Pesticides were used in the 1990–1991 Gulf War and Post-9/11 conflicts to combat the endemic insect and rodent populations in Iraq, Afghanistan, and other Persian Gulf areas; one goal was to control insects that were vectors for infectious diseases such as leishmaniasis, malaria, and sand fly fever. More than 35 types of pesticides were used during the Gulf War, while the number of pesticides used in the Post-9/11 conflicts has not been determined by the Volume 11 committee. The U.S. Environmental Protection Agency (EPA) had to approve each insecticide for use in the Gulf War, and the specific types of pesticides and the quantities shipped to military personnel were documented. Military personnel in the Gulf War were exposed to pesticides through both personal and field use. There were specific Department of Defense (DoD) guidelines regarding which pesticides should have been used and for what purposes, but there were reports of both misuse and lack of use. Additionally, reports suggest that pesticide use differed by service, living arrangements, race, rank, and season (Fricker et al., 2000). U.S. Army personnel used pesticides more frequently than members of the other service branches.
DoD indicated that the approximately 700,000 service members deployed to the Persian Gulf had access mainly to two pesticides: permethrin and DEET (IOM, 2003).1 Service members received spray cans of permethrin to treat their uniforms. It is estimated that 130,894 boxes of 5-oz cans (12/box) and 112,590 boxes of 2-oz tubes (12/box) of permethrin were sent to the Persian Gulf; 44% of service members were estimated to have had some level of exposure. Almost 55,000 boxes of DEET, used as a personal insect repellant, were sent to the Gulf War in the form of a stick or liquid; about 44% of service members were exposed (IOM, 2003). In addition, numerous other pesticides were used. For example, almost 4,000 cases (12 bottles/case) of chlorpyrifos, the most commonly used organophosphate (OP) pesticide, were also shipped to the Persian Gulf region, along with more than 111,000 2-oz cans of
1 The Volume 2 committee obtained information on the specific pesticides and solvents used during the Gulf War from a variety of sources, including veterans; DoD and, specifically, the Defense Logistics Agency; the Department of Veterans Affairs; the RAND Corporation; the Presidential Advisory Commission (Cecchine et al., 2000; PAC, 1996); and P.L. 105-277 (Gulf War Veterans Act) and P.L. 105-368 (Persian Gulf War Veterans Act) (see Chapter 1). See Gulf War and Health, Volume 2 (IOM, 2003) for more information on the use of pesticides in the Gulf War.
lindane, an organochlorine pesticide. Furthermore, some service members reportedly used pesticides that were not approved by DoD and were obtained on the local market, such as pet tick-and-flea collars.
Service members deployed to the Post-9/11 conflicts also used pesticides. For example, in Operation Iraqi Freedom in 2003, pesticides were used to repel and control sand flies. U.S. service members serving at Tallil Air Base in Iraq applied DEET to exposed skin, wore permethrin-impregnated uniforms, and slept under permethrin-treated bed nets. In April 2003 a requisition was made for 1,670 boxes of DEET insect repellent (each box contained 12 2-oz tubes), 2,500 boxes of permethrin (each box contained 12 individual dynamic absorption kits for the treatment of uniforms, insect bed nets, or both), and 10,000 insect bed nets. However, service members reported that these pesticides were not very effective against the sand flies. Area spraying of malathion, resmethrin, and pyronyl oil containing 3% permethrin was also conducted in and around the air base for sand fly control. Other pesticides used to treat residential and outdoor areas at the air base included carbaryl, lambda-cyhalothrin, cyfluthrin, bifenthrin, chlorpyrifos, deltamethrin, and cypermethrin (Schleier, 2009). In all, 29,287 oz of pesticides containing 8,258 oz of active ingredient were applied during 2003 at the air base (Coleman et al., 2006). Schleier et al. (2009) reported that these sand fly control efforts were largely unsuccessful. This is just one military installation among many in Iraq and Afghanistan where pesticides were used for insect control.
Although DoD sent rodenticides to the Persian Gulf and to Iraq, the committee did not review the health effects of rodenticide exposure. Inasmuch as those products were sent to the Persian Gulf in pellet form (Cecchine et al., 2000), exposure would have probably required ingestion. The use of rodenticides was also reported by Coleman et al. (2006) at Tallil Air Base in Iraq in 2003. Attempts to eliminate the mouse population on the base were made using Maki rodent pellets (whose active ingredient is 3-[3-(4’-bromo[1,1’- biphenyl]-4-yl)-3-hydroxy-1-phenylpropyl]-4-hydroxy- 2H-1-benzopyran-2-one) (Coleman et al., 2006). Because there were no accounts of military personnel consuming rodenticides or coming in direct contact, neither the Volume 2 committee nor the Volume 11 committee reviewed the adverse health effects of rodenticides. The Volume 11 committee was unable to identify further specific information in the literature on the exposure of Post-9/11 service members to pesticides, although it had no reason to conclude that any reproductive or developmental effects from exposure to the pesticides would be different for Post-9/11 veterans than for Gulf War veterans.
This chapter focuses on the active ingredients of pesticides (the mechanisms of action for these chemicals were discussed in Gulf War and Health, Volume 2); however, it is important to remember that the toxicity of a pesticide can be altered by its formulation. Agents contributing to formulation of a pesticide are often listed as “inert ingredients” (for example, petroleum products, xylenes, oils, and surfactants), but they can alter the toxicokinetics of a pesticide, potentially increasing the absorption of active ingredients, or can be toxic themselves (Petrelli et al., 1993; Ware, 1989). Many of these “inert ingredients” are considered elsewhere in this report because they act as solvents or fuels, and they may also be found in combustion products from oil-well fires and burn pits.
There are several overarching concerns when considering the possible reproductive, developmental, and generational health effects that may occur in Gulf War and Post-9/11 veterans and their descendants. First, there were several routes by which service members may have been exposed to pesticides, such as dermal contact, which could have occurred through the use of repellant or the wearing of contaminated clothing; inhalation from breathing polluted air or being downwind of spraying activities; and oral ingestion as a result of hand-to-mouth contact. The route of exposure can have significant implications for the toxicity of the pesticide. For example, the oral LD50 of carbofuran is 5 mg/kg in the rat, while the dermal LD50 is 120 mg/kg, indicating that oral exposure can be 24 times more toxic than dermal exposure (Roberts and Reigart, 2013). Second, as is evident from the description of the leishmaniasis control program used at Tallil Air Base, service members were often exposed to more than one pesticide
at a time or within a short time span. Exposures to mixtures of pesticides and to pesticides mixed with other chemicals are a common occurrence not only for service members but also for individuals who are occupationally exposed to toxicants, and even the general public. Nevertheless, many studies assessing adverse effects resulting from possible exposures to pesticides have focused on the effects of a single pesticide or group of pesticides (e.g., OPs), rather than on mixtures of pesticides or the combined exposure with other chemicals.
Most of the information on the exposure of service members to pesticides is based on information for the 1990–1991 Gulf War and described in detail in Volume 2. That committee concluded that there was inadequate/insufficient evidence to determine whether an association exists between exposure to pesticides and “male or female infertility after cessation of exposure,” “parental preconception exposure to pesticides and spontaneous abortion or other adverse pregnancy outcomes,” and “parental preconception exposure to pesticides and congenital malformations” (IOM, 2003). These conclusions were applicable to all the Gulf War pesticides considered in Volume 2 and were the only reproductive or developmental endpoints identified.
The Volume 11 committee notes that pesticides, unlike many other commonly used chemicals, must be registered for use by EPA, and those registrations are periodically renewed. Some of the pesticides sent to the 1990–1991 Gulf War were no longer registered for use by the time that the Post-9/11 conflicts began or else have had their registrations terminated since 2001.
In the sections below, the committee considers the reproductive and developmental effects that may result from exposure to the following pesticides: OPs, carbamates, pyrethroids, lindane, and DEET (N,N-diethyl m-toluamide).
The class of OP pesticides is large and includes a number of pesticides that are in wide use, primarily in agriculture. P.L. 105-277, the Persian Gulf War Veterans Act, lists four specific OP pesticides—chlorpyrifos, diazinon, dichlorvos, and malathion—that were used in the 1990–1991 Gulf War. Since that war, three of the OP pesticides—chlorpyrifos, diazinon, and dichlorvos—have had their use as residential pesticides restricted and therefore are unlikely to have been sent to the Post-9/11 conflicts. Beginning in 2000, most registered household use of diazinon and chlorpyrifos was to be phased out and eventually cancelled; most household use of dichlorvos was cancelled in 1995 (ATSDR, 1997b). Malathion continues to be used in public health pest control programs for controlling mosquito-borne illnesses. It is available to the general public for residential use, and it was sprayed on military installations in Iraq (Schleier, 2009). As all four OPs were used in the Gulf War and malathion was used in Post-9/11 conflicts, the Volume 11 committee considered the possible reproductive and developmental effects for all four OPs.
OPs are alkylating agents that can lead to deoxyribonucleic acid (DNA) damage (Ojha and Gupta, 2015). These agents were developed for their ability to inhibit the enzyme acetylcholinesterase (AChE). AChE is responsible for the breakdown of the neurotransmitter acetylcholine. When acetylcholine does not break down in the presence of OPs, the neuromuscular synapses become overstimulated, which can lead to nervousness, delirium, hallucinations, and psychoses (Ecobichon, 2001).
Exposure to OP pesticides may be assessed by AChE activity in erythrocytes as a proxy to indicate the enzyme status in the nervous system. Exposure to OPs may also be determined by detecting OP metabolites in urine or blood (IOM, 2003). OP metabolites, usually dialkylphosphates (DAPs), may be measured as individual metabolites or as a sum of several DAPs (ΣDAPs). Among the commonly measured urinary DAPs of OPs are dimethylphosphate (DMP), dimethylthiophosphate (DMTP),
dimethyldithiophosphate (DMDTP), diethylphosphate (DEP), diethylthiophosphate (DETP), and diethyldithiophosphate (DEDTP). Most of the OPs considered in this report may be metabolized to one or more DAPs, so the detection of a specific DAP does not necessarily indicate which OP an individual may have been exposed to. However, some other metabolites are specific for a given OP; for example, the detection of 3,5,6-trichloro-2-pyridinol (TCPy) in the urine is indicative of an exposure to chlorpyrifos (Bicker et al., 2005). These pesticides have a short half-life on the order of hours or days, and therefore a single measure may not adequately reflect exposure over a critical window of development. In addition, the nonspecific metabolites may reflect exposure to the metabolites themselves, not just the parent compound (Bradman et al., 2015; Eskenazi et al. 2014; Quirós-Alcalá et al., 2012), and excretion may differ from individual to individual according to their metabolism (Wessels et al., 2003).
In 1997, ATSDR published toxicological profiles of chlorpyrifos and dichlorvos. The profile for chlorpyrifos contained no information on reproductive effects in humans by any route of exposure. Similarly, the profile for dichlorvos also had no information on reproductive effects in humans by any exposure route. The 2003 ATSDR Toxicological Profile for Malathion included one study in which women had been exposed to malathion as a result of spraying for insect control in an urban area. There was a “moderate association” of malathion exposure with stillbirths, but not with spontaneous abortions. In a study that compared birth outcomes in 1,016 couples in which the men applied malathion as well as other pesticides with 1,020 couples in which the men were not exposed to pesticides, the exposed couples had more stillbirths (8.73% versus 2.65%) and more abortions (26.0% versus 15.0%) than the unexposed couples. The study also found a significantly lower percentage of fertile males (80.8% in the exposed males versus 94.9% in controls) and a decreased percentage of live births (53.0% versus 80.1%) (ATSDR, 2003a; Rupa et al., 1991). Finally, the 2008 ATSDR Toxicological Profile for Diazinon contained no studies of the effects of diazinon exposure on human reproduction.
Gulf War and Health, Volume 2 (IOM, 2003) reported on one study of time to pregnancy (TTP) in Ontario farm couples exposed to OPs (Curtis et al., 1999). Neither men nor women had a significantly altered time to pregnancy with OP exposure. Volume 2 reported no other reproductive effects in human or animals that may have occurred from exposure to the four OPs listed in the Persian Gulf War Veterans Act.
Reproductive Effects in Men and Women
The Volume 11 committee identified several new studies on reproductive effects and exposure to the OPs of interest. The new studies assessing OP exposure and reproductive effects in men and women are listed in Table 5-2. The effects of OPs on male fertility have been studied by several researchers, including a recent systematic review by Martenies and Perry (2013). In general, human studies have focused on sperm quality associated with either environmental or occupational exposures.
Meeker et al. (2004a) assessed TCPy, a urinary marker of chlorpyrifos exposure, in 272 men who were not occupationally exposed to pesticides and who were attending an infertility clinic in Massachusetts. They reported that increasing urinary TCPy levels were associated, although not significantly, with increased risk of having sperm concentrations below 20 million/mL, less than 50% motile sperm, and less than 4% of sperm with normal morphology. Meeker et al. (2004b) further studied sperm DNA integrity, which they assessed by neutral comet assay and reported as comet extent, percentage DNA in comet tail (Tail%), and tail distributed moment (TDM). Statistically significant increases in Tail% for an
interquartile range (IQR) increase in TCPy (β=2.8, 95% confidence interval [CI] 0.9–4.6) were reported, while a decrease in TDM was associated with IQR changes in TCPy (β= –2.5, 95% CI –4.7– –0.2).
Using the same population as Meeker et al. (2004b), Figueroa et al. (2015) examined whether exposure to OPs was associated with sperm chromosomal abnormalities—specifically XX18, YY18, XY18, and total disomy in sperm nuclei. Increasing levels of urinary metabolites, both individual and total sum of DAPs, were significantly associated with increased levels of disomy, particularly XX18 and XY18, although the increases were not monotonic. The highest significant association was observed between the third exposure quartile of DMTP (2.21–6.47 ng/mL) and XX18, with a 52% increase in the incidence rate ratio (95% CI 1.36–1.69).
Sperm integrity, specifically the sperm chromatin structure, was also evaluated in 33 men occupationally exposed to OPs in Mexico (Sánchez-Peña et al., 2004). About 75% of the men had semen with altered sperm chromatin structure (>30% of DNA fragmentation index [DFI]) that would classify them as having poor fertility potential, compared with only 4% of the unexposed men. Only the DAP concentrations for DETP were significantly associated with mean DFI (p=0.026) and standard deviation-DFI (p=0.022) and marginally associated with DFI% (p=0.079) after adjustment for potential confounders, such as the presence of immature germ cells (IGCs). No other DAPs had any associations with measured outcomes. No significant associations were found between semen quality parameters, including IGCs, and DETP or any other DAP, although IGC values were above the reference value for 82% of the samples versus only 9% of unexposed men. There was a negative correlation between mean DFI and sperm viability (r=0.254; p<0.027) and also seminal volume (r=0.333; p<0.004).
A further assessment of the effect of exposure to environmental OPs on sperm parameters and male reproductive hormones was conducted by Melgarejo et al. (2015). This was a study of 116 men attending an infertility clinic in southern Spain, all of whom had detectable concentrations of at least one urinary OP metabolite (the possible sources of exposure were not discussed). The serum levels of reproductive hormones were within the normal range, and only DEDTP showed a positive association with lutenizing hormone (LH) levels (β=11.4, 95% CI 0.81–22.1) and follicle-stimulating hormone (FSH) levels (β=3.2, 95% CI 0.08–6.2) and a negative association with the testosterone/LH ratio (β= –9.6, 95% CI –18.7− –0.50). However, the semen quality parameters were inversely correlated with the urinary OP metabolites. The sperm concentration and total sperm count were both significantly inversely associated with urinary concentrations of the DMP, DMDP, and DMDTP as well as ΣDAP. There was a significant inverse association between the percentage of motile sperm and DMTP, DMDTP, and DEP, and the total motile count was significantly inversely associated with urinary DMP and DMDTP concentrations (all significant at the ≤0.05 level). The seminal volume and percentage of sperm with normal morphology were not affected by DAP levels.
Although their study was not specific for any particular OP or carbamate pesticide (the latter are discussed in the next section), Miranda-Contreras et al. (2013) measured erythrocyte AChE and plasma butyrylcholinesterase (BuChE) levels in 64 male agricultural workers in Venezuela with exposure to these and other pesticides and also in 35 unexposed men. A total of 87.5% of the exposed workers showed exposure to OP and carbamate pesticides, and 82.8% had mild inhibition of BuChE activity. The levels of reproductive hormones and sperm quality were assessed in the workers. In the exposed group, there were significant adverse effects observed in semen pH, sperm motility, abnormal sperm, the number of live sperm, and the DNA fragmentation index versus those values in the unexposed group. Reproductive and thyroid hormones (testosterone, prolactin, free T4, and thyroid stimulating hormone) were normal in most of the exposed workers, although 44% had levels of LH and 21% had levels of FSH above reference values.
In a systematic review that assessed exposure to OPs and semen parameters, Martenies and Perry (2013) found that all six studies published between 2007 and 2012 reported significant adverse associations between OP exposure and at least one semen parameter (Hossain et al., 2010; Perez-Herrera et al., 2008; Perry et al., 2007, 2011; Recio-Vega et al., 2008; Yucra et al., 2008). The Volume 11 committee notes that exposures in most of the studies cited in this review were generally to more than one class of pesticide, e.g., OPs and carbamates, and to more than one pesticide within a class.
In a small biomonitoring study in China, Perry et al. (2007) examined sperm concentration in 18 men who did not work directly with pesticides but who had visited an agricultural area. Urine samples were analyzed for 13 parent OPs or their metabolites. Only the crude difference in sperm concentration for low versus high concentrations of DETP in urine was statistically significant (absolute sperm concentration difference= –1.0, 95% CI –1.8− –0.2). In a larger study of 94 exposed men and 95 controls in eastern China, Perry et al. (2011) confirmed that men with lower sperm concentration and motility were more likely to have higher urinary levels of DMP (odds ratio [OR]=1.30, 95% CI 1.02–1.65) than men with higher sperm concentrations; there was no association with five other urinary OP metabolites.
Recio-Vega et al. (2008) found similar effects on sperm quality in a study of 52 male agricultural workers in Mexico who were exposed to a number of pesticides, including diazinon. OP metabolites were found in the urine of 83% of the occupationally exposed participants. Of the seven semen parameters evaluated, only semen volume and total sperm count were significantly lower (p=0.004 and p=0.01, respectively) in the most exposed subjects compared with unexposed controls, and in the highly exposed group there was a negative correlation between semen volume and sperm count; there were no other significant associations. There were seasonal variations in the application of pesticides, with a corresponding variation in the levels of urinary metabolites and semen quality. Yucra et al. (2008) also found that exposure to OPs as reflected in the levels of urinary metabolites, particularly DETP and DEDTP, was associated with semen quality among 31 pesticide workers in Peru who were compared with 31 unexposed workers. Men with the highest levels of these two metabolites had lower volumes of semen and higher seminal pH than men with lower levels of DEDTP and DETP; there were no other significant associations. The reported use of the pesticides paraquat and/or malathion was correlated with semen quality in a group of 62 farmers in Malaysia who were compared with nonexposed controls (Hossain et al., 2010). The risk of having impaired semen parameters (lower semen volume, lower sperm concentration, a smaller percentage of motile sperm, and more abnormal sperm) was significantly increased (ORs 4.5–8.7) in exposed workers; seminal pH was not significantly changed. The committee notes that in the Hossain et al. study, exposure to the pesticides was not confirmed nor were the various pesticides distinguished by metabolite analysis.
Finally, Pérez-Herrera et al. (2008) evaluated the role of the PON1Q192R2 polymorphism on the susceptibility to OP toxicity in 54 male agricultural workers in Mexico. All participants had abnormal sperm morphology, 46% had low ejaculate volume, 30% had low sperm motility, and 13% had low sperm viability. OP exposure, either at the month of sample collection or during the preceding 3 months before sampling, was not associated with any parameters of semen quality or with DNA integrity. However, for men with the 192RR but not QQ genotype, there was a significant negative association between OP exposure during the 3 months before sampling (representing one spermatogenic cycle) and decreased sperm viability (p=0.086).
2 OP deactivation occurs primarily by serum paraoxonase (PON1), which is encoded by a single gene on chromosome 7q21–22. PON1 activity in serum shows a wide variation among individuals, which may be attributed to the presence of polymorphisms in the PON1 gene at positions 55 and 192; the latter has two isoforms (Q and R). The 192R alloenzyme hydrolyzes paraoxon (the parathion oxon metabolite) faster than the 192Q isoform, while diazoxon (the oxon of diazinon) is hydrolyzed faster by the 192Q isoform (Pérez-Herrera et al., 2008).
The Volume 11 committee identified two studies on female reproductive effects from exposure to OPs (Al-Hussaini et al., 2018; Hu et al., 2018). In the first study, Al-Hussaini and colleagues reported on 94 couples at an infertility clinic in Egypt. Follicular fluid from each woman was collected during in vitro fertilization/embryo transfer and analyzed for the presence of chlorpyrifos (mean=146.9±43 [s.d.] μg/L), diazinon (mean=284.7±108.2 μg/L), and malathion (mean=38.5±9.5 μg/L), among other pesticides and polychlorinated biphenyls. There were negative correlations between the follicular fluid concentrations of all three pesticides and endometrial thickness (p<0.05 for diazinon and malathion and p<0.01 for chlorpyrifos). The concentrations of diazinon and chlorpyrifos, but not of malathion, were negatively associated with the number of oocytes retrieved (p<0.01 for both pesticides) and with the implantation rate (that is, the number of sacs; p<0.01 for both pesticides). None of the OPs were associated with reduced fertilization or early embryo cleavage rates.
Hu et al. (2018) conducted a TTP study with 569 women recruited from two preconception care clinics in Shanghai, China, from August 2013 and April 2015. The subjects were enrolled in the study before conceiving a child and were followed for approximately 1 year. Exposure to organophosphates (and pyrethroids) was determined through the measurement of urinary metabolites. After adjustments were made for age, body mass index (BMI) before pregnancy, smoking, and education, women in the highest quartile for urinary levels of DETP were found to have significantly longer TTP (fecundability OR=0.68, 95% CI 0.51–0.92) and increased infertility (OR=2.17, 95% CI 1.19–3.93) than women in the lowest quartile.
Adverse Pregnancy Outcomes
Numerous researchers have examined the effects of exposure to OPs on a variety of pregnancy outcomes, such as infant body weight and length and head circumference, preterm birth, and duration of gestation. Some studies have included large cohorts of mother–child pairs and have followed the children for several years after birth. The postnatal developmental effects observed in these children are discussed in the section on developmental effects. In this section the committee focuses on fetal growth and birth outcomes, including spontaneous abortion, as important reproductive effects.
The committee notes that none of the studies considered in this section provide information on maternal or paternal exposures during the preconception period or the first trimester of pregnancy. Rather, the studies usually assess maternal and fetal exposure by measuring OP metabolites in maternal or cord blood or maternal urine samples during pregnancy. In addition, for the Columbia Center for Children’s Environmental Health (CCCEH) study the mothers wore personal air monitors during pregnancy to measure ambient exposures, and the researchers measured the parent compounds chlorpyrifos and diazinon in maternal blood within one day of delivery.
There are four major longitudinal U.S. birth cohort studies of pesticide exposure and child health that have examined the effects of prenatal exposure to OPs and other pesticides on fetal growth: the CCCEH study in New York City; the Center for the Health Assessment of Mothers and Children of Salinas (CHAMACOS) study in California; the Health Outcomes and Measures of the Environment (HOME) study in Cincinnati, Ohio; and the Children’s Environmental Cohort Study at Mount Sinai Hospital in New York City. Each of these studies has combined an assessment of maternal pesticide exposure with some measure of birth outcomes. Harley et al. (2016) evaluated birth outcomes across all four U.S. cohorts.
