PCDD/PCDFs and well-developed controls; (2) low PCDD/PCDFs and poorly developed controls (3) high PCDD/PCDFs and low TCDD; and (4) high PCDD/PCDFs and high TCDD.||Urinary amino acids in children with neurodevelopmental deficits at 3 years.
Mean histidine levels of group 4 significantly lower than group 1 (18.58 versus 187.93, p=0.004) and tryptophan (0.69 versus 2.99, p=0.007).
Ratio of histidine/glysine significantly lower in groups 4, 3, and 2, than 1 (1.37, 2.73, 1.70 versus 7.5; p<0.006 for all). Histidine levels associated with neurodevelopmental scores (cognitive β=11.8; receptive communication β=2.7; expressive communication β=1.5; language β=18.7; fine motor β=2.1; motor β=13.4; all p<0.05) same pattern shown by histidine/glysine ratio.
|Tran et al. (2016)||Cohort
176 mother–infant pairs in Da Nang, Vietnam.
(See Pham et al., 2015)
|PCDD/PCDFs in breast milk collected 28–33 days after birth.||Cognitive development (using the Kaufman Assessment Battery for Children, 2nd ed.) and motor coordination (using the Movement Assessment Battery for Children, 2nd ed.) at 5 years.
Boys with high exposure had lower mean scores than low exposure groups for total movement (7.7 versus 9.7, p<0.01) and balance scores (7.3 versus 9.3, p<0.01); nonverbal index scores (78.9 versus 90.1, p=0.01); and pattern reasoning scores (5.3 versus 7.1, p=0.041).
No associations between exposure and motor coordination or cognitive development in girls.
NOTE: ADHD=attention deficit–hyperactivity disorder; AhR=aryl hydrocarbon receptor; Bayley-III=Bayley Scales of Infant and Toddler Development, Third Edition; BMI=body mass index; CI=confidence interval; CYP1A1=cytochrome P450 1A1; GSTM1=glutathione S-transferase mu 1; IgE=immunoglobulin E; IQ=intelligence quotient; OR=odds ratio; PCB=polychlorinated biphenyl; PCDD=polychlorinated dibenzo-p-dioxin; PCDF=polychlorinated dibenzofuran; RR=relative risk; SR=sex ratio; TCDD=2,3,7,8-tetrachlorodibenzo-p-dioxin; TEQ=toxic equivalent.
The Volume 11 committee identified a few studies pertaining specifically to combustion engine exhaust that have been published since the last review by the Volume 3 committee in 2004 (see Table 6-8). Exhaust is a mixture that includes PM, PAHs, PCDD/PCDFs, and VOCs in addition to numerous other compounds that may result from the combustion of fuels such as gasoline, diesel, or jet fuel. No studies were conducted in veteran populations, and there were no studies that investigated possible generational effects.
Reproductive Effects in Men and Women
In an epidemiologic study, Guven et al. (2008) provided a cross-sectional look at the semen quality of 38 toll collectors and 35 office workers at the same company. Toll collectors had lower sperm density (p=0.002) and motility (p=0.003) and higher percentages of abnormal sperm morphology (head abnormalities p=0.01, mixed abnormalities p=0.04) than the office workers. No data relating these sperm effects to fecundity were reported.
In an animal study, gasoline exhaust was shown to have reproductive effects in male mice, including significant dose-dependent increases (p<0.05) in malondialdehyde and carbonyl protein in testes and DNA damage as measured by comet assay. Significant dose-dependent decreases (p<0.05) in superoxide dismutase and glutathione peroxidase in testicles were also reported (Che et al., 2009). Biodiesel fuel exhaust particles caused adverse effects on male reproductive function in mice, including having significant impacts (p<0.05) on sperm integrity, testosterone levels in testes, LH in serum, sperm DNA fragmentation, inflammatory cytokines in serum and testes, interstitial edema, degenerating spermatocytes, and dystrophic seminiferous tubules with arrested spermatogenesis (Kisin et al., 2015).
Adverse Pregnancy Outcomes
No new epidemiologic studies that specifically examined exposure to exhaust reported on preterm birth, birth weight, or other birth outcomes. Weldy et al. (2014) reported that diesel exhaust air pollution promotes significant placental injury (p<0.05), as manifested by hemorrhage, vascular compromise, focal necrosis, embryo resorption, inflammation, and oxidative stress in mice exposed in utero.