The CCCEH study followed a cohort of 263 inner-city African American or Dominican women enrolled between 1998 and 2002 (Perera et al., 2003). Women wore personal air monitors for 48 hours in the second or third trimester to determine their exposure to airborne toxicants. Blood and urine were
also sampled. Chlorpyrifos was detected in 98% of maternal blood samples and 94% of cord blood samples at birth. Prenatal chlorpyrifos exposure (pg/g blood) was associated with a significant decrease in birth weight (grams) overall (β= –0.04, p=0.01) and among African Americans (β= –0.05, p=0.04), but not among Dominican infants (β= –0.03, p=0.11). It was also associated with reduced birth length overall (β = –0.01, p=0.003) and in Dominicans (β= –0.02, p<0.001); chlorpyrifos exposure was not significantly associated with reduced head circumference in any group.
The CHAMACOS study investigated a group of 488 low-income Latina women and their infants in an agricultural area of Salinas Valley, California, in 1999−2000 (Eskenazi et al., 2004). DAPs were measured in maternal urine at about 13 weeks of gestation and again at 26 weeks of gestation, and cholinesterase (ChE) and BuChE were measured in maternal and umbilical cord blood at delivery. Increased DAP concentrations were associated with increased infant body length (p=0.06) and were significantly associated with an increased head circumference (p=0.03). A 10-fold increase in DMP, but not DEP, metabolites was associated with a 3-day decrease in gestational duration (β= –0.41 weeks/log10 unit increase; 95% CI –0.75– –0.02; p=0.02), but only after 22 weeks of gestation. DAP metabolites were not associated with birth weight, infant ponderal index (a weight–height index used to measure fetal growth), the risk of preterm delivery, low birth weight, or small-for-gestational-age births. Lower levels of ChE in umbilical cord blood were associated with a significantly shorter length of gestation (β=0.34 weeks/unit increase; 95% CI, 0.13–0.55; p=0.001), increased risk of preterm delivery (OR=2.3, 95% CI 1.1–4.8; p=0.02), and lower birth weight (OR=4.3, 95% CI 1.1–17.5; p=0.04). However, 6 of the 11 low-birth-weight infants were also preterm. BuChE levels in maternal and umbilical cord blood were not associated with any birth outcome.
A further analysis of the CHAMACOS cohort by Harley et al. (2011) examined whether PON1 would modify the effect of OPs on adverse fetal outcomes. Maternal and cord blood samples were analyzed for the PON1 genotype and enzyme activity. Associations of PON1 with decreased gestational age were found only in those infants who had susceptible PON1 genotypes (PON1-108TT); the presence of susceptible PON1 genotypes in mothers was not associated with decreased gestational age. No associations were found between any marker of PON1 genotype or activity as measured in maternal blood and the gestational age or infant birth weight, length, or head circumference. There was some evidence of effect modification with DAPs: maternal DAP concentrations were associated with shorter gestational age only among infants with the susceptible PON1-108TT genotype (p-value interaction=0.09). However, maternal DAP concentrations were associated with larger birth weight (p-value interaction=0.06) and head circumference (p-value interaction 0.01) in infants with a nonsusceptible genotype (PON1192QQ).
In the Cincinnati HOME study, 344 women less than 19 weeks pregnant were assessed for residential exposure to toxicants between 2003 and 2006 (Rauch et al., 2012). Urine samples collected at 16 and 26 weeks of gestation were analyzed for six DAPs; all mothers had detectable levels of at least one DAP. A 10-fold increase in total DAP concentration was significantly associated with a decrease in gestational age (β= –0.5 weeks, 95% CI –0.8– –0.1) and with a lower birth weight (β= –151 g, 95% CI –287– –16), and the association with decreases in gestation age was stronger for white mothers than for black mothers, but the reverse was true for birth weight. Decrements in birth weight and gestational age associated with ΣDAP concentrations were greatest among infants with the PON1192QR and PON–108CT genotypes.
Berkowitz et al. (2004) reported on the Children’s Environmental Cohort Study at Mount Sinai Hospital in New York City, a cohort of 404 multiethnic mother–infant pairs with babies born between 1998 and 2002. TCPy (a metabolite of chlorpyrifos) and 3-phenoxybenzoic acid (PBA; a pyrethroid metabolite) were each detected in more than 40% of prenatal maternal urines. There were no associations between maternal pesticide exposures or urinary levels of TCPy, PBA, or pentachlorophenol and adjusted birth weight, adjusted length, adjusted head circumference, or adjusted gestational age. A
positive trend was found between maternal blood PON1 activity and head circumference, but the trend was only significant for mothers whose TCPy levels were above the level of detection. No trends were seen for birth weight or birth length for TCPy or for the other metabolite levels when maternal PON1 genotype level was taken into account. Infant PON1 activity had no association with any of the fetal growth measures.
Harley et al. (2016) pooled data from the four U.S. cohort studies and found that there were no significant associations between DAP metabolites and birth weight, length, or head circumference. However, when the mothers were stratified by ethnicity, there was an association between increasing urinary DAP concentrations and decreased birth weight for non-Hispanic black women. Infants with the PON1192RR genotype were at greater risk for decreased birth length with increasing prenatal exposure to DAP metabolites. Gestational duration was not examined in the pooled analysis.
The committee identified one study in Shanghai, China, that also sought to determine whether prenatal exposure to pesticides resulted in any adverse fetal growth outcomes (Wang et al., 2012). Women (n=187) were recruited into the study between 2006 and 2007, and they provided urine samples at the onset of labor for DAP analysis. There were no significant associations between prenatal OP exposure and infant body weight or length or length of gestation, with the exception of DEP, which showed a significant inverse relationship with duration of gestation for infant girls but not boys (β= –1.79; p=0.001 versus β=0.17; p=0.164, respectively).
In an in vitro study with human villous cytotrophoblast cells isolated from human placentas, exposure to chlorpyrifos was found to cause marked changes in placental tissues including chorionic villi, suggesting the placenta may be a target of chlorpyrifos toxicity (Ridano et al., 2017).
Although the Volume 2 committee (IOM, 2003) did not report on human or animal studies of reproductive effects and OP exposure, information from animal studies was available at the time. In 1997, ATSDR had published toxicological profiles of dichlorvos and of chlorpyrifos that included the results of animal studies on the reproductive effects of those pesticides. Dermal exposure to chlorpyrifos had resulted in transient adverse effects on semen volume and quality in bulls. Oral exposure to chlorpyrifos had no effect on female reproduction in mice, rats or dogs; however, the doses of chlorpyrifos needed to cause death in pregnant mice were approximately six times lower than those needed to cause death in nonpregnant mice. Chlorpyrifos administered via inhalation had no effect on rat testes (ATSDR, 1997a). ATSDR concluded that in several animal species oral and inhalation exposures to dichlorvos did not have reproductive effects in either males or females (ATSDR, 1997b). The ATSDR Toxicological Profile for Malathion (2003a) indicated that malathion was not a reproductive toxicant in animals; however, several animal studies had found that the oral administration of malathion did produce transient testicular effects. Finally, ATSDR concluded that animal studies had provided no evidence that diazinon was a reproductive toxicant. The studies on which that conclusion was based had involved oral exposures in mice, rats, and dogs (ATSDR, 2008).
In the past 10 years, rodent research has focused on characterizing the testicular toxicity of OPs. As discussed below, these studies have treated rodents with OPs, usually over at least one sperm cycle, and found effects on sperm quality, serum hormones, testes and epididymis histopathology, biochemistry, and gene expression.
DNA damage in sperm, which has been associated with OP exposure in humans (Meeker et al., 2004b), has also been observed in animal studies after in vivo treatment with diazinon (Pina Guzman et al., 2005; Sarabia et al., 2009) and malathion (Giri et al., 2002) as well as in vitro with human sperm
after exposure to chlorpyrifos (Salazar-Arredondo et al., 2008). These findings in sperm agree with the in vitro genotoxicity that has been observed relating to chromosomal disruption in lymphocytes (Ohja and Gupta, 2015).
As in human studies such as Meeker et al. (2008a), rodent experiments in male animals have found decreased levels of serum testosterone as well as of the gonadotropins FSH and LH after short-term exposure to chlorpyrifos (Alaa-Eldin et al., 2017; Mandal and Das, 2012; Sai et al., 2014), dichlorvos (Dirican and Kalender, 2012), diazinon (Leong et al., 2013), and malathion (Bustos-Obregón and González-Hormazabal, 2003; Geng et al., 2015; Uzun et al., 2009). These hormone effects were seen in connection with sperm abnormalities and testicular histopathology. Hormone effects were observed consistently, as was testicular toxicity, when measured, with the exception of the two dichlorvos studies (Okamura et al., 2005, 2009).
Two mechanisms for OP reproductive toxicity have been studied. The suggestion that OPs generate reactive oxygen species (ROS) is supported by the presence of tissue oxidative damage markers in male reproductive organs after treatment with chlorpyrifos (Mandal and Das, 2011, 2012; Mosbah et al., 2016; Sai et al., 2014; Umosen et al., 2012). ROS-mediated genotoxicity has been established for in vitro chlorpyrifos exposures (Chauhan et al., 2016). Other studies have supported this hypothesis by showing that antioxidants provide protection against testicular toxicity (Alaa-Eldin et al., 2017; Dirican and Kalender, 2012; Dutta and Sahu, 2013; Elmazoudy et al., 2011; Elsharkawy et al., 2014; Mosbah et al., 2016; Umosen et al., 2012; Uzun et al., 2009). A second line of mechanistic work demonstrated a failure in the final steps of sperm maturation in the epidydimis after treatment with malathion and dichlorvos (Akbarsha et al., 2000). ROS are known to be involved in sperm capacitation (Aitken, 2017).
OP animal studies that included a mating trial after the males had been treated found that preconception exposure to chlorpyrifos reduced fertility (85% negative fertility using mating exposure test; Joshi et al., 2007) and decreased the number of live fetuses, increased the number of dead fetuses, and increased the number of early resorptions (Farag et al., 2010). The treatment of male rats with malathion prior to mating caused an 80% decrease in fertility (Choudhary et al., 2008).
As with humans, there is little information on female animal reproductive toxicity resulting from exposures prior to pregnancy. Chebab et al. (2017) found that 6.75 mg/kg chlorpyrifos administered to female rats for 30 days significantly decreased the serum concentrations of sex hormones and thyroid hormones.
A study with dichlorvos did not find any effects on pregnancy when female mice were treated before mating (Dean and Blair, 1976). Gomes and Lloyd (2009) and Gomes et al. (2008) found effects on both pregnancy outcome and the incidence of malformation when a mixture of OPs was given to female mice before mating. The mixture included chlorpyrifos and dichlorvos (both OPs used in the Gulf War) as well as other pesticides at LD50 levels. Such high doses coupled with the use of commercial formulations and the large number of agents make these findings difficult to interpret. Malathion has also been shown to affect the maturation of porcine cumulus–oocyte complexes in vitro by inducing oxidative stress (Flores et al., 2017).
The Volume 11 committee found a lack of relevant information in the animal literature on adverse birth outcomes associated with preconception or prenatal exposure to any of the relevant OPs considered in this chapter.
Numerous studies have examined developmental effects in children following prenatal exposures to OPs. The research for the most part has focused on potential neurodevelopmental effects that may occur
in infants and children exposed in utero. Several longitudinal studies have examined neurodevelopmental deficits in children as they grow. Respiratory effects, childhood cancers, birth defects, and ear infections in children have also been studied in the context of prenatal OP exposure. The Volume 11 committee identified only a few studies that related preconception exposure to OPs to health outcomes in children (e.g., Monge et al., 2007).
In the sections below and in Table 5-3, the Volume 11 committee reviews the literature on postnatal effects in children following prenatal exposures to OPs, with subsections devoted to each developmental effect. As noted in the introduction to this chapter, the Volume 2 committee (IOM, 2003) concluded that there was inadequate/insufficient evidence to determine whether an association exists between “parental preconception exposure to insecticides and childhood leukemia, brain and other central nervous system cancers, and non-Hodgkin’s lymphoma” or between “parental preconception exposure to insecticides and congenital malformations” in children, but that committee made no conclusions for OPs specifically.
The Gulf War and Health, Volume 2 committee did not discuss any studies on developmental effects that may result in children whose parents were exposed to OPs. On the other hand, that committee did consider a number of animal studies, which are discussed later in this section.
In November 2016, EPA released Chlorpyrifos: Revised Human Health Risk Assessment for Registration Review as part of the Federal Insecticide, Fungicide, and Rodenticide Act Section 3(g) Registration Review program. The review gave special consideration to chlorpyrifos’ neurodevelopmental effects on children. The risk assessment document (EPA, 2016) stated:
In the 2015 updated literature review (USEPA, D331251, 09/15/2015), the agency conducted a systematic review expanding the 2012/2014 review which was focused only on US cohort studies with particular emphasis on chlorpyrifos. The expanded 2015 review includes consideration of the epidemiological data on any OP pesticide, study designs beyond prospective cohort studies, and non-U.S. based studies. The updated literature review identified seven studies which were relevant (Bouchard et al. 2010; Fortenberry et al. 2014; Furlong et al. 2014; Guodong et al. 2012; Oulhote and Bouchard, 2013; Zhang et al. 2014; Shelton et al. 2014). These seven studies have been evaluated in context with studies from the 2012/2014 review (D. Drew et al. D424485, 12/29/2014). In addition, the agency has also reviewed more recent studies from CCCEH (Rauh et al. 2015) and a pooled analysis of U.S. cohort studies (Engel et al. 2015) (E. Holman, D432184, 03/25/2016). As discussed below, Rauh et al. (2015) provides further evidence of neurodevelopmental outcomes in the CCCEH study. The Engel et al. (2015) study shows relatively consistent results compared to previous studies conducted at 24 months (Engel et al. 2011; Rauh et al. 2006). . . . The agency continues to conclude that the 3 U.S. cohort studies (CCCEH, CHAMACOS, and Mt. Sinai) provide the most robust available epidemiological evidence.
The agency acknowledges the lack of established MOA/AOP pathway [mode of action/adverse outcome pathway], the inability to make strong causal linkages, and the unknown window(s) of susceptibility. These uncertainties do not undermine or reduce the confidence in the findings of the epidemiology studies. The epidemiology studies reviewed in the 2012/2014 and 2015 literature reviews represent different investigators, locations, points in time, exposure assessment procedures, and outcome measurements. Despite differences in study design, with the exception of two negative studies in the 2015 literature review (Guodong et al. 2012; Oulhote and Bouchard, 2013) and the results from the more recent Engel et al. (2015) study, all other study authors have identified associations with neurodevelopmental outcomes associated with OP exposure; these conclusions were across four cohorts and twelve study citations. Specifically, there is evidence of delays in mental development in infants (24-36 months), attention problems and autism spectrum disorder in early childhood, and intelligence decrements in school age children who
were exposed to OPs during gestation. Investigators reported strong measures of statistical association across several of these evaluations (odds ratios 2-4 fold increased in some instances), and observed evidence of exposures-response trends in some instances, e.g., intelligence measures. . . .
In summary, the EPA’s assessment is that the CCCEH study, with supporting results from the other 2 U.S. cohort studies and the seven additional epidemiological studies reviewed in 2015, provides sufficient evidence that there are neurodevelopmental effects occurring at chlorpyrifos exposure levels below that required for AChE inhibition.
On December 15, 2017, the California Environmental Protection Agency, Office of Environmental Health Hazard Assessment, Developmental and Reproductive Toxicant Identification Committee determined that chlorpyrifos was “clearly shown through scientifically valid testing according to generally accepted principles to cause reproductive toxicity,” based on the developmental endpoint (OEHHA, 2017).
The Volume 11 committee considered a number of studies on the developmental effects, particularly neurological outcomes, in children whose mothers or fathers had been exposed to organophosphate pesticides, primarily chlorpyrifos. In addition to studies conducted in Spain, Mexico, and China, researchers have also used four U.S. cohort studies to look at the effects of prenatal OP exposure on a variety of neurodevelopmental outcomes in children. Only studies not cited in Chlorpyrifos: Revised Human Health Risk Assessment for Registration Review (EPA, 2016) and considered by the committee to be relevant to its assessment are included in Table 5-3.
The CCCEH study assessed children born between February 1998 and May 2002 for psychomotor and cognitive development at 12, 24, and 36 months (Rauh et al., 2006), and at 7 years of age (Rauh et al., 2012). Younger children were administered the Bayley Scales of Infant Development II (BSID-II), and their mothers completed the Child Behavior Checklist (CBCL); the 7-year-olds were administered the Wechsler Intelligence Scale for Children, 4th edition (WISC-IV). Prenatal exposure of chlorpyrifos levels was measured in maternal blood or cord blood samples obtained at delivery, and the results were categorized as high exposure (>6.17 pg chlorpyrifos/g plasma) or low exposure (≤6.17 pg/g). At 12 and 24 months of age, there were no differences in mental or motor function between high- and low-exposure children. However, at 36 months of age, children with high prenatal exposure to chlorpyrifos scored significantly lower on the psychomotor development index (PDI; p=0.01) than children with low exposure; were in the clinical problems range for all five CBCL scales, particularly the attention problems and attention deficit–hyperactivity disorder (ADHD) scales (p=0.03 each); and were more likely than the low-exposure children to have mental delays (OR=2.4, 95% CI 1.12–5.08) and motor delays (OR=4.5, 95% CI 1.61–12.70) (Rauh et al., 2006). At 7 years of age, there were significant inverse associations between chlorpyrifos exposure and two of the WISC-IV scales: working memory (r= –0.21, p <0.0001) and full-scale intelligence quotient (r= –0.13, p=0.02) (Rauh et al., 2011). Horton et al. (2012) found a significant interaction between exposure and the child’s sex, suggesting that boys had a greater decrement in working memory than girls (β= –1.714, 95% CI –3.753–0.326). At 11 years of age, there was a significant association of prenatal chlorpyrifos exposure with tremor (Rauh et al., 2015). A subgroup of 40 CCCEH 7-year-old children (20 from the upper tertile and 20 from the lowest tertile) also underwent magnetic resonance imaging of the brain. Children in the high-exposure group had a significant enlargement of several brain areas (e.g., superior temporal, posterior middle temporal, and inferior postcentral gyri bilaterally; gyrus rectus) as well as frontal and parietal cortical thinning and an inverse dose–response relationship between chlorpyrifos and cortical thickness (Rauh et al., 2012).
A second New York City study, the Mount Sinai Children’s Environmental Health Cohort, recruited 404 mother–infant pairs during pregnancy in 1998–2001 (Engel et al., 2007, 2011). Prenatal exposure to OPs (including malathion) was determined by an analysis of maternal blood and urine samples taken at about pregnancy week 31. The Brazelton Neonatal Behavioral Assessment Scale (BNBAS) was
administered to 311 neonates before they left the hospital (Engel et al., 2007). BNBAS scores showed that neonates born to women with elevated urinary malathion dicarboxylic acid (MDA) levels had a greater number of abnormal reflexes (OR=2.24, 95% CI 1.55–3.24) than children born with elevated total DEP metabolites (OR=1.49, 95% CI 1.12–1.98). Increasing levels of total DAP and DMP metabolites were also associated with an increase in abnormal reflexes. In a further study of this cohort (n=169 children or more, depending on which assessment), neurodevelopment was assessed within 5 days of birth, at 12 and 24 months, and at 6–9 years of age (Engel et al., 2011). At 12 and 24 months, in black and Hispanic infants decreased mental development (as measured with the BSID-II) was associated with higher exposure to prenatal total DAP metabolites, whereas among whites, increasing exposure was associated with improved mental development index (MDI) scores. Furthermore, prenatal maternal DAP levels were inversely associated with children’s 24-month MDI but not their PDI. At 7–9 years of age, there were only modest effects of any prenatal OP metabolites on any psychometric measure in the children (Engel et al., 2011). A further evaluation of the 7- to 9-year-old children was done using the Social Responsiveness Scale (SRS) to assess for characteristics of autism spectrum and related disorders. No associations overall were found between any of the total DAP, DEP, or DMP metabolites and total SRS scores. However, increasing total DEP levels were associated with poorer SRS scores among blacks (β=5.1 points, 95% CI 0.8–9.4) and boys (β=3.5 points, 95% CI 0.2–6.8), although not among whites and Hispanics (β=0.2 points, 95% CI –2.8–3.2), or among girls (β= –0.4 points, 95% CI –4.1–3.3) (Furlong et al., 2014). Further analysis of the children found that DMP metabolites were negatively associated with internalizing factor scores (β= –0.13, 95% CI –0.26–0.00), but positively associated with executive functioning factor scores (β=0.18, 95% CI 0.04–0.31). However, DEP metabolites were negatively associated with the working memory index (β= –0.17, 95% CI –0.33– –0.03) (Furlong et al., 2017a).
The longitudinal CHAMACOS cohort study has tracked children born between October 1999 and October 2000 to mothers residing in an agricultural area of the Salinas Valley in California. The mothers were generally young, of Mexican origin, and from farm-working families, and most had not completed high school. Urine samples were collected from the mothers (n=381) during the first and the second half of pregnancy, and each sample was analyzed for DAP metabolites of OP pesticides. Children (n=327) were assessed for developmental parameters at 6, 12, and 24 months, 3.5 years, 5 years, and 7 years of age (Eskenazi et al., 2014). Newborn infants showed a significant increase in the number of abnormal reflexes on the BNBAS which was correlated with increasing levels of maternal total DAP, DMP, and DEP levels measured in urine (β=0.23, 95% CI 0.05–0.41; β=0.18, 95% CI 0.02–0.34; β=0.22, 95% CI 0.04–0.40, respectively); however, none of the maternal metabolite levels were correlated with any of the other BNBAS scores (Young et al., 2005). Eskenazi et al. (2007) found that maternal prenatal DAP levels were negatively associated with the child’s MDI on the BSID-II and that this association reached statistical significance at 24 months of age (per 10-fold increase in prenatal DAPs: β= –3.5 points; 95% CI –6.6– –0.5). Prenatal DAPs were also associated with pervasive developmental disorder (per 10-fold increase in prenatal DAPs: OR=2.3, p=0.05). Prenatal DAPs were not significantly associated with attention problems or ADHD when the children were 3.5 years of age, but they were associated with both outcomes by the time the children were 5 years of age (CBCL attention problems: β=0.7 points; 95% CI 0.2–1.2; ADHD: β=1.3, 95% CI 0.4–2.1), and some associations were found only among boys (Marks et al., 2010). Finally, at 7 years of age, a 10-fold increase in maternal DAP concentrations was associated with lower cognitive scores on all four cognitive subtests of the WISC-IV. The most significant associations were for verbal comprehension (β= –5.3, 95% CI –8.6– –2.0) and a decrease of 5.6 full-scale IQ points (95% CI –9.0– –2.2). Higher DAP levels later in pregnancy were associated with greater deficits in the children than were levels in the first half of pregnancy (Bouchard et al., 2011). The association between DAPs and full-scale IQ was strongest for the children of mothers
with the lowest-tertile levels of arylesterase, a measure of the enzyme activity of PON1. This relationship held for both diethyl and dimethyl DAPs and for all subscales of the WISC (Eskenazi et al., 2014). Further assessment of neurobehavior in 238 of the 7-year-old children in the CHAMACOS cohort found consistent but not significant negative associations between DNA methylation at several CpG sites close to the PON1 transcription start site in the children’s cord blood and full-scale IQ and other composite measures of cognition as measured with the WISC-IV. Nonsignificant associations were also found between cord blood methylation at cg15887283 and the children’s working memory scores (Huen et al., 2018). Studies using Pesticide Use Reporting data from California reported an association between the use of OPs in fields near the maternal residence during pregnancy and WISC Full Scale and Verbal Comprehension IQ scores in children at age 7 (Gunier et al., 2017).
A further study by Sagiv et al. (2018) of the CHAMACOS cohort examined whether there is an association between prenatal exposure to OP pesticides and autism spectrum disorder (ASD) traits in children and adolescents. Women were interviewed twice during pregnancy, once after delivery and then when the children reached the ages of 6 months, 1, 2, 3.5, 5, 7, 9, 10.5, 12, and 14 years. Exposure was assessed by measuring DAPs in urine and by collecting information on the mother’s residential proximity to OP application while pregnant, using California’s Pesticide Use Reporting data. The authors measured traits in 247 children based on the parents’ and teachers’ reporting on the child’s performance at ages 7, 10.5, 12, and 14 years on tests specifically aimed at evaluating a child’s ability to use facial expressions and to recognize the mental state of others. Results showed that prenatal exposure to DAPs was associated with unfavorable social behavior reported by parents and teachers. DAP levels were also associated with a 2.7-fold increase in social responsiveness among children at 14 years of age (β=2.7, 95% CI 0.9–4.5). The SRS subscales are social awareness, social cognition, social communication, social motivation, and restricted interests and repetitive behavior. Normal scores are below 59, and scores above 75 are interpreted as indicating severe deficiencies in reciprocal social behaviors. No association was found between proximity to an area where OPs were used during pregnancy and ASD traits in children. In short, this study presented inconsistent results on the association between OP pesticides and ASD-related traits.
The committee considered two studies that examined pregnant women recruited into the HOME study (Donauer et al., 2016; Millenson et al., 2017). Urine from 327 women was collected at about gestation week 16 and again around gestation week 26 and analyzed for metabolites of OPs. Children were administered the BSID-II at 1, 2, and 3 years of age; the Clinical Evaluation of Language Fundamentals at age 4; and the Weschsler Preschool and Primary Scale of Intelligence at age 5 (Donauer et al., 2016). No association was found between prenatal exposure and cognition at any age from 1 through 5 after adjustment for relevant covariates. Millenson et al. (2017) further examined the children for PON1 polymorphisms from cord blood. Children were assessed with the SRS. Total DAP levels were not associated with changes in SRS scores (β= –1.2, 95% CI –4.0–1.6), nor were the associations modified by the child’s PON1 genotype.
Engel et al. (2016) attempted to pool data from the CHAMACOS study, the HOME study in Cincinnati, and the CCCEH and Mount Sinai cohort studies in New York City. Those researchers found significant heterogeneity for estimates of the association between DAPs and MDI (p=0.09) and also found that PON1 genotypes, ethnicity, and the center performing the study all influenced the results. Despite the limitations, the authors concluded that some population subgroups were at risk of neurodevelopment problems as shown by changes in the MDI.
The Childhood Autism Risk from Genetics and Environment (CHARGE) study in California also attempted to link pesticide use near the mother’s residence to children having ASD (n=486) or developmental disorder (n=168) (Shelton et al., 2014). Each pregnancy was given an exposure profile based
on the pesticide applications made near the mother’s home and the days of her pregnancy on which those applications occurred. Children were eligible for the study if they were 2–5 years of age (mean 36–38 months). Any prenatal proximity (within 1.25 km) to OPs was associated with an increased risk for ASD (OR=1.6, 95% CI 1.02–2.51); the risks were higher the later in the pregnancy the exposure occurred, but except for third-trimester exposures, those risks were not significant (OR=1.37, 95% CI 0.76–2.50 for preconception exposure; OR=1.53, 95% CI 0.87–2.68 for first trimester; OR=1.57, 95% CI 0.87–2.83 for second trimester; and OR=1.99, 95% CI 1.11–3.56 for third trimester). For developmental disorder and residential proximity to agricultural pesticide application the ORs were 1.23 (95% CI 0.65–2.31) for all pregnancies, 1.20 (95% CI 0.54–2.65) for preconception exposure, 1.29 (95% CI 0.60–2.79) for the first trimester, 1.62 (95% CI 0.75–3.48) for the second trimester, and 1.10 (95% CI 0.46–2.67) for the third trimester. The authors concluded that prenatal exposure to OPs could be linked with neurodevelopmental disorders.
Several studies in China assessed the effect of prenatal exposure to OPs on child neurodevelopment. Zhang et al. (2014a) administered the Neonatal Behavioral Neurological Assessment (NBNA) to 3-day-old infants in Shenyang, China. Urine samples of 249 pregnant women were analyzed for five nonspecific OP metabolites. Higher prenatal DAP concentrations were associated with lower scores in all NBNA scales, especially the Behavior Scale (β for a 10-fold increase in concentration= –0.65, 95% CI –0.84– –0.45, p=0.01). In a further assessment of the effects of prenatal exposure to over 30 OPs on child neurodevelopment, Silver et al. (2017) analyzed cord blood from 237 Chinese infants born in Fuyang Maternal and Children’s Hospital between 2008 and 2011. Exposure to chlorpyrifos, as determined by gas chromatography, was associated with significant motor function deficits in 199 infants assessed at 6 weeks and 9 months of age using the Peabody Developmental Motor Scales, 2nd edition. In another study, Wang et al. (2017) measured six DAP metabolites in urine specimens from 436 women taken at delivery. Women were part of the Laizhou Wan Birth Cohort study in Shandong, China. Children were assessed for neurodevelopment at 12 months (n=297) and 24 months (n=262). Prenatal OP concentrations were not correlated with any neurodevelopmental deficits, as measured with the Gesell Developmental Schedules, at 12 months. At 24 months, however, prenatal, but not postnatal, OP exposure was associated with a reduced developmental quotient, especially among boys.
González-Alzaga et al. (2015) used the WISC-IV to study the effects of high and low exposure to OP applications on cognitive functioning in 305 6- to 11-year-old children in Spain. Maternal prenatal residential exposure to OPs was associated with a poorer score on the processing speed subtest for boys (β= –5.7; 95% CI –9.6– –1.8), but not for girls (β= –1.6, 95% CI –6.4– –3.3). No other associations were significant for either high or low exposures or any of the WISC-IV subtests. The authors concluded that there was a weak association between prenatal OP exposure and neurological deficits.
The Early Life Exposures in Mexico to Environmental Toxicants (ELEMENT) study measured metabolites in maternal urine to assess prenatal exposure to chlorpyrifos and ADHD in children ages 6–11 years old in Mexico City (Fortenberry et al., 2014). There were no significant associations between chlorpyrifos metabolite concentrations in maternal urine and ADHD outcomes in the children.
A recent integrative review by Burke et al. (2017) examined the developmental neurotoxicity of OPs with an emphasis on chlorpyrifos and summarized the extensive epidemiologic studies of chlorpyrifos (most of which are discussed earlier in this section). The authors concluded that a “disruption of the structural integrity of the brain can be an important determinant of the cognitive deficits” associated with prenatal exposure to the pesticide. The authors also considered eight studies of cognitive and locomotor phenotypes following prenatal exposure to chlorpyrifos in rats and mice. The authors argued that “it is unclear to what extent, if any, AChE inhibition contributes to the developmental neurotoxicity of chlorpyrifos” and concluded that this made it difficult to extrapolate from animal models to humans.
Respiratory effects in children following prenatal exposures were not reported in Gulf War and Health, Volume 2, nor were they discussed in EPA’s 2016 Chlorpyrifos: Revised Human Health Risk Assessment for Registration Review, ATSDR’s 2008 Toxicological Profile for Diazinon, or ATSDR’s 1997 Toxicological Profile for Dichlorvos.
The Volume 11 committee identified two recent publications by Raanan et al. (2015, 2016) that assessed respiratory effects in young children after prenatal exposure to OPs and other pesticides in the CHAMACOS cohort. The assessments of maternal exposures to OPs were based on an analysis of DAPs in maternal urine at gestation weeks 13 and 26. Mothers reported their children’s respiratory symptoms at 5 and 7 years of age, with 359 children assessed. Children’s respiratory symptoms were not associated with any maternal DAP concentrations in the first half of pregnancy. However, there were significantly increased risks of respiratory symptoms in children whose mothers had elevated prenatal total DAPs and diethyl metabolite levels during the second half of pregnancy (ΣDAPs OR for a 10-fold increase in concentration=1.77, 95% CI 1.06–2.95; DEPs OR=1.61, 95% CI 1.08–2.39); there was no increased risk observed for elevated dimethyl metabolites (Raanan et al., 2015). A spirometric assessment of lung function in the 7-year-old children found that the mothers’ DAP urinary concentrations, whether in the first half or second half of pregnancy, were not significantly associated with lung function measurements (FEV1, FVC, FEV1/FVC ratio, and FEF25–75) in their children (Raanan et al., 2016).
Genotoxicity studies of OPs reveal primarily chromosomal abnormalities and DNA fragmentation. IARC has classified two of the Gulf War OPs—diazinon and malathion—as probable human carcinogens (IARC, 2016a,b). IARC has also classified dichlorvos as a possible human carcinogen (IARC, 1991). No transplacental cancer studies were reviewed for any of the three OPs.
The committee found only one new study that assessed the risk of childhood cancers and exposure to OPs before birth. Monge et al. (2007) examined the effects of maternal exposures (n=876) and paternal exposures (n=762) to pesticides on the risk of childhood leukemia (n=334) in Costa Rica in 1995–2000. Parents were interviewed about risk factors for childhood leukemia, and their exposures were rated as to hazard value at the time of occurrence. The mothers’ exposure to OPs in the first trimester of pregnancy was significantly associated with total leukemia (OR=3.5, 95% CI 1.0–12.2) and acute lymphoblastic leukemia (ALL) (OR=3.7, 95% CI 1.0–13.1). Although the risk of any leukemia and ALL was elevated in children whose mothers were exposed in the year before conception (leukemia OR=2.3, 95% CI 0.9–6.0; ALL OR=2.5, 95% CI 0.9–6.7) and during the second trimester (leukemia OR=2.5, 95% CI 0.7–9.5; ALL OR=3.0, 95% CI 0.8–11.5), the increases were not significant. In the fathers, exposure to OPs was associated with a increased risk of leukemia at all time points: the year before conception (OR=1.5, 95% CI 1.0–2.2), the first trimester (OR=1.6, 95% CI 1.0–2.6), and the second trimester (OR=1.4, 95% CI 0.8–2.3). Further analysis of the fathers’ exposures to specific pesticides showed that exposure to malathion in the year before conception was associated with an increased risk of leukemia and of ALL for boys (OR=8.5, 95% CI 1.1–74.1 and OR=10.4, 95% CI 1.2–91.1, respectively), but not for girls (OR=0.9, 95% CI 0.2–4.9 and OR=0.5, 95% CI 0.1–4.8, respectively).
Other Developmental Effects
Carmichael et al. (2014) examined the association between exposure to pesticides during pregnancy and the prevalence of eight types of congenital heart defects in a population-based case-control study. Pesticide exposure within a 500-meter radius of each mother’s residence was estimated for the 3 months
prior to conception to delivery using geospatial coding for eight San Joaquin counties. The California Pesticide Use Reporting records were used to identify any pesticides applied at more than 100 lbs per year in any of the eight counties. Using data from the California Birth Defects Monitoring Program from 1997 to 2006, the exposures of the mothers of 569 case infants was compared with 785 mothers who had an infant without a birth defect. Prenatal exposure to chlorpyrifos was significantly associated with pulmonary valve stenosis (OR=2.4, 95% CI 1.0–6.3) and atrial septal defects, secundum (OR=1.9, 95% CI 1.1–3.4). A further assessment for five types of other birth defects was conducted (Carmichael et al., 2016), based on the exposure of 367 mothers of case infants compared with the same 785 control mothers. Prenatal exposure to OPs as a group (based on mechanism of action) was associated with a reduced risk of anorectal atresia/stenosis (OR=0.4, 95% CI 0.2–0.9) but was significantly associated with an increased risk of craniosynostosis (OR=1.9, 95% CI 1.1–3.2). Chlorpyrifos specifically was significantly associated with anotia (OR=2.0, 95% CI 1.1–3.8) and craniosynostosis (OR=2.4, 95% CI 1.2–4.6). The ORs for all other OPs or birth defects were not significant and ranged from 0.5 to 2.0.
Buscail et al. (2015) used the French PELAGIE mother–child cohort to examine the prevalence of otitis media in children at 2 years of age. They found that the children of 246 mothers who had exposure to OPs, determined by an analysis of DAP, DMP, and DEP metabolites in the mother’s urine taken before the 19th week of pregnancy, were not at an increased risk of having one or more occurrences of otitis media before age 2 years.
The Volume 11 committee identified one small study on epigenetic effects associated with prenatal exposure to pesticides. Declerck et al. (2017) examined 25 children born in Denmark between 1996 and 2001 to mothers who had occupational exposures to a variety of pesticides during the first trimester of pregnancy and 23 unexposed control children. Blood samples taken from the children at 6–11 years of age were analyzed for the PON1 Q192R genotype and genome-wide DNA methylation patterns related to prenatal pesticide exposure was determined. The majority of mothers (91%) were exposed to organophosphate pesticides, including chlorpyrifos, dichlorvos, and dimethoate, as well as pyrethroid pesticides (e.g., deltamethrin) and carbamate pesticides (e.g., methomyl). Prenatal pesticide exposure was found to be associated with changes in DNA methylation patterns in children with the PON1 192R-allele compared with the 192Q-genotype and the control children, and the hypermethylation changes were significantly associated with higher body fat content (p<0.001) and serum leptin concentrations (p=0.02).
In studies of pregnant women exposed to toxicants, including OPs, there are three confounding issues: exposures are seldom limited to one insecticide (i.e., exposures are typically to mixtures of toxicants); exposures are rarely limited to one period of time (i.e., preconception, pregnancy or trimester of pregnancy, or postnatal); and outcome assessments are typically made in early childhood and not at later stages of development, such as adolescence. These issues can be addressed in animal models. Socioeconomic status can also be a confounder in human studies, but it is not in animal studies.
The ATSDR Toxicological Profile for Malathion (2003a) indicated that malathion was not a developmental toxicant in animals at doses that did not induce maternal toxicity.
Animal models have been used extensively to study sensitive periods for chlorpyrifos exposure in pregnancy. This is important for female veterans whose deployments are terminated shortly after pregnancy is identified. Researchers using rodent models to compare the effects of chlorpyrifos exposure in early and late pregnancy found that early pregnancy exposure had various long-lasting effects on brain development in rats, including altered behavior (Icenogle et al., 2004), brain morphology (Chen et al., 2017; Qiao et al., 2004), modified brain chemistry (Aldridge et al., 2003, 2004, 2005; Meyer et
al., 2004a,b; Qiao et al., 2002), and altered brain development (Qiao et al., 2003). For example, adult rats that had been exposed to chlorpyrifos on gestational days (GDs) 9–12 or 17–20 showed significant changes in the adenylyl cyclase signaling cascade, which mediates the cellular responses to numerous neurotransmitters, in a wide variety of brain regions in adulthood (Meyer et al., 2004a). However, exposure during GD 9–12 was less sensitive and required doses that resulted in decreased maternal weight gain. In another study, early pregnancy exposure resulted in cardiac and hepatic effects in offspring (Meyer et al., 2004b).
Exposure to chlorpyrifos later in pregnancy (GDs 14–20) also produced anxiety-like symptoms in rats assessed on postnatal day (PND) 21, but the anxious behavior was reversed by PND 70 (Silva et al., 2017). There were no depressive symptoms.
Studies of mid- to late-pregnancy exposure (GDs 14–17) of CD-1 mice to chlorpyrifos at a dose of 6 mg/kg resulted in early effects on pup vocalizations (Venerosi et al., 2009) and, later, alterations in social vocalizations and social investigation in female offspring (Venerosi et al., 2006, 2015). There was “specific and sex-dependent vulnerability of affective/emotional domains” and a disruption of serotoninergic transmission in both male and female animals (Venerosi et al., 2010). Lan et al. (2017) also treated pregnant mice in mid-gestation and found a delayed development of neonatal reflexes; by PND 90 the mice at the high dose had reduced social behavior, while the low-dose mice showed increased restricted interest behaviors. Thus, these animal studies demonstrate that chlorpyrifos exposure during sensitive brain development periods may affect various behavioral domains, including cognition, motor behavior, anxiety, and social behavior, and these detrimental neurodevelopmental effects may persist to adulthood.
While chlorpyrifos has been the most extensively studied OP with regard to neurodevelopmental effects, one study did demonstrate long-term effects on behavior in rodents following early pregnancy exposure to dichlorvos (Lazarini et al., 2004).
The associations between early pregnancy exposure to OPs and malformations and transplacental cancer have also been studied. All registered pesticides are assessed for teratogenicity before EPA approval. For malathion, diazinon, and dichlorvos there are published studies that reported finding no teratogenesis (ATSDR 1997b, 2003a, 2008). Recently, studies using chlorpyrifos doses with high systemic toxicity during embryogenesis did produce malformations (Akhtar et al., 2006), and studies testing mixtures of OPs, or mixtures including OPs with other pesticides, were also positive (Gomes and Lloyd, 2009; Yu et al., 2017). At doses that produce high systemic toxicity, pregnancy outcome measures (resorption, fetal loss, neonatal death, growth retardation) were also affected by OPs. However, it is important to note that the studies that identified neurodevelopmental toxicity of OPs used doses below the threshold for systemic maternal toxicity.
Oxidative damage has been implicated as a possible mechanism of action for the developmental effects of OPs (Qiao et al., 2004). While it has been demonstrated that antioxidants protect against the reproductive effects of OPs in males (Mosbah et al., 2016), this has not been studied for neurodevelopmental effects in offspring.
Synthesis and Conclusions
Chlorpyrifos exposure has been shown to result in significant effects on sperm quality and lead to significantly increased levels of disomy in sperm chromosomes (Figueroa et al., 2015; Martenies and Perry, 2013; Meeker et al., 2004a,b; Melgarejo et al., 2015; Miranda-Contreras et al., 2013; Sánchez-Peña et al., 2004). Occupational exposure to OPs has been associated with altered sperm integrity,
specifically the sperm chromatin structure, although the association was only with the presence of the OP metabolite DETP (Sánchez-Peña et al., 2004). Adverse effects on sperm parameters and some reproductive hormones were also seen in agricultural workers with OP exposure, although the men were also exposed to other pesticides (Miranda-Contreras et al., 2013). Similar adverse effects on sperm parameters were seen in men with environmental exposures to OPs, although reproductive hormones were largely unaffected (Melgarejo et al., 2015). A systematic review of six studies assessing the effects of exposure to OPs on male reproduction concluded that all six studies reported significant negative associations between OP exposure, whether occupational or otherwise, and at least one semen parameter. However, the Volume 11 committee noted that in many of these studies exposure was assessed using a single nonspecific urinary measure of the metabolites to which most OPs devolve, and, as indicated above, these biomarkers have some inherent limitations (e.g., they reflect only recent exposure, they may reflect exposure to preformed metabolites, etc.). Meeker et al. (2004a,b) examined six sperm parameters and reported no significant associations between urinary levels of TCPy, a metabolite of chlorpyrifos, and sperm concentration, motility, morphology, or percentage DNA tail or comet extent in sperm; only a decrease in tail distributed moment reached significance.
Recent rodent studies also focused on the testicular toxicity of OPs. These studies treated rodents with a pesticide, usually over at least one sperm cycle, and found effects on sperm quality, serum hormones, testes and epididymis histopathology, biochemistry, and gene expression. Such rodent studies have been conducted for all the OPs recognized for Gulf War toxicants: chlorpyrifos (Mandal and Das, 2012; Sai et al., 2014), dichlorvos (Dirican and Kalender, 2012), diazinon (Leong et al., 2013; Zidan, 2009), and malathion (Bustos-Obregón and González-Hormazabal, 2003; Geng et al., 2015; Uzun et al., 2009). The Volume 11 committee found that these animal studies described findings that were supportive of male reproductive effects being associated with OPs. These studies used measures similar to those used in human studies (serum hormones, sperm quality, sperm DNA damage) and were able to add testes histopathologic and biochemical measures. Although sperm production is affected by chlorpyrifos exposure in both animal and human studies, the effects on spermatogenesis and male fertility after the discontinuation of exposure have not been elucidated in humans. The Volume 11 committee concludes that these animal studies are supportive of male reproductive effects associated with OPs in human studies, as both human and animal studies assess the same sperm parameters. Additionally, animal studies of ROS generation as the mechanism of testicular toxicity provide biological plausibility for the associations reported in human studies.
The Volume 11 committee identified two studies that assessed the effects of OP exposure on female reproduction. Although the study in China (Hu et al., 2018) showed increased TTP and reduced fertility, the second study did not show decreased fertility. OP effects on ovarian function either during or after exposure have not been examined in any human studies and have been examined in only a few animal studies.
The Volume 11 committee considered several major longitudinal U.S. birth cohorts—the CCCEH study in New York City, the CHAMACOS study in California, the Children’s Environmental Cohort Study at Mt. Sinai Hospital in New York City, and the HOME study in Cincinnati, Ohio—that examined the association between prenatal exposure to one or more pesticides, including OPs, and birth outcomes.
The effects of prenatal exposure to chlorpyrifos specifically, or OPs in general, on birth outcomes were not consistent across the studies. In particular, the Volume 11 committee notes that there were ethnic and sex differences for several of the outcomes related to specific DAPs. The Harley et al. (2016) pooled analysis found no significant associations between increases in maternal DAP concentrations and decreases in birth weight, birth length, or infant head circumference for white and Hispanic women, but it did find among black women that increased prenatal OP exposure was associated with decreased infant
birth weight, length, and head circumference. However, gestational duration could not be examined in this pooled analysis, and an association between OP exposure and shortened gestational duration was noted in two of the studies (CHAMACOS and HOME) that examined length of gestation (Eskenazi et al., 2004; Rauch et al., 2012). The Volume 11 committee also emphasizes that it was not possible in any of these cohort studies to say which particular OP the subjects were exposed to. Furthermore, maternal OP exposure may not be restricted to preconception, to the first trimester of pregnancy, or to pregnancy.
The Volume 11 committee found a lack of relevant information in the animal literature on adverse birth outcomes associated with preconception or prenatal exposure to any of the relevant OPs considered in this chapter.
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between exposure to OPs 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 OPs and reproductive effects in women, or with adverse pregnancy outcomes.
Research into the developmental effects in infants or children following prenatal exposure to OPs has focused primarily on neurodevelopmental effects, although effects on the respiratory system and the presence of childhood cancers have also been assessed (see Table 5-1).
The potential neurodevelopmental effects associated with prenatal exposure to OPs—and to chlorpyrifos in particular—have been the subject of several comprehensive reviews. In 2016 EPA concluded that the CCCEH study in New York City, along with other epidemiologic studies, provided “sufficient evidence that there are neurodevelopmental effects occurring at chlorpyrifos exposure levels below those required for AChE inhibition.” The Volume 11 committee reviewed a number of studies of child neurodevelopment and prenatal OP exposure conducted in Spain, Mexico, and China, as well as the CCCEH, CHAMACOS, and Mount Sinai, and HOME cohort studies. Results of many, but not all, of these studies indicate that prenatal exposure to OPs can result in lower scores on developmental indices such as IQ and other assessment scales and that these deficits may persist up to at least the 9th year of life. In addition, associations have been found with attention deficits and tremor. The CHARGE study in California found that close prenatal proximity to OP application was associated with an increased risk for ASD (Shelton et al., 2014), and Burke et al. (2017) stated that “the association between cord blood levels of chlorpyrifos and neurodevelopmental outcomes cannot be ignored.”