A prospective cohort study in eastern Massachusetts examined the effect of living near gasoline filling stations on growth and weight from birth through early childhood (Fleisch et al., 2015, 2017). The study (Project Viva) recruited more than 2,000 pregnant women from a medical practice and estimated third-trimester exposure based on modeled PM2.5, carbon black, traffic density, and residence distance to roadway. The highest levels of third-trimester carbon black and neighborhood traffic density were associated with lower mean fetal growth (–0.13 units, 95% CI –0.25– –0.01) and rapid mean weight gain to 6 months of age (0.25 units, 95% CI 0.01–0.49) (Fleisch et al., 2015). Living within 50 m (compared with greater than 200 m) of a major roadway at birth was significantly associated with greater BMI 0.3 kg/m2 (95% CI 0.0–0.7), 1.7 cm larger waist circumference (95% CI 0.6–2.8), 1.9 mm larger sum of skinfold thickness (95% CI 0.6–3.2), and 40.7% higher leptin concentration (95% CI 5.2–88.1) in early childhood (mean age 3.3 years), but other indicators of exposure were not related to growth (Fleisch
et al., 2017). This study did account for changes in address during the follow-up period, but it did not differentiate between prenatal or postnatal exposure.
Mice exposed prenatally to exhaust have demonstrated a variety of effects. Exposure to diesel exhaust PM2.5 particles can affect mothering by dams, resulting in altered energy metabolism in offspring. Prenatal mothering by exposed dams increased the mass of epididymal adipose tissue through hyperplasia, and postnatal mothering by diesel exhaust particle–exposed dams increased the mass of all tested fat pads through hypertrophy (Chen et al., 2017b). Reduced primordial and primary follicles in the offspring of diesel-exhaust-exposed mice were reported by Ogliari et al. (2013). Thirtamara Rajamani et al. (2013) reported that mice demonstrated increased locomotor activity, rearing behaviors, and repetitive self-grooming but no changes in social interaction or anxiety-like behaviors. Weldy et al. (2014) reported increased body weight, altered blood pressure, and increased susceptibility to pressure-overload-induced heart failure among adult mice with prenatal exposure to diesel exhaust.
One review inventoried the developmental effects of prenatal diesel exhaust, diesel exhaust particles, and carbon black that had been reported in 39 toxicologic studies (Ema et al., 2013). Male reproductive system effects included changes in the weight of the testes and accessory reproductive organs, in serum hormone levels, in spermatogenesis, in the histopathology of the testes, and in gene expression for male gonadal development. Female offspring had reproductive effects, including changes in postnatal behavior, in the histopathology of the cerebral cortex and hippocampus, in the development of the central dopaminergic system, in monoaminergic neurochemistry in the brain, and in the expression of steroid hormone–related and thyroid hormone–related genes in the brain. The reported changes in immune response in offspring were sometimes contradictory with regard to prenatal sensitization, but the effects included decreased thymus and spleen weight, an accelerated production of IgE against pollen, modified expression of immune-related genes, increased allergic susceptibility, altered inflammatory indices in the lung, and increased airway hyper reactivity. The authors noted that the data on genotoxicity were inconsistent (Ema et al., 2013).
Synthesis and Conclusions
Several studies have examined the reproductive or developmental effects of exposure to exhaust. The effects on sperm have been reported in a cross-sectional study of toll collectors, which showed adverse effects on sperm, and in studies of the effects of biodiesel and diesel exhaust in mice and rats (Che et al., 2009; Guven et al., 2008; Kisin et al., 2015). Prenatal exposure to high levels of exhaust from traffic has been associated with decreased growth at 6 months of age and a high BMI in childhood, although the analyses did not adjust for postnatal exposures (Fleisch et al., 2015, 2017). Studies in mice have found associations between prenatal exposures and adverse pregnancy outcomes as well as obesity, behavior changes, and reduced primordial germ cells and primary follicles in offspring (Chen et al., 2017b; Ogliari et al., 2013; Thirtamara Rajamani et al., 2013; Weldy et al., 2013). These studies demonstrate in utero susceptibility to exhaust. No studies examined effects following parental preconception exposure to exhaust. These studies on exhaust are limited by the lack of adjustment for PM, a major component of exhaust, in most studies.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between exposure to exhaust and reproductive or developmental effects.