Assessment of the CHAMACOS cohort found increased respiratory symptoms, but no impairment of lung function, in children at 5 and 7 years of age as reported by mothers, although the increase was not associated with maternal DAP measures in the first half of pregnancy. Studies on the effects of prenatal OP exposure and the risks of childhood leukemia were inconsistent, although the one positive study examined both the father’s and the mother’s exposure prior to conception and in the first trimester. There was little information on the effects of prenatal OP exposure on the risk of birth defects or on other adverse childhood health effects.
In addition to the numerous epidemiologic studies discussed above, the Volume 11 committee considered more than 50 new animal studies, many of which discussed a variety of reproductive and developmental endpoints. Because of the large number of studies that assessed developmental effects from prenatal exposure to OPs, the committee described only those that specifically showed effects in early
pregnancy compared with effects seen later in pregnancy. Animal studies indicate that early pregnancy exposures produce developmental toxicity, but it is unclear whether these effects would be expected for human pregnancies initiated after deployment ends since organ systems in animals and humans develop at different times during gestation.
The Volume 11 committee concludes that there is sufficient evidence of an association between prenatal exposure to OPs and neurodevelopmental effects.
The Volume 11 committee also concludes that there is inadequate/insufficient evidence to determine whether an association exists between prenatal exposure to OPs and other developmental effects.
TABLE 5-1 Summary of Neurodevelopmental Effects of Organophosphate Pesticides
|Cohort||OP Exposure Measurements||Child’s Age at Time of Assessment||Neurodevelopmental Effects|
|CCCEH Rauh et al. (2006, 2012, 2015)||Cord blood at delivery||12, 24, 36 months||At 36 months, decrease in PDI scores and motor skills with increase in ADHD and pervasive development disorder.|
|7 years||Decrease in IQ and working memory; decrease greater in boys than girls. Adverse brain morphology.|
|Mt. Sinai Engel et al. (2007, 2011); Furlong et al. (2014)||Blood/urine in third trimester||<5 days||Increase in abnormal reflexes.|
|12, 24 months, 6–9 years||Decrease MDI in nonwhites at 12, 24 months; whites had increased MDI; all effects more modest by 7–9 years old.|
|7-9 years||No association of DAPs with Social Responsiveness Scale scores in whites and Hispanics; association with blacks; boys scored poorly compared with girls.|
|CHAMACOS Eskenazi et al. (2007, 2014); Young, et al. (2005); Marks et al. (2010); Bouchard et al. (2011); Eskenazi et al. (2014); Sagiv et al. (2018)||Urine in the third trimester||3 days||Increased DAPs associated with increase abnormal reflexes.|
|6, 12, 24 months||Increased DAPs associated with decreased MDI at 24 months; increased risk of pervasive development disorder, but no effect on ADHD.|
|3.5 and 5 years||ADHD not increased at 3.5 years but increased at 5 years; some effects associated only in boys.|
|7 years||Increased DAP associated with decreased cognitive and IQ scores.|
|6 months–14 years||DAPs showed association with poor social behavior and test performance. No association between proximity and ASD-related traits.|
|7, 9, 10.5, 12, and 14 years||Increased DAPs associated with poor social skills but no other consistent associations with ASD symptoms.|
|HOME Donauer et al. (2016); Millenson et al. (2017)||Urine at 16 wks pregnancy||1, 2, 3, 4, 5 years||No association of DAPs with cognition at 1–5 years; at 5 years increased DAPs associated with increased verbal IQ, but attenuated after adjustment for covariates.|
|8 years||No association of DAPs with child’s social behavior.|
|Cohort||OP Exposure Measurements||Child’s Age at Time of Assessment||Neurodevelopmental Effects|
|CHARGE study (Shelton et al., 2014)||Residential proximity to OP application||2–5 years||OPs in first trimester not associated with ASD, but increased risk for ASD with any OP exposure during pregnancy, highest for third-trimester exposures.|
|González-Alzaga et al. (2015) Spain||Urine sample prenatal||6–11 years||Increased prenatal OP exposure associated with decreased processing speed performance for boys but not girls.|
|Zhang et al. (2014a) China||Urine samples during pregnancy||3 days||Increased DAPs associated with decreased scores on neonatal behavioral neurological assessment test for both boys and girls.|
|Fortenberry et al. (2014) Mexico||Urine samples||6–11 years||No association of TCPy with ADHD.|
NOTE: ADHD=attention deficit–hyperactivity disorder; ASD=autism spectrum disorder; CCCEH=Columbia Center for Children’s Environmental Health study; CHAMACOS=Center for the Health Assessment of Mothers and Children of Salinas study; CHARGE=Childhood Autism Risk from Genetics and Environment; DAP=dialkylphosphates; DEP=diethylphosphate; HOME=Health Outcomes and Measures of the Environment study; IQ=intelligence quotient; MDI=mental development index; Mt. Sinai=Mount Sinai Children’s Environmental Health Cohort; OP=organophosphate; PDI=psychomotor development index; TCPy=3,5,6-trichloro2-pyridinol.
TABLE 5-2 Summary of Reproductive Effects of Organophosphate Pesticides
|Reproductive Effects in Men|
|Meeker et al. (2004a)||330 men in subfertile couples seeking infertility diagnosis at Massachusetts General Hospital (Boston) between January 2000 and April 2003.||No men reported occupational exposure to pesticides. A single spot urine sample was collected from each subject on the same day as the semen sample and analyzed for TCPy.||Suggestive but not statistically significant associations between TCPy levels with sperm concentration (medium level OR=2.1, 95% CI 0.8–5.6; high level OR=2.4, 95% CI 0.9–6.3; p-value for trend=0.09), sperm motility (medium OR=1.6, 95% CI 0.8–3.0; high OR=1.7, 95% CI 0.9–3.2; p-value for trend=0.09), and sperm morphology (medium OR=1.2, 95% CI 0.5–2.7; high OR=1.9, 95% CI 0.9–4.0; p-value for trend=0.10) compared to those with low TCPy levels.|
|Meeker et al. (2004b)||See Meeker et al. (2004a); semen samples available for 260 men; 46 samples excluded with specific gravity >1.03 or <1.01.||See Meeker et al. (2004a). TCPy measured in urine (geometric mean of 2.59 μg/L). Semen analyzed for DNA damage by the neutral comet assay and reported as comet extent, percentage DNA in comet tail (Tail%), and tail distributed moment (TDM).||A statistically significant increase in Tail% was found for an IQR increase in TCPy (β=2.76, 95% CI 0.89–4.62), while a decrease in TDM was associated with IQR changes in TCPy (β= –2.5, 95% CI –4.71––0.22), the change in comet extent was not significant. A negative correlation between Tail% and TDM was present only when stratified by comet extent, suggesting that Tail% and TDM may measure different types of DNA damage within comet extent strata.|
|Figueroa et al. (2015)||Same population as Meeker et al. (2004a,b). 159 male subjects.||See Meeker et al. (2004a).||Of the 159 men, 10% (n=16) had sperm concentrations <15 million/mL, 21% (n=33) had <32% motile sperm, and 18% (n=28) had <4% normally shaped sperm based on the most recent WHO reference values for sperm concentration and the Tygerberg Strict Criteria for morphology. A significant inverse association was observed between DMP and total disomy.|
|Sánchez-Peña et al. (2004)||33 male agricultural workers in Durango, Mexico, ages 18–50.||Men provided semen and urine samples. Urine samples were analyzed for 6 DAPs: DMDTP, DMTP, DEDTP, DETP, DMP, and DEP. 87% of workers had at least one OP metabolite in their urine.||Most sperm samples had altered chromatin structure. About 75% of semen samples were classified as having poor fertility potential (>30% of DFI%) and 12% were classified as potentially fertile (<15% of DFI%), whereas only 4% of unexposed urban men were classified as having poor fertility potential. There were significant associations between urinary DETP concentrations and mean DFI (p=0.026), and standard deviation-DFI (p=0.022); no associations with any other DAP were found. 82% of exposed men versus 9% of unexposed men had immature germ cell values above the reference value. There were no significant associations found between DAP and semen quality parameters, although there was a negative correlation between mean DFI and sperm viability (r=0.254, p<0.027) and seminal volume (r=0.333; p<0.004).|
|Melgarejo et al. (2015)||Cross-sectional study in Murcia Region, Spain. Subjects recruited between December 2012 and July 2013. 162 men 25–38 years old (median age 35); 116 returned for semen analysis, had blood drawn, gave urine sample, and completed questionnaires.||Length of exposure not specified. Men attending infertility clinic. Urine analyzed for 6 DAPs.||At least one urinary OP metabolite detected for all men. Urinary DMP concentrations were signficiantly inversely correlated with percentage of motile sperm (r= –0.23, 95% CI –0.34– –0.05) and sperm with normal morphology (r= –0.20, 95% CI –0.36– –0.02); no other associations. Reproductive hormone levels within normal range for all men, but positive associations between DEDTP concentrations and LH (β=11.4, 95% CI 0.81–22.1) and FSH levels (β=3.2, 95% CI 0.08–6.2) and a negative correlation with the testosterone/LH ratio. Significant inverse association between concentrations of DMP, DMTP, and DMDTP, and ΣDAP and sperm concentration and total sperm count; between DMTP, DMDTP, and DEP concentrations and motile sperm; and between DMP and DMDTP concentrations and total motile sperm count.|
|Miranda-Contreras et al. (2013)||64 male farm workers (ages 18–52 years), 35 control males from city 90 km away.||Agricultural community in Venezuela; mixed exposure to OPs, carbamates, pyrethroids, and triazines. Exposed lived in area ≥2 years prior to the study, 53.5% of exposed males had worked in agriculture for >5 years; 75% sprayed pesticides 1–3 times/wk.||Levels of BuChE activity indicated that 87.5% of the pesticide sprayers had been exposed to OP and carbamate pesticides; 82.8% had mild BuChE inhibition. Significant differences in 3 of 9 semen parameters for exposed group: higher seminal pH (p<0.004), lower percent live sperm (p<0.0001), and rapid and progressive sperm motility (p < 0.04); the levels of testosterone, prolactin, FT4, and TSH appeared to be normal; a tendency for increased LH and FSH levels in the exposed group.|
|Perry et al. (2007)||Pilot biomonitoring study in 2004. 18 men living in Anhui, China; men were young and newly married.||Men lived in city and visited family in rural area where OPs and pyrethroids used. Urine samples analyzed for 24 pesticide metabolites, including 6 DAPs, TCPy, malathion dicarboxylic acid, 2-isopropyl-4-methyl-6-hydroxypyrimidinol (a diazinon metabolite); baseline sperm collected.||Significant difference in sperm concentration between men with high and low urine concentrations of DETP, but not DMTP (the only metabolites reported), with absolute sperm concentration difference for DETP= –1.0 (95% CI –1.8– –0.2).|
TABLE 5-2 Continued
|Perry et al. (2011)||Continuation of Perry et al. (2007); 94 men and 95 controls living in Anqing City, China, who returned to rural area in winter.||Baseline urine analyzed for 6 DAPs. Cases (those with sperm concentrations and motility below population median) and controls were based on their median residual value of sperm concentration and motility adjusted for covariates.||Lower sperm concentration and motility in men with higher urinary concentration of DMP in cases compared with men with lower urinary concentrations (OR=1.30, 95% CI 1.02–1.65). No significant differences were seen between cases and controls for the other DAPs.|
|Recio-Vega et al. (2008)||Longitudinal follow-up study. 52 men living in Durango, Mexico, who provided 2 or more semen samples (total 139 samples).||Men lived in agricultural area, 17 non-occupationally exposed, 16 occupationally exposed but not OP sprayers, and 19 OP sprayers (highly exposed). Semen collected at times of low, median, and high OP use. Urine collected morning of day semen collected and analyzed for 6 DAPs.||DAPs were found in urine of 83% of participants, and levels were highest during heavy spraying season. Only semen volume and total sperm count were significantly lower in the highly exposed group (p=0.004 and p=0.01, respectively), the other semen parameters (sperm concentration, viability, motility, dead sperm, and rapid progressive motility) were not significantly different.|
|Yucra et al. (2008)||Cross-sectional study of 31 male OP pesticide applicators in Peru (ages 20–60) and 31 unexposed controls.||Urine, blood, and semen samples collected day after last pesticide application; urine analyzed for 6 DAPs.||Nine semen parameters were analyzed. DEDTP (p=0.04) and DETP (p=0.02) were more related to occupational exposure than other OP metabolites. Seminal pH was the only parameter significantly associated with urinary concentrations of ethylated (p=0.02) or methylated (p=0.002) DAPs; methylated DAPs were also associated with lower seminal vesicle function (p=0.04, seminal fructose levels). Holm’s test used to correct for multiple comparisons.|
|Hossain et al. (2010)||Cross-sectional study of 62 male farmers in 3 communities in Malaysia with exposure to malathion (15 men) or paraquat (39 men) or both (8 men); 90 unexposed.||Pesticide exposure based on interviews with farmers; semen samples analyzed for volume, pH, sperm concentration, motility, morphology, and white blood count.||Results presented for exposed versus unexposed. The committee notes that grouping of all pesticide types together makes assessment impossible for malathion specifically. Exposed workers had a greater risk of abnormal sperm volume (OR=6.5, 95% CI 2.7–15.2), sperm concentration (OR=8.7, 95% CI 4–19), lower sperm motility (OR=5.1, 95% CI 2.5–10.5), sperm morphology (OR=4.9, 95% CI 1.6–14.6), and white blood cell count (OR=4.5, 95% CI 1.5–13.4).|
|Perez-Herrera et al. (2008)||Cross-sectional study; 54 agricultural workers (18–55 years old) in Yucatan, Mexico.||Men had chronic, year-round exposure to pesticides, mostly OPs (including CPF, malathion, and diazinon). Semen and blood samples collected.||Abnormal sperm morphology was seen in all 54 participants; 46% had low ejaculate volume, 30% had low sperm motility, and 13% had low sperm viability; however, sperm parameters were not associated with OP exposure during the month of semen collection or the 3 months prior to collection. For men with 192RR phenotype, there were significant associations between OP exposure at the month of sampling and NT-positive cells and sperm viability and a dose–effect relationship between OP exposure during 3 months before sampling and sperm quality parameters and NT-positive cells.|
|Al-Hussaini et al. (2018)||94 women in Egypt undergoing in vitro fertilization.||Follicular fluid collected from each woman at time of embryo transfer and analyzed for CPF, diazinon, and malation among other pesticides.||All three pesticide concentrations in follicular fluid were negatively correlated with endometrial thickness (p<0.05 for diazinon, p<0.024 for malathion, and p<0.098 for CPF). Diazinon and CPF, but not malation, were negatively associated with the number of oocytes retrieved and with a lower implantation rate (number of sacs) (p<0.01 for both pesticides and endpoints); none of the three were associated with reduced fertilization or early embryo cleavage rates.|
|Hu et al. (2018)||615 women (≥20 years of age) in Shanghai, China, recruited from two preconception care clinics during 2013–2015. 569 women were nulliparous. Women followed for 1 year prior to pregnancy.||4 DAPs measured in urine. Pyrethroid exposure also measured. Women asked about pesticide exposure and perceived stress and lifestyle.||Women in the highest quartile for DETP had longer TTP (fecundability OR=0.68, 95% CI 0.51–0.92) and increased infertility (OR=2.17, 95% CI 1.19–3.93) compared with women in lowest quartile. For nulliparous women only, the fecundability OR for TTP was 0.67 (95% CI 0.50–0.91), and the fertility OR was 2.30 (95% CI 1.23–4.30). There were no significant associations with the other 3 DAPs.|
|Adverse Pregnancy Outcomes|
|Perera et al. (2003)||CCCEH cohort of 269 African American and Dominican women, 18–35 years old, enrolled into the cohort between 1998 and 2002.||Women wore personal monitor during the daytime hours for 2 consecutive days and placed the monitor near the bed at night. Maternal blood (30–35 mL) was collected within 1 day postpartum, and umbilical cord blood (30–60 mL) was collected at delivery.||CPF was detected in 98% of the maternal and 94% of the cord samples. CPF was significantly associated with decreased birth weight overall (β= –0.04, p=0.01) and among African Americans (β= –0.05, p=0.04), but not among Dominicans (β= –0.03, p=0.11). CPF was also associated with reduced birth length overall (β= –0.01, p=0.003) and in Dominicans (β= –0.02, p<0.001), but it was not significantly associated with head circumference.|
TABLE 5-2 Continued
|Eskenazi et al. (2004)||CHAMACOS cohort of women living in agricultural area of Salinas Valley, CA; children born October 1999–October 2000; eligible n=488) women were ≥18 years of age, <20 weeks gestation at enrollment; DAP metabolites were measured in urine from mothers twice during pregnancy (mean=13 and 26 weeks gestation).||Gestational and preconception exposure, half of the women resided in United States for <5 years; approximately 28% had worked in the fields during pregnancy, another 14% worked at other agriculture jobs (e.g., packing shed, nursery, and greenhouse work). 85% of the women had agricultural workers living in their homes during their pregnancy.
Exposure to OPs was measured in 3 ways: (1) DAP metabolites in maternal urine during pregnancy; (2) 7 different pesticide-specific metabolites in maternal urine during pregnancy; and (3) ChE in whole blood and BuChE in plasma collected from mothers during pregnancy and at delivery, and from the umbilical cord.
|A 10-fold increase in average DAP metabolite concentration was associated with an increase in infant’s body length of 0.52 cm (p=0.06) and in head circumference of 0.32 cm (p=0.03); similar increases in body length and head circumference were seen when DMP and DEP were examined separately, although these increases were not statistically significant.
A 10-fold increase in average DMP but not DEP was associated with a 3-day decrease in gestational duration (p=0.02). After 22 weeks gestation, increasing levels of DMP had significant adverse association with gestational duration. DMP, DEP, and total DAP levels were not associated with birth weight, infant ponderal index, risk of preterm delivery, low birth weight, or small for gestational age births. Lower ChE levels in cord blood were associated with significantly shorter length of gestation, averaging 0.34 weeks (p=0.001) for each unit decrease in ChE (in moles/m/mL; range of ChE in cord blood is 4.4 units), with an increased risk of preterm delivery (OR=2.3, 95% CI 1.1–4.8, p=0.02), and with low birth weight (OR=4.3, 95% CI 1.1–17.5, p=0.04); however, 6 of the 11 low-birth-weight infants were also preterm. BuChE levels in maternal and cord blood were not associated with any birth outcome.
|Harley et al. (2011)||CHAMACOS cohort; see Eskenazi et al. (2004).||See Eskenazi et al. (2004).||All women had detectable levels of OP pesticide metabolites in their urine during pregnancy; average DAP concentration during pregnancy was 146 nmol/L (95% CI 133–160); associations of PON1 with birth outcome were found only with PON1 measured in infants, not in mothers. No associations were found between any marker of PON1 genotype or activity measured in maternal blood and gestational age or infant birth weight, length, and head circumference. Each 10-fold increase in maternal DAPs during pregnancy was associated with decrease in gestational age among PON1-108TT infants (β= –0.5, 95% CI –0.9–0.0) compared with PON1192QQ infants.|
|Rauch et al. (2012)||HOME Study, prospective birth cohort in Cincinnati area between March 2003 and January 2006. 344 women who met eligibility criteria.||Continuous residential exposure. Mothers gave blood and spot urine samples at 16 and 26 weeks gestation, and within 24 hours of delivery.
Measured 6 urinary DAPs.
|All mothers had detectable levels of at least one DAP metabolite. A 10-fold increase in ΣDAP concentrations was associated with a 0.5-week decrease in gestational age (95% CI –0.8– –0.1]); birth weight was also inversely associated with ΣDAP concentrations (–151 g, 95% CI –287––16); no evidence for main effects of PON1 genotype on either birth weight or length of gestation.|
|Berkowitz et al. (2004)||Mount Sinai Children’s Environmental Cohort Study of ethnically diverse mother–infant pairs in New York City. Babies born between May 1998 and July 2002 (n=404); maternal blood and urine samples taken at 26 weeks gestation.||Mothers in third trimester answered questionnaire about pesticide and other environmental exposures. 46.2% of women reported that they or a household member had used indoor pesticides during the pregnancy. Cord blood samples used to determine maternal and infant PON1 activity and polymorphisms.||No significant associations between adjusted birth outcomes and TCPy, PBA, or PCP levels above and below the LOD. A significant positive trend was found between maternal PON1 activity and head circumference among the offspring of mothers whose TCPy levels were above the LOD. A similar trend was seen for mothers with TCPy levels below the LOD, but this was not significant. The test for interaction among TCPy level, PON1 activity, and head circumference was not statistically significant (p>0.05). PON1 activity had no association with any of the fetal growth measures.|
|Harley et al. (2016)||See Engel et al. (2016);
(1) CHAMACOS, (2) HOME, (3) CCCEH, and (4) Mt. Sinai birth cohorts.
|See Engel et al. (2016).||No significant associations of ΣDEP, ΣDMP, or ΣDAPs with birth weight, length, or head circumference overall. Among non-Hispanic black women, increasing urinary ΣDAP and ΣDMP concentrations were associated with decreased birth length (β= –0.4 cm; 95% CI –0.9–0.0 and β= –0.4 cm; 95% CI –0.8–0.0, respectively, for each 10-fold increase in metabolite concentration). Among infants with the PON1192RR genotype, ΣDAP and ΣDMP were negatively associated with length (β= –0.4 cm; 95% CI –0.9–0.0 and β= –0.5 cm; 95% CI –0.9– –0.1).|
TABLE 5-2 Continued
|Wang et al. (2012)||187 healthy pregnant women from Shanghai, China, were recruited to participate in this study from September 1, 2006 to January 31, 2007.||Information about pesticide use for a woman herself or other household members during pregnancy and, if so, the types of pesticides and frequency of use; paternal OP exposure and occupation were also collected through personal interview. Urine samples collected at onset of labor; urine analyzed for 5 DAPs.||No significant associations between fetal growth (assessed as birth weight and body length) or length of gestation and any measure of in utero OP exposure. However, a significant inverse association was found between DEP level and duration of gestation (β= −1.79; p=0.001) for infant girls after stratification by gender; this association was not seen for boys (β=0.17; p=0.164).|
NOTE: BuChE=butyryl cholinesterase; CCCEH=Columbia Center for Children’s Environmental Health study; CHAMACOS=Center for the Health Assessment of Mothers and Children of Salinas study; ChE=cholinesterase; CI=confidence interval; CPF=chlorpyrifos; DAP=dialkyl phosphate; DEDTP=diethyldithiophosphate; DEP=diethylphosphate; DETP=diethylthiophosphate; DFI=DNA fragmentation index; DMDTP=dimethyldithiophosphate; DMP=dimethylphosphate; DMTP=dimethyl thiophosphate; DNA=deoxyribonucleic acid; FT4=free thyroxine; FSH=follicle-stimulating hormone; HOME=Health Outcomes and Measures of the Environment study; IQR=interquartile range; LH=luteinizing hormone; LOD=level of detection; NT=nick translation; OP=organophosphate; OR=odds ratio; PBA=3-phenoxybenzoic acid; PCP=pentachlorophenol; PON1=paraoxonase; TCPy=3,5,6-trichloro-2-pyridinol; TDM=tail distributed moment; TTP=time to pregnancy; TSH=thyroidstimulating hormone; WHO=World Health Organization.
TABLE 5-3 Summary of Developmental Effects of Organophosphate Pesticides
|Rauh et al. (2012)||CCCEH cohort, 40 children.||Subgroup of 20 children who had CPF exposure levels in the upper tertile of distribution and 20 children from the lowest tertile. All children had no/very low prenatal ETS exposure and low prenatal PAH exposure levels measured by third-trimester personal monitoring of mothers; children had usable MRI imaging data and 7-year cognitive testing using the WISC-IV.||Overall brain size did not differ significantly across exposure groups, adjusted for age, sex, and height; high-CPF group had significant enlargement of several brain areas: superior temporal, posterior middle temporal, and inferior postcentral gyri bilaterally, and enlarged superior frontal gyrus, gyrus rectus, cuneus, and precuneus along the medial wall of the right hemisphere. Group differences were derived from exposure effects on underlying white matter.
Significant exposure × IQ interaction was derived from CPF disruption of normal IQ associations with surface measures in low-exposure children. High-exposure children did not show expected sex differences in the right inferior parietal lobule and superior marginal gyrus, and they displayed reversal of sex differences in the right medial superior frontal gyrus; they also showed frontal and parietal cortical thinning and an inverse dose–response relationship between CPF and cortical thickness.
|Horton et al. (2012)||CCCEH cohort study. Subjects included 335 mother–child pairs selected from an ongoing prospective cohort study.
Pregnant women ages 18–35 years who self-identified as either African American or Dominican.
|Children assessed at age 7 with WISC-IV. Children’s home environments were evaluated at 3 years of age; maternal prenatal urine samples from third trimester.||Significant interaction between prenatal exposure to CPF and child sex (β= –1.714, 95% CI –3.753–0.326), suggesting males experience a greater decrement in working memory than females following prenatal CPF exposure.|
TABLE 5-3 Continued
|Engel et al. (2007)||Mount Sinai Children’s Environmental Health Cohort is a prospective multiethnic cohort study; 404 mother–infant pairs were recruited during pregnancy between May 1998 and July 2001.||Mothers answered exposure questionnaire and gave blood and urine samples about gestational week 31.2; neonatal behavior and/or primitive reflexes measured with BNBAS w/in 5 days of birth; mother’s urine analyzed for MDA and 6 DAPs: DMDTP, DMTP, DEDTP, DETP, DMP, and DEP.||Urine MDA levels above the LOD and ΣDEP metabolites were associated with RRs of 2.24 (95% CI 1.55–3.24) and 1.49 (95% CI 1.1–1.98), respectively, for the number of abnormal reflexes. Mothers with elevated MDA levels delivered infants who were 3.6 times more likely to have at least 2 abnormal reflexes (95% CI 1.5–8.8), and babies born to mothers with elevated ΣDEP were 2.3 times more likely to have at least 2 abnormal reflexes (95% CI 1.1–5.0). Higher levels of ΣDEP and ΣDAPs were associated with increase in abnormal reflexes. Strong interaction between PON1 expression levels and ΣDMP on risk of abnormal reflexes: infants born to women in the first (interaction p=0.002) and second (interaction p=0.01) tertiles (slower metabolizers) had a greater risk of abnormal reflexes than infants in the highest tertile (fast metabolizers). For women in the highest tertile of PON1 expression, no increased risk of abnormal reflexes with increasing exposure was found. There was no interaction between ΣDEP and PON1.|
|Furlong et al. (2017a)||Mt. Sinai cohort study; see Engel et al. (2007).||Children assessed at 6–9 years using WPPSI-III at age 6 and WISC-IV between the ages of 7 and 9 years. Parents completed the BRIEF questionnaire and BASC at child’s age 6 and 7–9 years.||Factor analysis used to derive 7 neurodevelopment phenotypes. Maternal DMP metabolites in urine were negatively associated with one phenotype internalizing factor (β= –0.13, 95% CI –0.26–0.00) but positively associated with the executive functioning factor (β=0.18, 95% CI 0.04–0.31). However, DEP metabolites were negatively associated with the working memory index of the WISC-IV (β= –0.17, 95% CI –0.33– –0.03).|
|Marks et al. (2010)||CHAMACOS cohort, see Eskenazi et al. (2004).
Children were assessed at ages 3.5 years (n=331) and 5 years (n=323).
|Mothers completed the CBCL. NEPSY-II visual attention subtest administered to children at 3.5 years and K-CPT at 5 years. The K-CPT yielded a standardized ADHD confidence index score. Psychometricians scored behavior of the 5-year-olds during testing using the Hillside Behavior Rating Scale.
Total DAPs were measured in maternal urine collected at two time points during pregnancy and in child urine collected at the 3.5-year and 5-year visits.
|Prenatal DAPs were not associated with maternal reports of attention problems and ADHD at age 3.5 years but were significantly related at age 5 years (CBCL attention problems: β=0.7 points, 95% CI 0.2–1.2; ADHD: β=1.3, 95% CI 0.4–2.1). Prenatal DAPs were associated with scores on the K-CPT ADHD confidence index >70th percentile (OR=5.1, 95% CI 1.7–15.7) and with a composite ADHD indicator of the various measures (OR=3.5, 95% CI 1.1–10.7). Some outcomes exhibited evidence of effect modification by sex, with associations found only among boys.|
|Bouchard et al. (2011)||CHAMACOS cohort; mothers of 329 7-year-old children were slightly older and breastfed longer than those from the initial cohort.||Most commonly used OP pesticides in the Salinas Valley are CPF and diazinon (which devolve to DE) and malathion and oxydemeton-methyl (which devolve to DM).
Maternal urine samples from first (median, 13th week of gestation) and second half (median, 26th week of gestation) of pregnancy analyzed for 3 DAPs.
Administered WISC-IV at the 7-year study visit.
|DAP levels were associated with lower cognitive scores on all subtests in children at 7 years old, but were most significant for verbal comprehension (13-wk β= –2.6, 95% CI –5.1– –0.1; 26-wk β= –3.1, 95% CI –6.4–0.2) and full-scale IQ (13-wk β= –2.4, 95% CI –4.9–0.2; 26-wk β= –3.5, 95% CI –6.9– –0.1). Higher prenatal DAP concentrations were associated with lower scores on all four cognitive domains, the strongest associations being for verbal comprehension (β for a 10-fold increase in concentration= –5.3; 95% CI –8.6– –2.0). A 10-fold increase in DAP concentrations was associated with 5.6-point decrease in full-scale IQ (95% CI –9.0––2.2).|
|Eskenazi et al. 2014||CHAMACOS cohort; 296 children assessed at 5 years of age and 327 assessed at 7 years of age.||Children administered the K-CPT at 5 years and the WISC-IV at 7 years; DAPs measured in maternal and child urine samples, and analyzed PON1192 and PON1−108 genotypes and enzyme activity [arylesterase (ARYase), paraoxonase (POase)] in maternal and child blood.||Maternal and child PON1 genotype was not related to performance on K-CPT or WISC. Pregnancy ARYase levels were positively associated with all WISC subscales (e.g., 4.0 point increase in full-scale IQ per ARYase increase, 95% CI 1.6–6.4). Pregnancy POase levels were positively associated with WISC processing speed only. The association between DAPs and full-scale IQ was strongest for children of mothers with lowest-tertile ARYase levels (p interaction=0.27).|
TABLE 5-3 Continued
|Huen et al. 2018||CHAMACOS cohort; 238 children assessed at 7 years of age||Cord blood and blood samples from children collected at 7 years old assessed for PON1 DNA methylation using Illumina 450K and EPIC BeadChip arrays; administered WISC-IV full-scale IQ and 4 composite measures (verbal comprehension, perceptual reasoning, working memory, and processing) at the 7-year study visit.||Cord blood PON1 DNA methylation CpG sites were not significantly associated with any decrements in the 5 neurobehavioral scores after adjustment for false discovery rate (p-values all >0.05).|
|Gunier et al. (2017)||CHAMACOS cohort, see Eskenazi et al. (2004)
283 mother–child pairs.
|Children assessed for neurodevelopment at 7 years old using WISC-IV.
Agricultural pesticide use within 1 km of mother’s residence during pregnancy estimated using California Pesticide Use Reporting data from 1999–2001. Maternal urine analyzed for 6 DAPs.
|Increased exposure to OPs during pregnancy was significantly associated with an estimated 2.2-point decrease in full-scale IQ (95% CI –3.9 – –0.5) and a 2.9-point decrease in verbal comprehension (95% CI –4.4– –1.3). Only increases in CPF and diazinon use were significantly associated with a decrease in verbal comprehension, malathion was not. There were no significant changes in other WISC domains.|
|Sagiv et al. (2018)||CHAMACOS cohort, see Eskenazi et al. (2004) 246 mother–child pairs.||Children assessed at 7, 10.5, and 14 years old. Parents reported child’s social behavior with SRS-2 at age 14 years and BASC-2 at 7, 10.5, and 14 years.
Teachers also used BASC-2 for 7-year-olds. At 9 years old, children tested with Evaluación Neuropsicológica Infantil Facial Expression Recognition Test and at age 12 years, the NEPSY-II Affect Recognition subtest.
|In 14-year-olds, a 10-fold increase in prenatal DAPs was associated with 2.7-point increase (95% CI 0.9–4.5) in parent-reported SRS-2 T-scores (a measure of ASD behaviors). Scores on the social communication and interaction test and the restricted and repetitive behaviors tests were not significantly associated with DAP concentrations.
Increased DAPs were associated with parent and teacher reports of BASC-2 social skills T-scores, which indicate poorer social skills. In 9-year- and 12-year-olds, DAPs were not associated with affect recognition measured with the Evaluación Neuropsicológica Infantil or the NEPSY-II, respectively. There were no consistent associations across sex for any of the outcomes.
|Donauer et al. (2016)||HOME study. Prospective pregnancy and birth cohort study. Subjects were 327 mother–infant pairs in Cincinnati, Ohio, between February 2003 and January 2006.||Spot urine samples were collected from the mothers at 2 time points about 16 and 26 weeks’ gestation. BSID-II MDI and PDI administered at ages 1, 2, and 3 years, the Clinical Evaluation of Language Fundamentals-Preschool, Second Edition, at age 4, and the WPPSI-III at age 5.||No associations between prenatal exposure to OPs and cognition at 1–5 years of age. At 2-year visit, total DMP and total DEP were associated with BSID-II MDI at the p<0.20 level after controlling for relevant covariates (β=0.002 and β=0.002). At 5 years, total DAP and total DM were positively associated with full-scale IQ at the p<0.20 level after controlling for relevant covariates (β=0.001 and β=0.002, respectively). Finally, total DAP, total DM, and total DE were positively associated with verbal IQ at the p<0.20 level after controlling for relevant covariates (β=0.003, β=0.003, and β= –0.027, respectively). No associations between prenatal exposure to OPs and cognition at 1–5 years of age.|
|Millenson et al. (2017)||Prospective cohort study. 224 pregnant women in the Cincinnati HOME study. Child PON1 polymorphisms determined from cord blood. Children 8 years old.||6 OP insecticide metabolite concentrations from two urine samples collected at ~16 and ~26 weeks gestation; analyzed for 6 DAPs. Children administered SRS.||ΣDAP concentrations were not associated with SRS scores (β= –1.2, 95% CI –4.0–1.6). Among 8-year-old children with the PON1–108TT genotype, ΣDAP concentrations were associated with 2.5-point higher (95% CI –4.9–9.8) SRS scores; however, the association was not different from the 1.8-point decrease (95% CI –5.8–2.2) among children with PON1–108CT/CC genotypes (ΣDAP × PON1-108 p-value=0.54). The association between ΣDAP concentrations and SRS scores was not modified by PON1192 (ΣDAP × PON1192 p value=0.89). Urinary DAP concentrations were not associated with children’s social behaviors; these associations were not modified by child PON1 genotype.|
TABLE 5-3 Continued
|Engel et al. (2016) (Analysis of 4 cohort studies)||Pooled data from (1) CHAMACOS; (2) HOME in Ohio; (3) Columbia Center for Childrens’ Environmental Health study of urban population residing in South Bronx or Northern Manhattan in New York City; (4) urban population consisting of women receiving antenatal care at the Mount Sinai Hospital or two private obstetric practices in New York City.||Variable, maternal residential exposures preconception, during pregnancy and postnatal. (1) 2 urine samples at 13 and 26 weeks gestation averaged on a creatinine corrected basis; (2) 2 urine samples at 16 and 26 weeks gestation averaged on a creatinine-corrected basis; (3) 1 urine sample at 32 weeks gestation; (4) 1 urine sample at 31 weeks gestation.||10-fold increase in prenatal exposure to ΣDAPs is associated with an approximate 1-point decrease in the 24-month MDI, whether pooled across four cohorts without adjustment for race/ethnicity (β= –1.28, 95% CI –2.58–0.03) or pooled across race/ethnicities without adjustment for cohort (β= –1.48, 95% CI –2.77– –0.19). Evidence of heterogeneity in associations of ΣDAPs with the MDI according to PON1–108C/T genotype among whites, heterogeneity with ΣDEPs and MDI according to PON1Q192R in the whole population, and with ΣDAPs and PON1Q192R among both whites and blacks. There was no evidence of an association with the psychomotor development index.|
|Silver et al. (2017)||Zhejiang province, China. 237 infants born in Fuyang Maternal and Children’s Hospital between 2008 and 2011.||30 OPs analyzed for in cord blood. Motor function assessed at 6 weeks and 9 months with PDMS.||CPF detected in 36.7% of samples. At 6 weeks, no significant associations between CPF and PDMS outcomes. At 9 months, chlorpyrifos-exposed infants had significantly lower PDMS scores than unexposed infants for reflexes (p=0.04), locomotion (p=0.02), grasping (p=0.05), visuomotor integration (p<0.001), gross motor (p=0.007), fine motor (p=0.002), and total motor (p<0.001) scores.|
|Wang et al. (2017)||Laizhou Wan Birth Cohort study in Shandong, China. 436 mothers recruited from March 2011 to December 2013, 297 12-month-old infants and 262 24-month-old children.||Urine samples collected from mothers at delivery and from 12- and 24-month-old children. Analyzed for 6 DAP metabolites. Gesell Developmental Schedules (motor, adaptive, language, and social domains) used to assess neurodevelopment.||Prenatal OP exposure was negatively associated with 24-month-old children’s development quotients, especially among boys. A 10-fold increase in prenatal diethyl and total DAPs was associated with a 2.59- and 2.49-point decrease in social domain development quotient scores in 24-month-old children (n=262), respectively. Prenatal exposure to OPs was not related to 12-month-old infants’ development quotients.|
|González-Alzaga et al. (2015)||This was an ambispective study. Subjects included 305 (156 boys and 149 girls) children ages 6–11 years old who were randomly selected from public schools in southeastern Spain.
Average age was 8.1 years for boys and 8.4 for girls. 291 (149 boys and 142 girls) children provided a urine sample in the period of low exposure and 270 (137 boys and 133 girls) in the period of high exposure.
|OP exposure assessed by biomonitoring of maternal urinary levels of metabolites of these compounds at two time points of the same crop season, representing low and high pesticide use (January–February and October 2010).
Both prenatal and postnatal residential exposure to pesticides estimated by developing a GIS-technology-based index that integrated distance-weighted measure of agricultural surface, time-series of crop areas per municipality and year, and land-use maps.
Cognitive performance measured with WISC-IV.
|Prenatal exposure index was significantly associated with a poorer score in processing speed subtest in boys (β= –5.7, 95% CI –9.6 – –1.8) but not in girls (β= –1.6, 95% CI –6.4–3.3). There was a nonsignificant inverse association with performance in the remaining subtests of the WISC-IV scale for full-scale IQ, verbal comprehension, and perceptual reasoning, but not for working memory.|
|Raanan et al. (2015)||CHAMACOS cohort; see Eskenazi et al. (2004); examined relationship between early-life exposure to OPs and respiratory outcomes.||See Eskenazi et al. (2004). (Note: Only data on prenatal exposure in first half of pregnancy is given as closest to Gulf War exposures; postnatal exposures are not relevant for this report.)||Higher prenatal DAP concentrations in the first half of pregnancy were nonsignificantly associated with respiratory symptoms in the previous 12 months at 5 or 7 years of age (OR per 10-fold increase=1.11, 95% CI 0.72–1.72) and with exercise-induced coughing (OR=1.24, 95% CI 0.58–2.66).|
|Raanan et al. (2016)||See Raanan et al. (2015).||Spirometry was performed at age 7 years.||Mothers’ DAP urinary concentrations in either half of pregnancy were not significantly associated with lung function measurements (FEV1, FVC, FEV1/FVC ratio and FEF25–75) in their children.|
TABLE 5-3 Continued
|Monge et al. (2007)||N=334 subjects with cases of leukemia ages 0–14 years old. Diagnosed in Costa Rica during 1995–2000. Population controls (N=579).||22 pesticides were identified as high priority on the basis of historical data on use and toxicity data. OPs included malathion and dichlorvos.||Mothers’ exposures to any pesticides during the year before conception and during the first and second trimesters were associated with risk of leukemia (OR=2.4, 95% CI 1.0–5.9; OR=22, 95% CI 2.8–171.5; OR=4.5, 95% CI 1.4–14.7, respectively) and during anytime (OR=2.2, 95% CI 1.0–4.8).
Association with fathers’ exposures to any pesticides during the second trimester (OR=1.5, 95% CI 1.0–2.3). Increased risk with mother’s exposure to OPs during the first trimester (OR=3.5, 95% CI 1.0–12.2) and for fathers during the year before conception and the first trimester (OR=1.5, 95% CI 1.0–2.2 and OR=0.6, 95% CI 1.0–2.6, respectively).
|Other Developmental Effects|
|Carmichael et al. (2014)||569 cases of children born 1997–2006 with 1 of 7 congential heart defects in 1 of 8 counties in San Joaquin Valley, CA, and 785 infants born without birth defects. Birth defects (heterotaxia; tetralogy of Fallot; hypoplastic left heart syndrome; coarctation of the aorta; pulmonary valve stenosis; ventricular spetal defect, perimembranous; and atrial septal defect, secundum) based on California Birth Defects Monitoring Program.||Pesticide exposure was estimated for each mother on the basis of her residence from 3 months prior to conception to delivery using geospatial coding for eight San Joaquin counties and the Pesticide Use Reporting records for pesticides applied at greater than 100 lbs in any of the eight counties.||Prenatal exposure to CPF was significantly associated with pulmonary valve stenosis (OR=2.4, 95% CI 1.0–6.3) and with atrial septal defects, secundum (OR=1.9, 95% CI 1.1–3.4). Prenatal exposure to OPs as a group was associated with an increase in pulmonary valve stenosis as well (OR=2.5, 95% CI 1.2–5.0).|
|Carmichael et al. (2016)||367 cases of children born with 1 of 5 birth defects. See Carmichael et al. (2014).
Birth defects (anotia/microtia; anorectal atresia/stenosis; transverse limb deficiency; craniosynostosis; and diaphragmatic hernia) based on California Birth Defects Monitoring Program.
|See Carmichael et al. (2014).||Prenatal exposure to OPs as a group (based on mechanism of action) was not significantly associated with anorectal atresia/stenosis (OR=0.4, 95% CI 0.2–0.9), but was significantly associated with craniosynostosis (OR=1.9, 95% CI 1.1–3.2); CPF specifically was significantly associated with anotia (OR=2.0, 95% CI 1.1–3.8) and craniosynostosis (OR=2.4, 95% CI 1.2–4.6). The ORs for all other OPs or birth defects were ≥0.5 or ≤2.0 or the CI included 1.0.|
|Buscail et al. (2015)||PELAGIE cohort of subsample of 248 pregnant women from Brittany, France, from January 2002 to February 2006, recuited before the 19th week of gestation; women provided urine sample, cord blood sample, and umbilical cord sample.||Urine analyzed for 6 DAPs; maternal exposure during pregnancy reported via questionnaire.||No association between any outcome and residential proximity to crops. Positive association between the highest tertile of DMPs and recurrent otitis media occurrence was found in the complete-case analysis (OR=2.90, 95% CI 1.12–7.48, p-value for trend=0.02), but it was no longer statistically significant after multiple imputations (OR=1.76, 95% CI 0.82–3.77, p-value for trend=0.13). No statistically significant association was found between urinary concentrations of other OP insecticide metabolites and either otitis media or recurrent otitis media.|
NOTE: ADHD=attention deficit–hyperactivity disorder; ASD=autism spectrum disorder; BASC=Behavior Assessment System for Children; BNBAS=Brazelton Neonatal Behavioral Assessment Scale; BRIEF=Behavior Rating Inventory of Executive Functioning; BSID-II=Bayley Scales of Infant Development II; CBCL= Child Behavior Checklist; CCCEH=Columbia Center for Children’s Environmental Health study; CHAMACOS=Center for the Health Assessment of Mothers and Children of Salinas study; CI=confidence interval; CPF=chlorpyrifos; CpG=cytosine-adjacent-to-guanine; DAP=dialkyl phosphate; DE=diethyl alkyl metabolites; DEDTP=diethyldithiophosphate; DEP=diethylphosphate; DETP=diethylthiophosphate; DM=dimethyl alkyl metabolites; DMDTP=dimethyldithiophosphate; DMP=dimethylphosphate; DMTP=dimethylthiophosphate; ETS=environmental tobacco smoke; FEF25–75=forced expiratory flow at 25–75%; FEV1=forced expiratory volume in 1 second; FVC=forced vital capacity; GIS=geospatial information system; HOME=Health Outcomes and Measures of the Environment study; IQ=intelligence quotient; K-CPT=Conners’ Kiddie Continuous Performance Test; LOD=level of detection; MDA=malathion dicarboxylic acid; MDI=mental development index; MRI=magnetic resonance imaging; Mt. Sinai=Mount Sinai Children’s Environmental Health Cohort; NEPSY-II=Developmental NEuroPSYchological Assessment; OP=organophosphate; OR=odds ratio; PAH=polyaromatic hydrocarbons; PDI=psychomotor development index; PDMS=Peabody Developmental Motor Scales; PON1=paraoxonase; SRS=Social Responsiveness Scale; WISC-IV=Wechsler Intelligence Scale for Children, 4th edition; WPPSI-III=Wechsler Preschool and Primary Scales of Intelligence-III.
Three carbamate pesticides—propoxur, carbaryl, and methomyl—are among the agents listed in Gulf War legislation P.L. 105-277 and P.L. 105-3683 and therefore were considered by the Volume 2 and Volume 11 committees. Carbamates have a similar mechanism of toxicity to that of OPs—that is, they both inhibit acetylcholinesterases. Not all carbamates are pesticides; some, such as pyridostigmine bromide (discussed as a distinct deployment-related exposure in Chapter 4), have medicinal uses. Unlike OPs, the effects of carbamate poisoning tend to be of shorter duration because their inhibition of nervous tissue acetylcholinesterase is reversible, and carbamates are more rapidly metabolized (Roberts and Reigart, 2013).
Indoor residential uses of propoxur were cancelled in 2007, and as of 2016, it is no longer used in pet flea collars. However, it continues to be sold for commercial and residential control of cockroaches and other insect infestations. Methomyl is used as fly bait and in agricultural applications, but it has no residential uses. Carbaryl has both agricultural and residential uses and is widely sold under the brand name Sevin®.