TABLE 6-8 Summary of Reproductive and Developmental Effects of Exhaust
|Guven et al. (2008)||Cross-sectional
38 healthy toll collectors and 35 healthy controls working as office personnel at the same company.
|Exposure categorized by job title.||Semen quality.
Toll collectors had lower sperm density (p=0.002) and motility (p=0.003) and higher percents of abnormal morphology (head abnormalities p=0.01, mixed abnormalities p=0.04).
|Fleisch et al. (2015)||Prospective cohort (Project Viva)
2,115 pregnant women from a medical practice in Boston.
|Estimated third-trimester exposure based on modeled PM2.5, carbon black, traffic density, and residence distance to roadway. Exposure estimates accounted for changes in address.||Fetal growth and infant growth at 6 months Black carbon and PM2.5 not related to outcome measures.
Lower fetal growth and rapid weight gain to 6 months of age associated with the highest quartiles of exposure for third trimester carbon black (OR=2.0, 95% CI 0.82–4.84) and traffic density (OR=3.01, 95% CI 1.08–8.44).
Adjustments for household income, paternal BMI, maternal education, race/ethnicity, prepregnancy BMI, gestational weight gain, gestational diabetes, smoking, gestational age. Did not adjust for any postnatal factors.
|Fleisch et al. (2017)||Prospective cohort (Project Viva)
1,418 pregnant women from a medical practice in Boston.
|Estimated prenatal through mid-childhood exposure based on modeled PM2.5, carbon black, traffic density, and residence distance to roadway. Exposure estimates accounted for changes in address after birth.||Early and mid-childhood BMI and metabolic measures (median ages 3.3 and 7.7).
Prenatal exposures to carbon black, PM2.5, or high traffic density not associated with increased BMI or cardiometabolic measures in early or mid-childhood.
Living within 50 m of a major roadway at birth was significantly associated with increased BMI, waist circumference, skinfold thickness, and leptin change in childhood.
NOTE: BMI=body mass index; CI=confidence interval; OR=odds ratio; PM=particulate matter.
Several petroleum-based fuels were used by the U.S. military in the Persian Gulf region during Operation Desert Shield and Operation Desert Storm, including gasoline, kerosene, diesel, and the jet-propulsion fuels JP-4, JP-5, and JP-8. Those fuels powered aircraft, ground vehicles, tent heaters, and cooking stoves. They were also used for less conventional purposes, such as suppressing sand, cleaning equipment, and burning trash. Military personnel serving in the Gulf War theater of operations could have been exposed to uncombusted fuels, the combustion products from burning those fuels, or a combination of uncombusted and combusted materials (IOM, 2005).
The Volume 3 committee found several studies of parental exposure to fuel that examined adverse reproductive outcomes which may result from genetic alterations in either sperm or egg, including infertility, spontaneous abortion, childhood leukemia, neuroblastoma, and Prader-Willi syndrome. The committee found that it was difficult to come to conclusions about the health outcomes because of the small number of studies that were available for each health outcome, the possibility of recall bias, and the lack of specificity of exposure to the agents of concern in this report. Thus, the Volume 3 committee concluded that there was inadequate/insufficient evidence to determine whether an association exists between exposure to fuels and adverse reproductive or developmental outcomes, including infertility, spontaneous abortion, childhood leukemia, central nervous system tumors, neuroblastoma, and Prader-Willi syndrome (IOM, 2005).
The Volume 11 committee’s assessment of the reproductive and developmental effects of fuels was augmented by a 2017 ATSDR profile that described the extensive literature on jet fuels. The committee identified only one additional study on jet fuels not assessed by ATSDR (Tracey et al., 2013). Because the ATSDR profile is recent and broad, the committee did not re-review those materials and instead cites the profile to describe the reproductive and developmental effects of jet fuels.
The remainder of the literature in this section pertains to exposure to automotive gasoline and additives (see Table 6-9). No new studies describing the effects of diesel or kerosene exposure on reproductive or developmental effects were identified. A handful of studies on diesel exhaust are discussed in the previous section on exposure to exhaust. Other than the ATSDR profile on jet fuels, no systematic reviews or meta-analyses of fuels were identified.