As with the OP pesticides, metabolites of carbamate pesticides detected in the urine and blood can be used to indicate exposure. Urinary metabolites of carbaryl include 1- and 2-naphthol, and 2-isopro-poxyphenol is a propoxur metabolite that has been detected in human blood. The elimination half-life in humans of 1-naphthol in urine is about 3.6 hours for an oral carbaryl dose of 0.008 mg/kg (Sams, 2017).
There are no ATSDR toxicological profiles available for any of the carbamates. The California Office of Environmental Health Hazard Assessment lists carbaryl as a male and female toxicant as well as a developmental toxicant (OEHHA, 2018).
Table 5-4 at the end of this section summarizes the studies used by the Volume 11 committee to draw its conclusions.
Reproductive Effects in Men and Women
The Volume 2 committee found little relevant information on the reproductive and developmental effects of human or animal exposure to propoxur and methomyl. The committee did review several epidemiologic studies that assessed reproductive effects associated with exposure to either carbamates in general or to carbaryl specifically. The Volume 11 committee found a similar dearth of epidemiologic information on propoxur and methomyl and only one set of studies on carbaryl published since Volume 2.
In Volume 2, two studies on sperm effects in the same cohort of male carbaryl production workers were assessed (Whorton et al., 1979; Wyrobek et al., 1981); they reported that compared with unexposed controls there was an exposure-dependent increase in the number of oligospermic men among the male carbaryl production workers (p=0.07). There was also an increase in teratospermic men (having more than 60% abnormal sperm), but the increase was not dose-dependent. The latter study also found that men who were younger or who had more years of exposure had a greater proportion of abnormal sperm. The Volume 2 committee concluded that both of these cross-sectional studies had flaws in their methodology, including the choice of chemical workers as controls and crude measures of exposure.
The Volume 11 committee identified two new studies that examined reproductive effects in men (see Table 5-4). Meeker and colleagues (2004a,b, 2006, 2008a) completed a series of reports on the effects of low-level environmental exposure on human sperm and male reproductive hormones. Men (n=330) who were recruited from an infertility clinic provided semen and urine samples; the latter were analyzed for
3 Persian Gulf War Veterans Act of 1998—P.L. 105-277; Veterans Program Enhancement Act of 1998—P.L. 105-368.
1-naphthol, a urinary metabolite of carbaryl. None of the subjects reported being occupationally exposed to carbaryl or other pesticides. Compared with the lowest-exposure tertile, ORs for the medium and high tertiles of 1-naphthol and below-reference sperm concentration were 4.2 (95% CI 1.4–13.0), and 4.2 (95% CI 1.4–12.6; p-value for trend=0.01); for 1-naphthol exposures and sperm motility, the ORs were 2.5 (95% CI 1.3–4.7) and 2.4 (95% CI 1.2–4.5; p-value for trend=0.01). There was no association between 1-naphthol and abnormal sperm morphology. However, increasing carbaryl exposure was associated with increasing DNA damage as determined by the neutral comet assay. Reduced testosterone and estradiol levels were associated with carbaryl exposures but were not dose-dependent.
In a study of 16 production workers in China who had occupational exposure to carbaryl for at least 1 year, no significant differences in semen volume, sperm concentration, sperm number per ejaculum, or sperm motility were seen between the carbaryl-exposed workers and members of two control groups (12 unexposed workers in the same plant and 18 workers outside the plant) (Xia et al., 2005). However, morphological defects such as sperm abnormalities were significantly higher in the exposed workers than in the external control group (p=0.008), as were the frequencies of disomic sex-chromosome-bearing (highest the XY-bearing) and disomic chromosome 18 spermatozoa (p<0.05 and/or p<0.01, respectively).
The Volume 2 committee also considered several studies on fertility and carbamate exposure. Several researchers reported on the large Ontario Farm Family Health Study, which assessed both male and female exposures to pesticides during the month of attempted conception or at any time during the previous 2 months. Neither maternal-only nor paternal-only exposure to carbaryl was associated with significant effects on TTP; the close correlation of exposures between members of a couple made it difficult to determine whether any effects were specifically maternally or paternally mediated (Curtis et al., 1999).
Adverse Pregnancy Outcomes
In the Ontario Farm Family Health Study reviewed by the Volume 2 committee, both male and female exposures to pesticides were assessed during the month of attempted conception or at any time during the previous 2 months (Curtis et al., 1999). For men who reported working with carbaryl, there was a significant increase in spontaneous abortions, although there was no association with altered sex ratios, small for gestational-age deliveries, or preterm births (Savitz et al., 1997). A further analysis of preconception exposure for the couple as a unit showed no increased risk for overall, early, or late spontaneous abortions (Arbuckle et al., 2001).
Bell et al. (2001) studied the relationship between fetal death and maternal residential proximity to agricultural areas in California with pesticide applications. They found that exposure to carbamates from the third to eighth week of pregnancy was not significantly associated with an increased risk of fetal death due to congenital abnormalities. Wives of pesticide applicators were at increased risk of spontaneous abortion, although the Volume 2 committee noted that the study had few details on its methodology.
The Volume 11 committee reviewed one study (see Table 5-4) and one EPA summary on the effects of exposure to carbamate pesticides on fetal growth and birth outcomes.
Whyatt et al. (2004, 2005) assessed birth outcomes in women enrolled in the CCCEH study between 1998 and 2002. All women had detectable levels of chlorpyrifos, diazinon, and propoxur in personal air samples collected at their homes over a 48-hour period during their third trimester. The insecticides were detected in 45–74% of blood samples collected from the mothers and newborns at delivery. Among the birth outcomes measured in newborns (birth weight, birth length, and head circumference), for those born before 2001 only the association between 2-isopropoxyphenol (a propoxur metabolite) and decreased birth length was statistically significant (β= –0.73 cm/unit, p=0.01); for those born after January 1, 2001, the association remained inverse, but the magnitude of the effect was less and no longer significant (β= –0.30 cm/unit, p=0.56).
Volume 2 reported on two animal studies on the effects of carbaryl on sperm. When 50 mg or 100 mg carbaryl/kg was fed to rats 5 days per week for 60 or 90 days, dose- and age-dependent decreases in sperm count and motility and increased abnormal sperm structure were observed, particularly in young animals (Pant et al., 1995, 1996). In a sperm-abnormality assay, rats fed carbaryl for 3, 6, 9, 12, 15, and 18 months at 12.5, 25, and 250 mg/kg per day had genotoxic effects (Luca and Balan, 1987).
The Volume 2 committee further reported that the reproductive toxicity of carbaryl has been examined in rats, gerbils, dogs, and primates following a variety of protocols, including multigeneration studies and studies to determine whether carbamate pesticides act as endocrine modulators. The oral administration of carbaryl to rats for 1 year resulted in a decreased sperm count, decreased motility, and abnormalities (Kitagawa et al., 1977). In a three-generation reproductive study in rats, even the lowest dose (100 mg/kg per day) decreased weaning weights (Collins et al., 1971). In a three-generation study of gerbils, a dose of 150 mg/kg per day was found to be the no-observed-effect level; higher doses were associated with decreases in fertility, litter size, viability, and survival (Collins et al., 1971). In two dog studies, teratogenesis was seen at doses similar to those that did not result in fetotoxicity (Imming et al., 1969; Smalley et al., 1968).
The EPA Integrated Risk Information System (IRIS) Chemical Assessment Summary (EPA, 1987) reported on one dog study in which doses of methomyl up to 1,000 ppm had no effect on reproductive organs. A 2016 EPA fact sheet for propoxur reported that no information was available on the reproductive or developmental effects of propoxur in humans and that no adverse reproductive or developmental effects had been observed in an oral study of rabbits; however, a few studies of rats orally exposed to propoxur found fetotoxic effects, decreased numbers of pups, and depressed fetal weight (EPA, 2000). No studies on adverse effects on maternal reproduction or fetal outcomes were identified.
The Volume 11 committee identified no additional studies of carbaryl exposure in animals, but it did identify one on propoxur and one on methomyl. Male rats exposed to propoxur at a dose of 5.2 mg/kg body weight had significant (p<0.05) decreases in body weight gain, sperm density, serum and intratesticular total cholesterol concentration, and intratesticular total protein. However, there were no significant effects on gestation, fertility and parturition indices, average birth weight, litter size, and pups’ sex ratio for untreated females mated with treated males (Ngoula et al., 2007).
Male rats that were administered methomyl orally (17 mg/kg in saline) daily for 2 months had a significant decrease in their levels of testosterone, while their levels of FSH, LH, and prolactin were significantly increased. The testes had degenerative changes in the seminiferous tubules up to total cellular destruction (Mahgoub and El-Medany, 2001).
The 1992 EPA IRIS Chemical Assessment for Baygon (a brand name for propoxur) briefly summarized three animal studies conducted by the chemical industry: a three-generation reproduction study in rats that found decreased pup numbers at 37.5 mg/kg/day (the generation in which the effects were seen was not stated); a no-observed-effect level of 500 mg/kg/day in a teratogenicity study in rabbits; and a fetotoxicity study in rats that showed a significantly lower average fetal weight, at 150 mg/kg/day (EPA, 1992).
Volume 2 identified several studies that reported an increased risk of childhood leukemia and brain cancers in children born to mothers who reported exposure to pesticides, including carbamates, during pregnancy. However, none of the studies examined preconception exposure, and the exposure to carbamates specifically was based on the researchers’ assessment of likely uses around the house; none of the studies actually confirmed the use of carbamate pesticides.
The Volume 11 committee identified several studies of developmental effects and carbamate exposure (see Table 5-4). One case-control study found no significant association between the prevalence of hypospadias in children born to parents living in agricultural areas where carbaryl was used and in children from areas where carbaryl was not used (OR=0.80, 95% CI 0.20–3.18) (Meyer et al., 2006). Exposure was based on agricultural pesticide application records and geographic information system modeling.
In a second case-control study of the association between exposure to agricultural pesticides in the San Joaquin Valley in California and the presence of any of five birth defects, Carmichael et al. (2016) found that prenatal exposure to methomyl was significantly associated only with craniosynostosis (OR=3.0, 95% CI 1.1–7.8).
Lafiura et al. (2007) analyzed umbilical cord blood samples obtained from 49 infants, 39 of whom had prenatal exposure to propoxur, as determined by meconium analysis and maternal blood samples. The mothers lived in an agricultural area of the Philippines. Cord blood was analyzed for the presence of t(8;21) (q22;q22), a common cytogenetic abnormality in childhood acute myeloid leukemia (AML). Propoxur-exposed infants had a nonsignificant twofold increase in the incidence of t(8;21) compared with unexposed infants (20.5% versus 10%). A follow-up analysis of this cohort found that at 24 months of age, prenatal exposure to propoxur was negatively related to motor development (β= –0.14, p<0.001) but was unrelated to social and performance development in the children (Ostrea et al., 2012). The committee notes that this was a small number of infants, which made it difficult to extrapolate the results to the general population.
In a further follow-up of the CHAMACOS study in California described in the section on OPs, Gunier et al. (2017) analyzed the Pesticide Use Reporting data from California to determine the use of OPs and carbamate pesticides in fields near the maternal residence during pregnancy. Greater application of carbamates was associated with a significant decrease at age 7 in children’s WISC verbal comprehension scores (β= –2.4, 95% CI –3.9– –1.0) but not in scores on the other scales. Greater application of OPs and carbamates combined also was associated with significantly lower WISC full-scale IQ (β= –2.1, 95% CI –3.7– –0.4) and verbal comprehension scores (β= –2.9, 95% CI –4.5– –1.4) at age 7; however, by age 10 there were no significant differences for any WISC score and maternal residential exposure to OPs and carbamates combined (Rowe et al., 2016).
Monge et al. (2007), discussed in the OP section, also examined the effects of maternal (n=876) and paternal (n=762) exposures to pesticides on the risk of childhood leukemia (n=334) in Costa Rica in 1995–2000. Parents were interviewed about risk factors for childhood leukemia and their exposures were rated by hazard value and time of occurrence. The fathers’ exposure to carbamates in the year prior to conception or at any time during the mother’s pregnancy was not significantly associated with an increased risk of total leukemia in their children. Specific information on the mother’s exposures to carbamates and any increased risk of leukemia in their children was not reported.
No animal data on developmental effects from exposure to carbamate pesticides was reported in Volume 2, nor did the Volume 11 committee identify any new animal studies on the relevant carbamates.
Synthesis and Conclusions
Information on the reproductive and developmental toxicity of the carbamate pesticides is limited almost exclusively to carbaryl, with little information on propoxur or methomyl. With the exception of
occupational studies in carbaryl workers—the CCCEH study and the studies by Meeker et al.—virtually all other epidemiologic studies on male or female reproductive effects relied on self-reports of exposure to carbamates, often in conjunction with other potential pesticide exposures. This poses considerable challenges for the interpretation of the findings. The Volume 2 committee found adverse effects on sperm in the two occupational studies it reviewed, but it concluded that the studies were flawed. The Volume 11 committee identified one new occupational study in China that found an increase in sperm abnormalities (Xia et al., 2005) and a second study of men with no occupational exposure to carbaryl who attended an infertility clinic in the United States (Meeker et al., 2004a,b). The latter study also found sperm damage and decreased sperm concentrations to be associated with carbaryl exposure. Animal studies with carbaryl, propoxur, or methomyl supported the adverse effects on sperm seen in humans; other studies had inconsistent results for other reproductive effects in male animals, and there were no studies on female reproduction. Additional animal evidence indicates that carbamates, specifically carbaryl, act as an endocrine-disrupting compound, providing biological plausibility and mechanistic support for the reproductive effects in humans. There were no relevant animal studies on female reproductive effects.
Studies on adverse pregnancy outcomes associated with carbamate exposure reviewed by the Volume 2 committee showed that paternal occupational exposure to carbaryl increased the risk of spontaneous abortion, but when couples were analyzed as a unit, there was no increased risk (Curtis et al., 1999). Maternal exposure had no significant effects on birth outcomes. The Volume 11 committee identified only one new study that examined birth outcomes for women with exposure to chlorpyrifos, diazinon, and propoxur during pregnancy. Only propoxur was significantly associated with birth length, and it was not associated with birth length or head circumference at birth. The animal data provide inconsistent evidence of an association between exposure to carbamates and adverse reproductive effects in male and female animals. Adverse pregnancy outcomes were reported in two multigeneration studies of carbaryl.
The Volume 11 committee concludes that there is sufficient evidence of an association between exposure to carbamates 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 carbamates and reproductive effects in women, or with adverse pregnancy outcomes.
The epidemiologic studies on developmental effects in children exposed prenatally to carbamates, particularly the prevalence of childhood cancers, are of concern. While the CHAMACOS study (Gunier et al., 2017) indicated that there are some early neurocognitive deficits associated with maternal residential proximity to carbamate application, these deficits resolved with age (Rowe et al., 2016). The small study by Lafiura (2007) also showed neurodevelopmental effects with maternal carbamate exposure. Other studies were not restricted to carbamates alone. There is a lack of animal data on developmental effects in offspring following parental exposure to carbamates reported in Volume 2, and the Volume 11 committee did not identify any new animal studies.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between prenatal exposure to carbamates and developmental effects.
TABLE 5-4 Summary of Reproductive and Developmental Effects of Carbamate Pesticides
|Reproductive Effects in Men|
|Meeker et al. (2004a)||330 men who were partners in subfertile couples seeking infertility diagnosis at Massachusetts General Hospital (Boston) between January 2000 and April 2003.||1-N, a metabolite of carbaryl, was detected in 99.7% of men’s urine.||Compared with men in the lowest 1-N concentration tertile, men in both the medium and high 1-N tertiles were more likely to have below-reference sperm concentration (ORs low exposure=1.0; medium exposure=4.2, 95% CI 1.4–13.0; high exposure=4.2, 95% CI 1.4–12.6) and sperm motility (low=1.0, medium=2.5, 95% CI 1.3–4.7; high=2.4, 95% CI 1.2–4.5). Sperm morphology was not significantly associated with 1-N.|
|Meeker et al. (2004b)||See Meeker et al. (2004a); semen samples available for 260 men; 46 samples excluded with specific gravity >1.03 or <1.01.||See Meeker et al. (2004a).||For an IQR increase in 1-N, Tail% significantly increased by 4.13% (95% CI 1.92–6.32). An IQR increase in 1-N resulted in a nonsignificant reduction in TDM (–2.18 mm, 95% CI –4.88–0.50) and a nonsignificant decrease in comet extent (–0.92 mm, 95% CI –7.39–5.55).|
|Meeker et al. (2006)||See Meeker et al. (2004a); 330 men with semen, urine, and nonfasting blood samples taken on same day.||See Meeker et al. (2004a).||Regression model results show an inverse association between 1-N and testosterone. An IQR range increase in 1-N was associated with a 6.0% (–11% to –1%) decrease in median testosterone level.|
|Meeker et al. (2008a)||See Meeker et al. (2004a).||See Meeker et al. (2004a).||No significant association between increased 1-N levels and decreased estradiol (p=0.09). Highest four quintiles of 1-N were associated with declined estradiol but not in a dose-related manner. No statistically significant associations between 1-N and prolactin.|
TABLE 5-4 Continued
|Xia et al. (2005)||46 sperm donors ages 21–48 years in Changzhou, China; 16 were carbaryl-exposed workers; 12 internal controls were clerical or official workers in the same pesticide factory but far away from the pesticide workshop; 18 external control group were selected from professions other than pesticide workers.||Exposed workers worked in the plant for >1 year and had been working continuously for 6 months before biological sampling.||No significant differences in semen volume, sperm concentration, sperm number per ejaculum, and sperm motility between carbaryl-exposed and two control groups. Carbaryl group had significantly more morphological defects (e.g., sperm abnormalities) than external control group (p=0.008), significantly higher mean percentage of spermatozoa with fragmented DNA than internal (p=0.035) or external controls (p=0.030), and significant differences in the frequencies of disomic sex-chromosome-bearing (highest the XY-bearing) and disomic chromosome 18 spermatozoa (p<0.05 and p<0.01, respectively). Nullisomies of sex chromosomes and chromosome 18 were significantly higher than those in the external (p<0.01) but not internal controls. No significant differences in percentage of normal spermatozoa with X- or Y-chromosome or diploidy rates between any two groups.|
|Adverse Pregnancy Outcomes|
|Whyatt et al. (2004)||CCCEH cohort; 314 African American and Dominican women, 18–35 years old, enrolled into the cohort between 1998 and 2002. Umbilical cord blood collected at birth; and maternal blood collected within 2 days postpartum.||85% reported using some form of pest control measures during pregnancy, and 35% reported using an exterminator; 56.5% of users reported using one or more of the more highly toxic pest control methods (can sprays, pest bombs, or sprays by exterminator). Measured 2-isopropoxyphenol, a metabolite of propoxur.||2-isopropoxyphenol in cord plasma was inversely associated with birth length (p=0.05). Among newborns born before 2001, the association between (ln)2-isopropoxyphenol and birth length was statistically significant (β= –0.73 cm/unit, p=0.01). Among newborns born after January 1, 2001, the association remained inverse, but the magnitude of effect was less and not significant (β= –0.30 cm/unit, p=0.56). No association was seen between infant head circumference and levels of propoxur or its metabolite in maternal personal air and cord blood samples.|
|Whyatt et al. (2005)||CCCEH cohort; see Whyatt et al. (2004).||See Whyatt et al. (2004)||No significant associations between 2-isopropoxyphenol levels in cord plasma and birth weight for infants born before or after 2001 (p=0.12 and p=0.27, respectively). For infants born before 2001, statistically significant inverse relationship between 2-isopropoxyphenol in cord blood and birth length (β= –0.73, p=0.01), but not for infants born after 2001 (p=0.56).|
|Meyer et al. (2006)||354 cases of hypospadias in male children born between 1998 and 2002 in eastern Arkansas; 727 controls were selected from birth certificates.||Classified exposure based on pounds of pesticides (estimated by crop type) applied or persisting within 500 m of each child’s home during gestational weeks 6–16.||Hypospadias were not significantly associated with increases in carbaryl application (OR=0.80, 95% CI 0.20–3.18).|
|Carmichael et al. (2016)||367 cases of children born 1997–2006 with 1 of 5 birth defects in any one of 8 counties in San Joaquin Valley, CA, and 785 infants born without birth defects. Birth defects (anotia/microtia; anorectal atresia/stenosis; transverse limb deficiency; craniosynostosis; and diaphragmatic hernia) based on California Birth Defects Monitoring Program.||Pesticide exposure within 500-m radius was estimated for each mother’s residence from 1 month before to 2 months after her reported date of conception using geospatial coding for San Joaquin Valley counties and the Pesticide Use Reporting records for pesticides applied at >100 lbs in any of the counties.||Prenatal exposure to methomyl was significantly associated with craniosynostosis (OR=3.0, 95% CI 1.1–7.8) but not with other birth defects.|
|Lafiura et al. (2007)||Pregnant women prospectively enrolled into the study at mid-gestation, living in Bulacan, an agroindustrial province of the Philippines.||49 meconium and umbilical cord blood samples were obtained of which 39 were exposed prenatally to propoxur and 10 were unexposed.||Incidence of t(8;21)(q22;q22), a common cytogenetic abnormality seen in AML, was twofold higher in the exposed children than in the unexposed children. Levels of AML1-ETO fusion transcripts resulting from t(8;21) abnormality were positively associated with propoxur concentrations in meconium.|
TABLE 5-4 Continued
|Ostrea et al. (2012)||Same population as Lafiura et al. (2007). 696 mother–infant pairs assessed.||Cyfluthrin/propoxur major pesticide used in 73% of homes/farms. Neurodevelopment at 24 months assessed with Griffith Mental Development Scales: (1) locomotor subscale (gross motor development, including balance and coordination of movements), (2) social development (personal–social [daily activities and child interaction] and language subscales [receptive and expressive language]), and (3) performance subscale (visuospatial skills and reaction time).||Exposure to propoxur negatively associated with motor development (β= –0.14, p<0.001), but not with social and performance development.|
|Gunier et al. (2017)||CHAMACOS cohort of women living in agricultural area of Salinas Valley, CA; children born October 1999–October 2000; eligible n=488) women were ≥18 years of age, <20 weeks gestation at enrollment.||Gestational and preconception exposure, half of the women resided in United States for <5 years; approximately 28% had worked in the fields during pregnancy, another 14% worked at other agriculture jobs (e.g., packing shed, nursery, and greenhouse work). 85% of the women had agricultural workers living in their homes during their pregnancy. Exposure to OPs was measured as DAP metabolites in maternal urine during pregnancy (mean=13 and 26 weeks gestation) and ChE in whole blood and BuChE in plasma collected||Increased exposure to carbamates during pregnancy was significantly associated with an estimated 2.4-point decrease in verbal comprehension (95% CI –3.9– –1.0), but not with the other WISC subscales of full-scale IQ, working memory, processing speed, and perceptual reasoning.|
|from mothers during pregnancy and at delivery, and from the umbilical cord. Children assessed for neurodevelopment at 7 years using WISC-IV. Pesticide use within 1 km of mother’s residence during pregnancy estimated using California Pesticide Use Reporting data from 1999–2001.|
|Rowe et al. (2016)||CHAMACOS cohort, see Grunier et al. (2007). 501 10-year-old children.||Exposure based on kgs of OPs and carbamates used near the residence where each woman lived the longest during pregnancy. WISC-IV used to test neurodevelopment at 10 years of age.||WISC scores were significantly decreased only at the fourth quintile of combined OP and carbamate use (highest usage) during pregnancy (full-scale IQ β= –3.0, 95% CI –5.6– –0.3; verbal comprehension β= –1.3, 95% CI –4.1–1.4; perceptual reasoning β= –4.0, 95% CI –7.6– –0.4; working memory β= –2.8, 95% CI –5.6––0.1; and processing speed β= –0.7, 95% CI –3.8–2.4).|
|Monge et al. (2007)||334 cases of leukemia and 579 population controls. Children ages 0–14 years diagnosed in Costa Rica in 1995–2000.||Childhood leukemia identified from Cancer Registry and the Children’s Hospital of Costa Rica. Parents interviewed in 2001–2003 about occupational, environmental, and home pesticide exposures of both parents from 12 months prior to conception to diagnosis of cancer.||No increased risk of childhood leukemia associated with paternal exposure to carbamates at any time prior to birth (ORs=1.3, 95% CI 0.6–2.5) or for year before conception (OR=1.2, 95% CI 0.4–3.3 for first trimester; OR=1.6, 95% CI 0.6–4.1 for second trimester; and OR=2.0, 95% CI 0.7–5.8 for third trimester).|
NOTE: 1-N=1-naphthol; AML=acute myloid leukemia; BuChE=butyryl cholinesterase; CCCEH=Columbia Center for Children’s Environmental Health study; CHAMACOS=Center for the Health Assessment of Mothers and Children of Salinas study; ChE=cholinesterase; CI=confidence interval; DAP=dialkyl phosphate; DNA=deoxyribonucleic acid; IQR=interquartile range; OP=organophosphate; OR=odds ratio; TDM=tail distributed moment; WISC-IV=Wechsler Intelligence Scale for Children, 4th edition.
Pyrethrins are naturally occurring pesticides found in chrysanthemum flowers. They degrade rapidly in the environment, particularly in the presence of sunlight. Pyrethroids are commercially manufactured pesticides with a chemical structure similar to pyrethrins, which have been in use for more than 40 years. They are more toxic to insects and mammals than pyrethrins, and they persist longer in the environment (ATSDR, 2003b). Pyrethroids were used extensively during the 1990–1991 Gulf War, with estimates that each service member deployed to the Persian Gulf received 2.2 cans of permethrin (as 6-oz aerosol cans) to treat their uniforms (PAC, 1996). Service members in the Post-9/11 conflicts also used permethrin-treated uniforms and bed nets to help prevent sand fly bites.
The acute neurotoxicity of pyrethroids to human adults is well established (Shafer et al., 2005). Permethrin is metabolized by the cytochrome P450 enzyme and esterases to at least 80 metabolites (Casida et al., 1983), but most of them are inactive. Half-lives vary among the different pyrethroid compounds, but they range from 6.4 to 16.5 hours in humans, with elimination mostly completed within 5 days (ATSDR, 2003b). Among the pyrethroid metabolites found in urine and blood and discussed in this section are PBA, cis- and trans-3-(2,2-dichlorovinyl)-2,2-dimethylcyclopropane carboxylic acid (CDCCA and TDCCA, respectively), and cis-2,2-dibromovinyl-2,2-dimethylcyclopropane-carboxylic acid (DBCA).
Based on the Defense Logistics Agency list of pesticides sent to the Gulf War (IOM, 2003) and the reported use of pyrethroids in Iraq in 2003 (Schleier et al., 2009), the Volume 11 committee paid particular attention to the pyrethroids permethrin, cypermethrin, deltamethrin, and resmethrin.
The Volume 2 committee did not consider specific reproductive effects that may result from the exposure of men or women to pyrethroid insecticides. The ATSDR publication Toxicological Profile on Pyrethrins and Pyrethroids (2003b) reported that standard reproductive toxicity studies, including some that were performed for three successive generations, did not indicate that pyrethrins or pyrethroids are of particular concern to reproductive success.
Reproductive Effects in Men and Women
The Volume 11 committee considered several papers that explored the association between environmental exposure of men to pyrethroids and the impacts on their reproductive systems, particularly sperm quality (see Table 5-5). The majority of the studies assessed exposure to pyrethroids by analyzing for urinary metabolites, including PBA, CDCCA, TDCCA, and DBCA. In general, the men in the studies were recruited from infertility clinics.
Xia et al. (2008) reported a dose–response relationship between increasing urinary PBA quartiles and decreased sperm concentration in 376 Chinese men. ORs for the first, second, third, and fourth concentration quartiles were 1.00, 1.31 (95% CI 0.65–2.64), 1.73 (95% CI 0.87–3.45), and 2.04 (95% CI 1.02–4.09), respectively (p-value for trend=0.027), whereas sperm volume, sperm number per ejaculum, and sperm motility showed no significant associations with PBA quartiles. In another study of Chinese men, Ji et al. (2011) found a significant inverse correlation between urinary PBA levels and sperm concentration (β= –0.27, 95% CI –0.41– –0.12, p<0.001) and a significant positive correlation between PBA levels and sperm DNA fragmentation (β=0.27, 95% CI 0.15–0.39, p<0.001).
Imai et al. (2014) found in a sample of 323 Japanese men (university students) that environmental exposure to pyrethroid insecticides as determined by urinary PBA did not affect semen volume, sperm concentration, motility, the total number of sperm, or the total number of motile sperm. Furthermore,
the researchers did not find any association between PBA and the levels of FSH, LH, testosterone, sex hormone-binding globulin, inhibin B, or calculated free testosterone. More than 91% of the men had detectable levels of PBA in their urine (Yoshinaga et al., 2014).
In a study of 195 men in Poland conducted to evaluate sperm DNA damage (Jurewicz et al., 2015) and aneuploidy (Radwan et al., 2015), there was a positive association between CDCCA greater than the 50th percentage concentration distribution in urine and the percentage of medium DFI, the percentage of immature sperm, and disomy of chromosome 18 (p=0.04, p=0.04, and p=0.05, respectively). A level of PBA was positively related to the percentage of high DFI (p=0.03). The TDCCA, DBCA levels, and the sum of pyrethroid metabolites were not associated with any sperm DNA damage measures. PBA concentrations in the urine that were both below and above the 50th percentile were associated with disomy of sex chromosomes, although to different degrees—XY disomy (p=0.05 and p=0.02, respectively), Y disomy (p=0.04 and 0.02, respectively), disomy of chromosome 21 (p=0.04 and p=0.04, respectively), and total disomy (p=0.03 and p=0.04, respectively); while PBA levels that were above the 50th percentile were associated with disomy of chromosome 18 (p=0.03). Urinary levels of TDCCA above the 50th percentile were related to XY disomy (p=0.04) and disomy of chromosome 21 (p=0.05) but not with any sperm DNA damage. The levels of DBCA and the sum of pyrethroid metabolites also were not associated with any sperm DNA damage measures. The three metabolites were also significantly associated with an increase in the number of sperm having abnormal morphology and a decrease in sperm concentration and levels of testosterone (Radwan et al., 2014).
One study of U.S. men also found environmental exposure to pyrethroids to be associated with semen quality. In a study of 207 men attending an infertility clinic in Massachusetts, Meeker et al. (2008b) found that increasing levels of urinary PBA and CDCCA were associated with increased sperm DNA damage. TDCCA levels showed a dose-dependent increase in the risk of having sperm concentration, motility and morphology below the reference values. Further assessment of these men showed that all three metabolites were associated with an increased incidence of disomy of chromosomes X, Y, and 18. Specifically, disomy was increased in YY18 for PBA, CDCCA, and TDCCA (incidence ratio rates [IRRs]=1.28, 95% CI 1.15–1.42; 1.18, 95% CI 1.05–1.31; and 1.19, 95% CI 1.06–1.34, respectively). Both CDCCA and TDCCA were significantly associated with increased disomy for XY18 and total disomy, but PBA was not. None of the metabolites were associated with disomy for XX18 or 1818 (Young et al., 2013).
Meeker et al. (2009) assessed the effects of pyrethroids on serum hormone levels in men and found that PBA, TDCCA, and CDCCA were all positively associated with FSH levels (all p-values for trend <0.05). CDCCA was also significantly associated with luteinizing hormone; however, CDCCA and TDCCA were inversely associated with inhibin B (p-values for trend=0.03 and 0.02, respectively).
In a systematic review of three papers—Ji et al. (2011); Meeker et al. (2008b); Xia et al. (2008)—published between 2007 and 2012 on the effects of exposure to pyrethroid insecticides on sperm parameters, Martenies and Perry (2013) concluded that although all three studies reported inverse associations between pyrethroid metabolites in urine and decreased sperm concentrations, the results were suggestive but not definitive.
The Volume 11 committee identified only one study on the effects of pyrethroids on female reproduction. Hu et al. (2018) conducted a TTP study of 615 women in Shanghai, China, who were recruited from two preconception care clinics between August 2013 and April 2015. Women were enrolled in the study before conceiving a child and were followed for approximately 1 year after they had stopped using a contraceptive. Exposures to pyrethroids were determined by measuring PBA, TDCCA, and CDCCA in urine. After adjustments were made for age, BMI before pregnancy, smoking, and education, women who had never been pregnant and were in the highest quartile for PBA concentrations had significantly
longer TTP (fecundability OR=0.72, 95% CI 0.53–0.98) and increased infertility (OR=2.03, 95% CI 1.10–3.74) than women in the lowest quartile.
Adverse Pregnancy Outcomes
The ATSDR review Toxicological Profile for Pyrethrins and Pyrethroids (2003b) reported that no studies had been identified on the reproductive effects, including birth outcomes, of pyrethroid pesticides following inhalation, oral, or dermal exposures in humans. Several studies in Poland, China, and Japan have been published since the ATSDR profile. The studies examine the impact of prenatal exposure to pyrethroid pesticides on fetal outcomes.
In a 2003 study of the effect of first- and second-trimester exposures to pesticides on birth outcomes in Poland, women (n=95) who reported residential or occupational exposure to pesticides, including pyrethroids, were found to be at an increased risk (p=0.01) of delivering a low-birth-weight infant, but the risk was not significant for pyrethoid exposure specifically (p=0.286), and there was no effect on pregnancy duration (Dabrowski et al., 2003). The study is limited in that exposures were to mixtures of pesticides and based on the mother’s recall of pesticide use at the time of delivery. As such, pesticide-specific effects cannot be evaluated.
The Volume 11 committee considered one study of birth outcomes in babies born to women in rural China who had prenatal exposure to pyrethroids (Ding et al., 2015). Maternal urine samples were analyzed for CDCCA, TDCCA, and PBA. The authors found that log10 unit increases in the total concentration of the metabolites, but not individual metabolite levels, were associated with a decrease in birth weight (β= –96.76, 95% CI= –173.15– –20.37). No associations were found between individual or total metabolite levels and birth length, head circumference, or gestational duration.
Zhang et al. (2014b) analyzed the urine of 147 Japanese women in their first trimester of pregnancy for PBA metabolites. Thyroid hormone levels—free thyroxine (FT4) and thyroid-stimulating hormone (TSH)—were determined in neonatal and maternal blood at birth. The levels of both thyroid hormones were within the normal range for all infants except one. There was no significant correlation between PBA concentrations in maternal urine and neonatal thyroid hormone concentration or with body size at birth (p=0.05), nor was there a significant correlation between the levels of maternal PBA and other neonatal outcomes of body weight, body length, head circumference, or chest circumference. However, in a multiple-regression analysis the authors found that PBA concentrations in maternal urine sampled in the first trimester were significantly associated with reduced birth weight (β=0.183, p=0.017) and head circumference (β=0.243, p=0.010).
The Baltimore THREE (Tracking Health Related to Environmental Exposures) study collected cord blood from 300 singleton live births delivered between November 2004 and March 2005 at Johns Hopkins Hospital. The cord blood was analyzed for the cis- and trans-isomers of permethrin, for the permethrin synergist piperonyl butoxide, and for nine cytokines. Infants were evaluated for birth weight, length, head circumference, ponderal index, and gestational age. Permethrin was negatively associated with the anti-inflammatory cytokine IL-10 (β= –0.14, 95% CI –0.22– –0.05). The permethrin concentrations measured in the cord blood had no significant associations with gestational age, birth weight, length, head circumference, or ponderal index (Neta et al., 2011).
The Volume 2 committee considered several studies on the reproductive effects of pyrethroids in animal models. Pyrethroids at 100–10,000 nM were found not to be estrogenic or antiestrogenic, nor
did they alter human estrogen-receptor α-mediated mechanisms in vitro (Saito et al., 2000; Sumida et al., 2001). The Volume 2 committee cited numerous studies that indicated that pyrethroids are not gonadotoxic, embryotoxic, or teratogenic (EPA, 1983; Hallenbeck and Cunningham-Burns, 1985; Kaloianova and El Batawi, 1991; Kidd and James, 1991; Litchfield, 1985; Miyamoto, 1976; Polakova and Vargova, 1983; Ray, 1991), although high oral doses (250 mg/kg per day) of permethrin during GDs 6–15 were found to reduce the number of offspring (Wauchope et al., 1992). A three-generation study of resmethrin reported slight increases in premature stillbirths and a decrease in pup weight (Ray, 1991). The Volume 2 committee concluded that pyrethroids are unlikely to cause reproductive and teratogenic effects (IOM, 2003).
The ATSDR review Toxicological Profile on Pyrethrins and Pyrethroids (2003b) reported that a few studies have indicated that relatively low-level, intermediate-duration oral exposure of adult male rats to some pyrethroids may result in damage to reproductive organs, abnormal sperm characteristics, reduced plasma testosterone levels, and reduced fertility. Relative to controls, male rats administered oral doses of deltamethrin had significantly reduced sperm cell concentrations, live cell percentage, and motility index; had a significantly higher percentage of total sperm abnormalities; and had significantly reduced plasma testosterone levels. Male fertility, determined by successful matings to untreated female rats, was 50% that of controls (Abd El-Aziz et al., 1994).
ATSDR also reported on an EPA study that showed decreased pup survival in rats following parental oral exposure to resmethrin at 70.8 mg/kg/day prior to mating and throughout gestation and lactation; rats receiving 80 mg/kg/day during GDs 6–15 had slight increases in skeletal variations and delayed ossification (EPA, 1994).
The Volume 11 committee identified several animal studies that assessed the reproductive toxicology of all three pyrethroids of interest. One study examined the effect of 20 or 40 mg/kg/day permethrin on rat ovaries (Kotil and Yon, 2015). After treatment for 14 days via gavage, dose-dependent degenerative changes were seen in the follicular and corpus luteum cell morphology, with evidence of apoptotic cell death. All the other studies assessed the reproductive toxicity of pyrethroids in male animals.
In studies of male reproductive effects, Zhang et al. (2007, 2008) dosed male ICR mice with cispermethrin at 0, 35, or 70 mg/kg/day for 6 weeks. Caudal epididymal sperm count and sperm motility, testicular testosterone production, and plasma testosterone concentrations were all significantly and dose-dependently decreased, with an increase in LH. Testicular mitochondrial mRNA expression was also reduced, with mitochondrial membrane damage in the Leydig cells. Cis-permethrin administered at 90 μmol/kg/day for 6 weeks did not affect plasma testosterone levels, although the co-exposure of animals with 10 μmol/kg/day diazinon reduced plasma testosterone levels to 45% of those in the control animals. Cis-permethrin alone also reduced translocator protein mRNA levels by 46% compared with controls but did not reduce P450scc mRNA levels. Administration of permethrin to adult male rats at 35 mg/kg/day for 60 days resulted in reductions in reproductive organ weights, the number of live sperm and Leydig cells, sperm motility, and serum testosterone levels and in increased abnormal sperm morphology (Khaki et al., 2017).
Deltamethrin was also found to have adverse effects on male reproduction in animal studies, including reduced sperm quality, sex hormones, and libido and increased testicular damage (Ben Slima et al., 2017). Issam et al. (2009) treated male rats with deltamethrin at 2 ppm for 30 days, 20 ppm for 45 days, and 200 ppm for 60 days and reported a significant decrease in FSH, LH, and testosterone at the highest dose. The deltamethrin treatment of rats also arrested spermatogenesis and produced a significant increase in malondialdehyde levels that was related to dose, the length of treatment, and lipid peroxidation. Abdallah et al. (2010) studied the effect of deltamethrin at doses of 0, 10, 50, 100 and 200 μmol on the oxidative stress response in rat spermatozoa after 3 hours of incubation. This in vitro exposure caused
a significant decline in sperm motility and viability and increased abnormal sperm morphology as well as levels of malondialdehyde, superoxide dismutase, and catalase. The administration of deltamethrin at 5 mg/kg/day to male rats for 4 weeks resulted in decreased serum testosterone, LH, and FSH levels and significantly increased testicular total oxidant capacity, poly(ADP-ribose) polymerase, lactate dehydrogenase, and DNA damage (Ismail and Mohamed, 2012). A dose of 1 mg/kg deltramethrin administered intraperitoneally for 5 days/week for 1 month increased the weight of rat testes and produced adverse histopathologic changes. There was also a significant reduction in the number of seminiferous tubules per unit area and adverse morphologic changes in the tubules (Kumar and Nagar, 2014).
Cypermethrin was found to have similar effects to deltamethrin on male rodents. Elbetieha et al. (2001) studied the effects of cypermethrin on reproductive and fertility parameters in male mice receiving 0, 13.15, 18.93, and 39.66 mg cypermethrin in drinking water for 12 weeks. All doses resulted in a significant reduction in the number of viable fetuses in females mated to the treated males, lower body weight gain in the treated males, a significant increase in preputial gland weights, and a significant decrease in epididymal and testicular sperm counts and daily sperm production. At the two higher doses, there was a significant increase in the weight of the testes and seminal vesicles and a significant reduction in the perimeter and number of cell layers of the seminiferous tubules, along with hemorrhage in the testes and a large number of immature spermatids in the tubules. The two lower doses (13.15 and 18.93 mg) produced significantly reduced fertility, as indicated by fewer females being impregnated by the treated males, while 39.66 mg cypermethrin caused a significant reduction in the number of implantation sites. On the other hand, Al-Hamdani and Yajurvedi (2010) did not see a decrease in fertility in mice administered cypermethrin by gavage at 1.38, 2.76, and 5.52 mg/kg/day for 6 or 12 weeks. All doses and both durations caused a significant reduction in epididymal spermatozoa count and an increase in abnormal spermatozoa, but 6 weeks after dosing ceased, the counts returned to normal in mice administered the two lower doses but not the highest dose. All mice had normal fertility, although the weight of the litters was significantly lower than for controls; the gestation period and the litter size were unaffected. Dahamna et al. (2010) also found that cypermethrin administered to male mice at 20% and 5% of the LD50 (485 mg/kg) for 2 weeks and 4 weeks, respectively, and at 20% of the LD50 for 12 weeks, resulted in a 20% decrease in sperm count and an increase in abnormal sperm and in testicular and epididymal morphology. Similar results were seen by Hu et al. (2013) who treated rats with cypermethrin at 0, 6.25, 12.5, 25 and 50 mg/kg/day by gavage for 15 days. At 25 and 50 mg/kg there was a significant reduction in sperm production and changes in the seminiferous tubules, including reduced and deformed spermatogonia. Furthermore, at 50 mg/kg cypermethrin, androgen receptor levels were significantly reduced, as were serum testosterone levels. Rodriguez et al. (2017) also found adverse effects on Sertoli cells in mice administered 20% of the LD50 intraperitoneally.
Sharma et al. (2018) found that the administration of cypermethrin, deltamethrin, or a combination of the two pesticides significantly reduced sperm count and motility and concentrations of some steroidogenic enzymes and increased the number of abnormal sperm; the combination of the two pesticides had a greater effect on all sperm parameters and enzyme levels. Both pesticides and their combination also significantly reduced testosterone and LH concentrations, but only the combination of the two pesticides significantly reduced concentrations of FSH.
The mechanism by which pyrethroids act on male fertility was explored by Taylor et al. (2010) using an in vitro mouse Sertoli cell assay. Pyrethroids, at both low and high serum concentrations, affected the expression of the two estrogen receptors; however, the influence on estrogen receptor gene expression was different from the effect of exposure to 17beta-estradiol.
The ATSDR review Toxicological Profile on Pyrethrins and Pyrethroids (2003b) did not report any studies of developmental effects in humans following maternal or paternal exposure to pyrethroids by any route of exposure. The majority of the studies published since the ATSDR profile examined neurodevelopmental outcomes in children following prenatal exposures to pyrethroids, although the committee did consider a few other studies that examined other developmental outcomes in children.
The Volume 11 committee considered several studies of the developmental effects of pyrethroids. The studies are summarized in Table 5-5 and discussed in the following section.
The CCCEH study assessed 348 children born between February 1998 and May 2002 for psychomotor and cognitive development at several times during their childhoods (Horton et al., 2011). In this study, the children were administered the BSID-II to generate MDI and PDI scores. Prenatal exposure to permethrin was measured in maternal and umbilical cord plasma collected on delivery, and permethrin and piperonyl butoxide levels were measured in personal air collected during pregnancy. When the children were assessed at 36 months of age, the authors found no significant associations between their MDI or PDI scores and permethrin concentrations in either the personal air samples or the maternal or umbilical cord plasma samples. However, the highest prenatal piperonyl butoxide exposure was associated with a 3.9-point mean decrement in MDI scores (95% CI –0.25– –7.49), and the most highly exposed children had the greatest risk (OR=3.11, 95% CI 1.38–6.98) of delayed mental development.
In a further follow-up of the CHAMACOS study in California described in the section on OPs, Gunier et al. (2017) used the Pesticide Use Reporting data from California to determine the use of pyrethroids in fields near the maternal residence during pregnancy. Neurodevelopment in children at age 7 was assessed using WISC-IV. Pyrethroid use was associated with lower scores on the full-scale IQ (β= –2, 95% CI –3.7– –0.3), verbal comprehension (β= –1.8, 95% CI –3.4– –0.3), and perceptual reasoning (β= –2.1, 95% CI –4.0– –0.2) subscales, but not on the working memory or processing speed scores.
The Mt. Sinai Children’s Environmental Health Cohort, described in the section on OPs, assessed 162 mother–infant pairs during pregnancy in 1998–2001. Prenatal exposure to pyrethroids was determined by an analysis of maternal urine samples taken in the third trimester of pregnancy; the samples were analyzed for PBA, TDCCA, and CDCCA. Children were assessed for neurodevelopment at 4, 6, and 7–9 years of age. Parents reported on their children’s development using the Behavioral Assessment System for Children and the Behavior Rating Inventory of Executive Function. Detectable levels of PBA during pregnancy were associated with lower scores for the indices of internalizing (β= –4.50, 95% CI –8.05– –0.95), depression (β= –3.21, 95% CI –6.38– –0.05), somatization (β= –3.22, 95% CI –6.38– –0.06), behavioral regulation (β= –3.59, 95% CI –6.97– –0.21), emotional control (β= –3.35, 95% CI –6.58– –0.12), shifting (β= –3.42, 95% CI –6.73– –0.11), and monitoring (β= –4.08, 95% CI –7.07– –1.08). Detectable levels of CDCCA were associated with worse externalizing (β= –4.74, 95% CI –9.37– –0.10), conduct problems (β= –5.35, 95% CI –9.90– –0.81), behavioral regulation (β= –6.42, 95% CI –11.39– –1.45), and inhibitory control (β= –7.20, 95% CI –12.00– –2.39) (Furlong et al., 2017b).
The impact of maternal exposure to pyrethroids on the development of 102 18-month old Japanese infants was examined by Hisada et al. (2017) using the Kinder Infant Development Scale, which is based on a questionnaire completed by the caretaker. Infants whose mothers had higher levels of PBA in their urine at 10–12 weeks of gestation had higher development quotient scores than infants whose mothers had lower PBA concentrations (p=0.017), which the authors considered to be unexpected.