ATSDR’s review of jet fuels reported a lack of data showing clear reproductive effects in humans. It included one study of military and civilian personnel working at a U.S. Air Force base with occupational exposure to fuels in which the exposed women were found to be not at increased risk for menstrual disorders but did have decreased LH levels before ovulation. However, the authors noted that the effects could have been related to jet fuel, other fuels, or products of complete or incomplete combustion (ATSDR, 2017; Reutman et al., 2002; U.S. Army, 2001).
Animal studies on mice, rats, and dogs have examined the reproductive effects of jet fuels using inhalation, dermal, or oral exposures that ranged from 14 days to 103 weeks. No gross, microscopic, or histological changes in reproductive organs or effects on sperm parameters were reported. One study of preconception exposure for 90 days before mating and throughout gestation, with doses of up to 1,500 mg/kg/day JP-8 by gavage to female rats, did not find any significant effect on pregnancy rate, gestation length, or litter size (ATSDR, 2017; Mattie et al., 2000).
Exposure to jet fuels has not been associated with changes in pregnancy rate, gestation length, or litter size in animal models (ATSDR, 2017).
Reproductive Effects in Men and Women
One epidemiologic study compared the levels of reproductive hormones in filling station workers with office workers, matched by age and by glutathione S-transferase T1 and glutathione S-transferase M1 genotypes (GSTT1 and GSTM1; genes that code for enzymes that detoxify some of the toxic chemical compounds found in petrochemicals) in order to investigate potential genetic sensitivity (Saadat and Monzavi, 2008; Saadat et al., 2013). The first analysis found lower testosterone and higher LH levels in the filling station workers. The authors reported a significant statistical interaction between the GSTT1 genotype and occupation for LH levels, but not for testosterone or FSH levels (Saadat and Monzavi, 2008). In the second analysis, lower testosterone levels were reported for filling station workers with the GSTM1 genotype, regardless of GSTT1 genotype, but no other significant differences were seen (Saadat et al., 2013). Both analyses compared mean hormone levels and did not report any ORs.
One ecological study of birth weight among live births in Montreal found no impact of living near gasoline filling stations based on maternal residence at birth (Huppe et al., 2013).
In a study of male rats exposed to gasoline fumes for 0, 1, 3, 5, or 9 hours per day for 12 weeks, investigators observed significant dose-dependent effects on serum reproductive hormones, oxidative parameters in testicular tissue, sperm parameters (including motility, count, and morphology), and testicular histopathological changes. The authors attributed these changes to oxidative stress (Owagboriaye et al., 2017).
Leachates from gasoline-, diesel-, and kerosene-dispensing sites (0.5 mL of 0%, 1%, 5%, 10%, 25%, or 50% leachate by intraperitoneal injection) were found to adversely affect sperm parameters (morphology and count) 30 days after exposure in mice. Significant effects were noted at 5% for gasoline and 10% for diesel and kerosene. However, the leachate contained high concentrations of heavy metals, PAHs, and benzene, which may be responsible for the observed effects (Alabi et al., 2017).
In recent animal studies, exposure to gasoline during pregnancy did not affect maternal weight gain or food consumption, litter size, pup weight, or pup loss (Bushnell et al., 2015), except at the highest doses (Roberts et al., 2014a,b).
ATSDR did not report on any studies of developmental effects in humans, and it concluded that JP-8 was not embryotoxic or teratogenic according to standard developmental endpoints in animal models. However, several studies in animals have shown other developmental effects of jet fuels after in utero exposure, including decreased fetal weight and average litter size, suppressed immune function, decreased thymus and spleen weight, and transient motor delay (ATSDR, 2017).
Two case-control studies examined the relationship between parental vehicle refueling and childhood cancer (Bailey et al., 2011; Greenop et al., 2015). Vehicle refueling from a year before pregnancy up to the time of the child’s cancer diagnosis was assessed. No association was reported between refueling and childhood ALL (Bailey et al., 2011). Childhood brain cancers were related to paternal refueling (≥4 times/month OR=1.59, 95% CI 1.11–2.29), but not to maternal exposure (Greenop et al., 2015). Neither study differentiated among preconception, prenatal, and postnatal exposures.
Adult rats exposed to gasoline in utero through GD 20 were tested for a variety of outcomes, including being given a series of cognitive and neuropsychological tests, tests of immune function, and physiologic assessments; they showed no effect for any of the outcomes (Bushnell et al., 2015; Herr et al., 2016; Oshiro et al., 2015). Even at very high exposure levels (20,000 mg/m3) adverse maternal and developmental effects were not observed in mice and rats (Roberts et al., 2014a,b).
A two-generation rat study of gasoline evaluated neuropathology and regional brain glial fibrillary acidic protein content. Parental males and females were exposed, as were F1 animals through lactation, to high doses of gasoline (up to 20,000 mg/m3). No adverse reproductive or developmental effects were noted in any of the three generations (Gray et al., 2014).
An in vitro study of mouse embryonic stem cells (ES-D3) exposed to petroleum substances (containing varying concentrations of PAHs) showed an inhibition of cell differentiation into cardiomyocytes that was dose-dependent and related to PAH concentration. By contrast, gas-to-liquid products (without PAHs) had no effect on viability or cell differentiation. Because the results correlated with those of in vivo studies, the authors concluded that the prenatal developmental toxicity of petroleum substances is due to their PAH components (Kamelia et al., 2017).
A systematic review of the evidence for inherited transgenerational effects found that although there is no human evidence concerning the transgenerational effects of JP-8, there are four studies in rats, all from the same group of investigators, that examined the transgenerational effects of a gestational exposure to JP-8. The effects the studies examined included growth and development, reproductive effects in offspring, and renal effects in offspring (Walker et al., 2018). The observed effects on body weight were inconsistent with one study that had reported an increase (Tracey et al., 2013) and another that reported no change (Manikkam et al., 2012c). Results suggest that the effects of JP-8 exposure are related to increased apoptosis of germ cells in male offspring, but not to reproductive organ weight, prostate disease, or puberty (Manikkam et al., 2012c; Tracey et al., 2013). In female offspring, JP-8 exposure was associated with decreased follicle number and increased ovarian cysts in female offspring, but its effects on puberty were inconsistent, and no effects on fertility or reproductive organ weight were observed (Manikkam et al., 2012c; Nilsson et al., 2012; Tracey et al., 2013). Transgenerational effects on kidney disease were not observed with JP-8 exposure, even though such effects were observed in earlier generations (Tracey et al., 2013). Inconsistencies across studies (heterogeneity of endpoints, differences in measurement of endpoints, and age at observation), a lack of studies from other researchers, and a lack of analysis using the litter as the unit of analysis limit the confidence in these findings.
Epigenetic effects were associated with exposure to JP-8 in rats (Tracey et al., 2013). No new studies were found that reported epigenetic or transgenerational effects of gasoline, diesel, or kerosene. Rodent and cell-line models suggest that alterations in genome methylation during critical developmental windows, in part through repetitive regions, have profound effects on both the offspring and the subsequent generations at least to the F3. These effects include immediate F1 toxicity which is later followed by an increase in disease frequency. Of note, F3 females exposed to plastics, dioxin, or jet fuel exhibit an increased rate of polycystic ovary syndrome, with 100% affected with exposure to jet fuel. The rate of F3 male obesity increased 10–15% following exposure to jet fuel (Tracey et al., 2013). However, these outcomes were not correlated with genome methylation.
Synthesis and Conclusions
Since the Volume 3 review in 2004 and the ATSDR review in 2017, little new information has been published regarding the reproductive, developmental, and generational effects of fuels. There are no new high-quality epidemiologic studies that provide evidence of additional effects. The Volume 3 committee concluded that there was “inadequate/insufficient evidence” to permit a conclusion regarding an association between exposure to “fuels and reproductive or developmental effects (including infertility, spontaneous abortion, childhood leukemia, CNS [central nervous system] tumors, neuroblastoma, and Prader-Willi syndrome)” (IOM, 2005). ATSDR’s review similarly notes a dearth of data pertaining to the reproductive and developmental effects of jet fuel in humans. Although a few animal studies have shown a variety of developmental effects associated with prenatal JP-8 exposure (decreased fetal and pup weight, suppressed immune function, delayed motor coordination), this fuel is not embryotoxic or teratogenic by standard endpoints, and there are no data on the developmental effects of JP-5 (ATSDR, 2017). Animal studies have shown that epigenetic and transgenerational effects are possible (Walker et al., 2018).
Between the three case-control studies (Bailey et al., 2011; Greenop et al., 2015; Saadat and Monzavi, 2008; Sadaat et al., 2013) and one ecological study (Huppe et al., 2013) that studied reproductive hormone levels, PTB, childhood ALL, and brain tumors, few associations have been reported. Animal studies have shown a few effects, mostly at very high doses, that are of questionable relevance to humans.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between exposure to fuels and reproductive or developmental effects.
TABLE 6-9 Summary of Reproductive and Developmental Effects of Fuels
|Saadat and Monzavi (2008)||Case-control
114 filling station workers and 112 office workers as age-matched controls in Iran, 2006.
|Exposure categorized by job title; GSTT1 genotype.||Testosterone, LH, and FSH in serum by job and GSTT1 genotype.
Testosterone: levels lower in filling station workers than for office workers for GSTT1-present (mean=4.23 versus 4.67, p=0.54) and GSTT1-null genotypes (mean=4.26 versus 5.48, p=0.001); no interaction.
LH: levels higher in filling station workers than for office workers for GSTT1-present (mean=2.82 versus 1.17, p=0.50) and the GSTT1-null genotype (mean=3.52 versus 1.07, p=0.025); significant interaction between GSTT1 genotype and exposure (p=0.003).
FSH: no differences between exposure groups or genotypes; no interaction. No ORs reported.
|Saadat et al. (2013)||Case-control
114 filling station workers and 100 office workers as age- and sex-matched controls in Iran, 2006.
|Exposure categorized by job title; GSTM1 and GSTT1 genotype.||Testosterone, LH, and FSH in serum by job and GSTM1 and GSTT1 genotypes.
Among workers with the null GSTM1 genotypes, levels of testosterone were higher in unexposed workers than filling station works (p<0.001), regardless of GSTT1 genotype.
No other differences noted among exposure, genotypes, or hormones. No ORs reported.
|Adverse Pregnancy Outcomes|
|Huppe et al. (2013)||Ecological
267,478 live births on the Island of Montreal, 1994–2006.
|Distance from gasoline filling stations based on residential zip code at birth.||PTB (<37weeks gestation) determined from administrative live birth file.
No association between residential distance to gasoline filling stations and PTB (adjusted for maternal education, country of birth, marital status, age, parity, infant sex, unemployment rate, renters’ proportion paying 30% or more of income on rent, and mean street-block income).
Adjusted for SES as estimated by census tract and postal code.
|Bailey et al. (2011)||Case-control
389 cases of ALL and 876 controls matched by age, sex, and state of residence, <15 years old.
2003–2007 in Australia
|Self-reported exposures by questionnaire and computer-assisted telephone interview about vehicle refueling in the year before and during pregnancy.
Does not distinguish among preconception, prenatal, or postnatal exposure.
No association between childhood ALL and parental vehicle refueling in the year before or during pregnancy.
|Greenop et al. (2015)||Case-control
306 cases of cases of childhood brain cancer and 950 controls matched by age, sex, and state of residence, <15 years old.
2005–2010 in Australia.
Same controls as Bailey et al. (2011).
|Self-reported exposures by questionnaire and computer-assisted telephone interview about vehicle refueling in the year before, during pregnancy, and to the time of diagnosis.
Does not distinguish among preconception, prenatal, or postnatal exposure.
|Childhood brain tumors.
No association between maternal refueling before or during pregnancy and brain tumors.
Paternal refueling ≥4 times/month OR=1.59, 95% CI 1.11–2.29; adjusted for matching variables, father’s age, education, occupational exposure to solvents and diesel exhaust.
NOTE: ALL=acute lymphoblastic leukemia; CI=confidence interval; FSH=follicle-stimulating hormone; GSTM1=glutathione S-transferase mu 1; GSTT1=glutathione S-transferase theta 1; LH=luteinizing hormone; OR=odds ratio; PTB=preterm birth; SES=socioeconomic status.
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