Other studies have found in utero pyrethroid exposure to be associated with adverse effects on neurodevelopment. Xue et al. (2013) found that among 497 mother–infant pairs in China, increasing levels of pyrethroid metabolites in maternal urine during pregnancy were inversely correlated with the development quotient of the infants at 1 year of age (β= –0.1527, p=0.05), although there was no such correlation for the MDI. The Development Screen Test scale was used to assess the intellectual development of infants. Children born to mothers residing in rural areas had greater impairment than children born to mothers living in cities. The impairment was correlated with higher levels of pyrethroid metabolites in the urine of mothers residing in rural areas.
The impact of maternal exposure to pyrethroids on child neurodevelopment was also assessed by Watkins et al. (2016) in the ELEMENT longitudinal study of a mother–child cohort in Mexico City recruited from 1997 to 2001. PBA was measured in the urine of 187 mothers during their third trimester, and children were administered the BSID-II Spanish version (BSID-IIS) to assess the developmental functioning of infants and children at 24 and 36 months of age using the MDI and PDI. Children whose mothers had medium and high urinary PBA levels had nonsignificantly lower MDI scores (β= –3.54, 95% CI –7.86−0.78 and β= −3.80, 95% CI –8.44−0.84, respectively) at 24 months than those with mothers in the low (reference) PBA category (p-value for trend=0.07), and the effect was lower at 36 months (p-value for trend=0.14). At 24 months the associations were significantly stronger for female children (β= –6.20, 95% CI –12.3− –0.14) than for males (β= –2.58, 95% CI –9.10−3.94) of mothers with medium PBA levels, but the differences were not significant at 36 months. The third-trimester prenatal level of maternal PBA was not associated with PDI scores at either time point.
Eskenazi et al. (2018) assessed the effects of prenatal exposure to multiple pesticides on neurodevelopment, as measured by the BSID-II, in 752 children in South Africa participating in the Venda Health Examination of Mothers, Babies and Their Environment (VHEMBE) birth cohort study. Maternal exposure to pyrethroids during pregnancy was determined through the identification of four urinary pyrethroid metabolites—DBCA, CDCCA, TDCCA, and PBA. Maternal urine samples were collected around the time of delivery. When the children were 1 year of age, each 10-fold increase in CDCCA, TDCCA, and PBA was significantly associated with a decrement in social–emotional scores (β= –0.63, 95% CI –1.14– –0.12; β= –0.48, 95% CI –0.92– –0.05; and β= –0.58, 95% CI –1.11– –0.06, respectively). At 2 years of age, each 10-fold increase in maternal CDBCA levels was associated with significant decrements in language composite scores (β= –1.74, 95% CI –3.34– –0.13) and expressive communication scores (β= –0.40, 95% CI –0.77– –0.04). The authors noted that “significant differences by sex were estimated for pyrethroid metabolites and motor function scores at 2 years of age, with higher scores for boys and lower scores for girls.”
The PELAGIE study in Brittany, France, is a large mother–child cohort that has assessed child development since the children were born between 2002 and 2006 (Viel et al., 2015). Maternal exposure to pyrethroids was determined on the basis of urine samples collected between 6 and 19 weeks of pregnancy and adjusted for exposure to OPs. At a child’s sixth birthday, the mother and child participated in a neuropsychological follow-up using the WISC-IV; a total of 287 mother–child pairs took part. None of the WISC scores were associated with maternal pyrethroid metabolite concentrations (PBA, TDCCA, or CDCCA). The children’s behavior was further evaluated for three subscales of the Strengths and Difficulties Questionnaire: prosocial behavior, internalizing disorders, and externalizing disorders (Viel et al., 2017). Maternal prenatal CDCCA was associated with internalizing difficulties, while maternal prenatal PBA was associated with externalizing difficulties and higher odds of abnormal or borderline social behavior per the maternal report on the Strengths and Difficulties Questionnaire (Viel et al., 2017).
The CHARGE study in California attempted to link pesticide use near the mother’s residence with children having ASD (n=486) or developmental disorder (n=168) (Shelton et al., 2014). Prenatal
residence proximity (within 1.5 km) to areas sprayed with pyrethroids was associated with an increased risk for ASD if the exposure was prior to conception (OR=1.82, 95% CI 1.00–3.31) or in the third trimester (OR=1.87, 95% CI 1.02–3.43). Exposure at 1.75 km was associated with an increased risk for ASD only in the third trimester (OR=1.83, 95% CI 1.04–3.23). For developmental disorders and residential proximity to agricultural pesticide application, there was a significant risk only for a distance of 1.75 km during the third trimester (OR=2.34, 95% CI 1.18–4.67), indicating that the timing of exposure during pregnancy had a significant effect on the risk of ASD and developmental disorder.
Other Developmental Effects
Researchers have linked other developmental endpoints to prenatal exposure to pyrethroids. Carmichael et al. (2014), discussed in the section on OPs, compared the exposure of mothers of 569 infants with any of eight congenital heart defects with 785 mothers who had an infant without a birth defect. Prenatal exposure to permethrin was nonsignificantly associated with ventricular septal defect, perimembranous (OR=2.2, 95% CI 0.8–5.7), but not with any of the heart defects. Monge et al. (2007) examined the effect of maternal (n=876) and paternal (n=762) exposures to pesticides on the risk of childhood leukemia (n=334) in Costa Rica in 1995–2000. Parents were interviewed about risk factors for childhood leukemia, and their exposures were rated by hazard value and by time of occurrence. Neither maternal nor paternal exposure to pyrethroid insecticides, primarily deltamethrin, at any time prior to birth was significantly associated with childhood leukemia for either boys or girls.
The association between prenatal exposure to permethrins and piperonyl butoxide, a synergist for residential pyrethroid pesticides, and the presence of childhood cough at 5–6 years of age was examined by Liu et al. (2012) as part of the CCCEH birth cohort. Prenatal exposure to cis- and trans-permethrin was not significantly associated with a noninfectious cough at age 5–6 years, but there was a positive association between a noninfectious cough and prenatal piperonyl butoxide in children ages 5–6 years (OR=1.27 for unit change in log piperonyl butoxide, 95% CI 1.09–1.48). Exposure to piperonyl butoxide or permethrins at ages 5–6 years was not associated with an increase in childhood cough.
Coker et al. (2018) assessed the effects of prenatal exposure to multiple pesticides on child body weight and body composition in 708 children in South Africa’s VHEMBE birth cohort. Maternal exposure was determined by the identification of four urinary pyrethroid metabolites—DBCA, CDCCA, TDCCA, and PBA. Maternal urine samples were collected around the time of delivery. The length or height and the weight of the children were measured at 1 and 2 years of age; maternal exposure to dichlorodiphenyltrichloroethane (i.e., DDT) was also assessed. Using a single-pollutant linear mixed-effects model, BMI-for-age and weight-for-height were inversely related to DBCA and TDCCA levels; the effects were more pronounced for boys than for girls for both measures and metabolites. Weight-forage was also inversely related to TDCCA for boys only, but not overall. None of the four metabolites significantly affected any of the measures for girls.
The Volume 2 committee reported that pyrethroids administered to female rats during pregnancy did not affect the development of their offspring (Extoxnet, 1994). No other animal studies on the developmental effects of pyrethroids were cited in Volume 2.
The ATSDR review Toxicological Profile on Pyrethrins and Pyrethroids (2003b) did not find the typical signs of developmental toxicity in animals exposed to pyrethrins or pyrethroids, particularly deltamethrin. There were no studies of reproductive or developmental effects following inhalation.
ATSDR reported that numerous studies of pyrethorids administered orally in rats and mice did not indicate any developmental toxicity at doses below those which cause maternal toxicity. In utero exposure of rats resulted in cellular changes indicative of compromised immunological function. Santoni et al. (1997, 1998, 1999) reported treatment-related increases in natural killer (NK) cell and antibody-dependent cytotoxic activity, impaired thymocyte function, increased levels of T cells in peripheral blood, and decreased levels of T cells in the spleen in rats after their mothers were given 50 mg/kg/day cypermethrin orally during gestation—a dose schedule that did not result in clinical signs of maternal toxicity. Malaviya et al. (1993) observed significant increases in the levels of dopamine and muscarinic receptors of striatal membrane in rat pups that had been exposed to 15 mg cypermethrin/kg on GDs 5–21.
The Volume 11 committee identified three new studies and a systematic review that assessed the effects of preconception and prenatal exposure to permethrin on the development of offspring. In a two-generation study with mice, males and females were given 0, 4.9, 9.8, and 19.6 mg/kg/d permethrin by gavage for 4 weeks before mating (Farag et al., 2006). The highest dose produced signs of toxicity in the F0 mice, with body weight loss in the females during gestation and lactation. There was a significant (p<0.05) reduction in fetal weight and in the number of live pups born to females receiving the two highest doses. Postnatal assessment of the F1 offspring showed that at 9.8 and 19.6 mg/kg there was a significant reduction in righting reflexes, swimming ability, and open field activity and a slowed development of cliff avoidance; these effects were not seen in pups whose dams received 4.9 mg/kg permethrin. Imanishi et al. (2013) found other effects of prenatal permethrin exposure on neurodevelopment. Pregnant mice received a single dose of 0, 2, 10, 50, or 75 mg/kg permethrin on GD 10.5. On GD 17.5, exposure to 10 mg/kg resulted in significant increases in abnormalities of the anterior communicating arteries in the brain of male fetuses but not of female fetuses. Fetal brains obtained on GD 17.5 had anatomical abnormalities, including altered vascular formation, and on PND 7 rat brains had alterations in the thickness of the neocortex and hippocampus and significantly increased norepinephrine and dopamine levels. Even the lowest dose of 2 mg/kg showed significant influences on vascular development in fetuses and on adult offspring behaviors. At 8 and 12 weeks of age, male mice had a significant reduction in standing ability and locomotor activity at the 2 mg/kg dose; these effects were not observed in female offspring.
One new study on the effects of deltamethrin on adult male offspring of treated dams was identified (Ben Slima et al., 2012). Pregnant mice received 5 mg/kg/day deltamethrin by gavage from GD 3 to GD 21. Male offspring were assessed for fertility when the animals achieved sexual maturity. Although the body weights of the dams were significantly reduced, there were no significant differences in absolute and relative epididymides weights in offspring. However, there were significant differences in absolute and relative testes weights (reduced to approximately half of the control value, p≤0.05); reduced spermatozoa counts and percent of motile and viable sperm; increased abnormal sperm; and increases in testicular histopathology.
Saillenfait et al. (2017) assessed the effects of prenatal exposure to cypermethrin on fetal testicular steroidogenesis. Cypermethrin (0.1, 1, 5, or 10 mg/kg/day) administered on GDs 13–19 had no effect on anogenital distance of the male fetuses assessed on GD 19, but the two highest doses reduced testosterone production. Testicular mRNA expressions of HMG-CoA synthase and reductase, and the genes SRB1, StAR, P450scc, 3βHSD, P450 17A1, and 17βHSD, were not affected at any dose.
A systematic review of inherited transgenerational effects was conducted by the National Institute of Environmental Health and Sciences (Walker et al., 2018). The researchers found that there was no evidence of transgenerational effects of pesticide exposure in humans. However, there were three rodent studies—all from the same group of investigators—that examined whether transgenerational effects might occur following gestational exposure to a mixture of permethrin and DEET. The studies assessed
the effects of the pesticide mixture on growth and development, reproductive effects in offspring, and renal effects in offspring. Two studies evaluated the body weight of F3 offspring and reported no impact of exposure on post-weaning weight at PND 21, but an increased weight at 12 months of age (Manikkam et al., 2012a,b). Two studies reported on male reproductive effects with no associations for most endpoints (Manikkam et al., 2012a,b). Three studies (Manikkam et al., 2012a,b; Nilsson et al., 2012) reported decreased ovarian follicle counts and increased ovarian cysts in female rats of the F3 generation of F0 gestating females exposed to 190 mg/kg/day permethrin and DEET, but inconsistent effects on puberty and no effects on fertility and ovarian weight were reported. Confidence in these findings is limited by inconsistencies across the studies (heterogeneity of endpoints, differences in the measurement of endpoints), a lack of studies from other research groups, and a lack of analysis using the litter as the unit of analysis.
Synthesis and Conclusions
Pyrethroid insecticides have been used for many years and were widely used in both the 1990–1991 Gulf War and the Post-9/11 conflicts. In particular, permethrin was used to treat uniforms and bed nets to control sand flies, which were vectors for infectious diseases such as leishmanisis.
Several studies have assessed the effects of exposure to pyrethroids on human male reproductive parameters. Most of the studies assessed pyrethroid exposure by measuring the levels of the metabolite PBA in urine. With the exception of one study in Japanese men (Imai et al., 2014), which found no association between PBA levels and semen parameters or male reproductive hormones, the other five studies considered by the committee found that increasing levels of pyrethroid metabolites in urine were associated with decreased sperm concentrations and motility and increased sperm abnormal morphology and DNA damage and disomy. Numerous animal studies in rats and mice also indicate that exposure to permethrin, cypermethrin, and deltamethrin has adverse effects on male reproduction, especially sperm parameters and reproductive hormones. One limitation of these studies is that the PBA reflects only recent exposure, given the short half-life of pyrethroids, and in most cases only a single measurement of pyrethroid metabolites was made.
The Volume 11 committee identified one study of the effects of pyrethroid exposure on female reproduction and it showed increased TTP and reduced fertility (Hu et al., 2018).
Several epidemiologic studies on the effects of prenatal exposure to pyrethroids on birth outcomes were considered. When the assessed exposure to pyrethroids was based on the concentration of metabolites in urine, there was a negative association between exposure and birth weight, although this was not seen in a study where permethrin exposure was based on its concentration in cord blood. There were no animal studies that added to the evidence base for effects on female reproduction or birth outcomes.
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between exposure to pyrethroid pesticides 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 pyrethroid pesticides and reproductive effects in women, or with adverse pregnancy outcomes.
The effects of prenatal pyrethroid exposure on the neurodevelopment of children were assessed in several studies with mixed results. Hisada et al. (2017) found that mothers’ exposure to pyrethroids during the first trimester of pregnancy was positively associated with their children’s development scores at 18 months of age—that is, greater exposure in the mothers was associated with better performance in the children. This conclusion was contradicted by several other studies, including that of Watkins et al. (2016), who found that children born to mothers with higher PBA concentrations during any trimester of pregnancy had lower scores on developmental tests at 24 months of age; a study by Xue et al. (2013) in China that reported an association between pyrethroid metabolite concentrations in the mother’s urine and the developmental quotient in their children at 1 year of age; a study by Gunier et al. (2017) that used geospatial data to estimate prenatal exposure and found negative associations with child IQ at age 7 years; a study by Furlong et al. (2017b) that found increased PBA and CDCCA were both associated with decrements in children’s neurodevelopment at 4, 6, and 7–9 years of age; and a study by Shelton et al. (2014) that found that preconception or third-trimester exposure (but not first- or second-trimester exposure) to pyrethroids (as determined by proximity to areas of agricultural pesticide application) was associated with an increased risk of ASD. However, the PELAGIE study in France found no significant association between a mother’s urinary pyrethroid metabolite concentrations during early pregnancy and cognitive outcomes in their children at age 6 years, with the exception of one developmental subscale (Viel et al., 2015, 2017), and a study from Japan (Hisada et al., 2017) found better developmental quotients associated with prenatal exposure.
Prenatal exposure to pyrethroids was not associated with childhood leukemia. Studies of the effects of prenatal exposure to permethrin and piperonyl butoxide on noninfectious cough at ages 5–6 years were inconclusive. Although exposure to pyrethroids was associated with reduced child growth at 1 and 2 years of age, the effect was limited to boys only and for only two of the four metabolites measured (Coker et al., 2018).
Animal studies reviewed by the Volume 2 committee and ATSDR did not indicate that prenatal exposure of animals to pyrethroids causes postnatal developmental effects, although one research group did find effects on immune function that deserve further study. The Volume 11 committee considered two additional animal studies of pyrethroids that indicated that prenatal exposure at levels that do not cause maternal toxicity may result in neurodevelopmental effects (structural brain abnormalities, poor performance on neurodevelopmental tests) in offspring. Although there is evidence in rodent models that transgenerational effects are possible (Walker et al., 2018), their relevance to human health is not yet clear.
The Volume 11 committee concludes that there is limited/suggestive evidence of an association between prenatal exposure to pyrethroid pesticides and developmental effects.
TABLE 5-5 Summary of Reproductive and Developmental Effects of Pyrethroid Pesticides
|Reproductive Effects in Men|
|Xia et al. (2008)||376 eligible men provided semen and urine samples at infertility clinic in Nanjing, China; recruited between March 2004 and March 2006.||Urinary concentrations of PBA were assessed.||Compared with men in the lower PBA quartiles, men in higher quartiles were more likely to have below-reference sperm concentration (ORs for increasing exposure quartiles, 1.00, 1.31 [95% CI 0.65–2.64]; 1.73 [95% CI 0.87–3.45]; 2.04 [95% CI 1.02–4.09]; p-value for trend=0.027). No statistical significance was found between the three higher 3 PBA quartiles and the lowest PBA quartile in terms of below-reference sperm number per ejaculum and sperm motility. For semen volume, although the OR of quartile 3 was significantly higher than that of quartile 1, the exposure–response trends were not monotonic.|
|Ji et al. (2011)||240 men recruited from an infertility clinic in China between April 2005 and March 2007.||Urinary concentrations of PBA were assessed.||Significant inverse correlation was observed between the urinary PBA level and the sperm concentration (β= –0.27, 95% CI –0.41– –0.12, p<0.001). A significant positive correlation between urinary PBA level and sperm DNA fragmentation (β=0.27, 95% CI 0.15–0.39, p<0.001) was found.|
|Imai et al. (2014)||323 men from a larger cross-sectional, multi-center study. Inclusion criteria were (1) university student, (2) 18–24 years old, and (3) he and his mother were born in Japan. Recruitment in Tokyo May 1999–May 2000 and April 2002–May 2003.||Urine and semen collection. Urine analyzed for PBA; PBA was detected in 294 of 322 samples of urine (91.3%).||No significant difference in PBA levels between men with a semen parameter below the 2010 WHO lower reference limit (n=23–52, depending on the parameter) and those with a value greater than the lower reference limit (t-test, p>0.05).|
|Yoshinaga et al. (2014)||See Imai et al. (2014) above.||Urine, blood, and semen samples collected. Concentrations of LH, FSH, SHBG, testosterone, and inhibin B in serum.||There was no significant difference between the four quartiles of urinary PBA excretion for any of the hormones except for testosterone/LH). Mean testosterone/LH was significantly higher in the 3rd quartile than in the 4th quartile (p<0.05); however, no linear trend was found.|
TABLE 5-5 Continued
|Jurewicz et al. (2015)||See Radwan et al. (2014); 286 men participated.||See Radwan et al. (2014). Comet assay procedure used to assess sperm DNA damage.||A positive association was observed between CDCCA >50th percentage distribution of concentration and the percentage of medium DFI and the percentage of immature sperm (p=0.04, p=0.04, respectively). The level of PBA >50th percentage distribution in urine was positively related to the percentage of high DFI (p=0.03). TDCCA and DBCA levels and the sum of pyrethroid metabolites were not associated with any sperm DNA damage measures.|
|Radwan et al. (2015)||See Radwan et al. (2014); 195 men participated.||Semen and urine analysis as in Radwan et al. (2014). Sperm aneuploidy was measured by multicolor fluorescence in situ hybridization (FISH) test.||Association between CDCCA >50th percentile and disomy of chromosome 18 (p=0.05), but TDCCA in urine >50th percentile was related to XY disomy (p=0.04) and disomy of chromosome 21 (p=0.05). PBA level ≤50th and >50th percentile was related to disomy of sex chromosomes: XY disomy (p=0.05 and p=0.02, respectively), Y disomy (p=0.04 and 0.02, respectively), disomy of chromosome 21 (p=0.04 and p=0.04, respectively), and total disomy (p=0.03 and p=0.04, respectively). Disomy of chromosome 18 was positively associated with PBA >50th percentile (p=0.03). No association was found between DBCA level in urine and any of analyzed chromosome disomy.|
|Radwan et al. (2014)||334 men who attended infertility clinics in Lodz, Poland, for diagnostic purposes. The mean age of men was 32.2 years. Semen and urine samples were collected.||A spot urine sample was collected from each subject and analyzed for PBA, CDCCA, TDCCA, and DBCC.||Positive association was observed between CDCCA levels in urine >50th percentile and the percentage of sperm with abnormal morphology (p=0.05). The level of TDCCA in urine >50th percentile was negatively related to sperm concentration and the level of testosterone (p=0.04 and p=0.04, respectively). In addition, the level of TDCCA in urine <50th percentile and >50th percentile increases the percentage of sperm with abnormal morphology (p=0.01 and p=0.02, respectively). The concentrations of PBA and DBCC >50th percentile were negatively associated with one of the computer-aided semen analysis parameters, LIN (p=0.05 and p=0.01, respectively). The sum of pyrethroids >50th percentile was positively associated with the percentage of sperm with abnormal morphology (p =0.04).|
|Meeker et al. (2008b)||207 men between 18 and 54 years of age were recruited between January 2000 and April 2003 from Massachusetts General Hospital.||Comet assay procedure used to assess sperm DNA damage. A spot urine sample was collected from each subject and analyzed for PBA, CDCCA, and TDCCA.||Men in the highest PBA category had declines in sperm concentration and sperm motility. TDCCA concentration >75th percentile was also associated with a suggestive 20.3 million sperm/ml decline in sperm concentration (95% CI 235.3– +2.6) and a statistically significant 10% (95% CI –19% – –21%) reduction in sperm motility. In the comet assay, PBA and CDCCA were associated with increased sperm DNA damage measured as Tail%, but not comet extent or TDM. Association of PBA categories and Tail% was monotonic (p-value for trend=0.02). Regression coefficients for medium and high PBA groups were equivalent to increases in Tail% of 7.5% (95% CI –10–25%) and 20% (3–37%), respectively, relative to the study population median of 32.3% of DNA in comet tail.|
|Young et al. (2013)||75 men who were selected from a parent study of 341 patients at Massachusetts infertility clinic (see Meeker et al., 2008b). 90% of the men were white and non-Hispanic (96%), average age 35 years.||Three pyrethroid metabolites—PBA, CDCCA, and TDCCA—measured in urine.
Multiprobe fluorescence in situ hybridization for chromosomes X, Y, and 18 was used to determine XX, YY, XY, 1818, and total sex chromosome disomy in sperm nuclei.
|Disomy rates were increased by 7% to 30% across the aneuploidy outcomes, with the highest risk seen for TDCCA and XX18 (IRR=1.30, 95% CI 1.17–1.46), the lowest risk seen for TDCCA and XY18 (IRR=1.07, 95% CI 1.00–1.14), and no association with 1818. In men with PBA concentrations >LOD versus those <LOD, rates of YY18 disomy were 1.28 times higher (95% CI 1.15–1.42), whereas a lower rate (IRR=0.82, 95% CI 0.77–0.87) was seen for XY18 and total disomy (IRR=0.93, 95% CI 0.87–0.97). No association was observed between PBA and XX18 or 1818 disomy.|
TABLE 5-5 Continued
|Meeker et al. (2009)||161 men from an infertility clinic recruited during 2000–2003 from Massachusetts General Hospital; 18–54 years of age; measured serum reproductive and thyroid hormone levels.||Measured pyrethroid metabolites PBA, CDCCA, and TDCCA in spot urine samples.||Compared with men who had values below the median, men in the highest PBA category had significantly higher FSH levels (p=0.001; p-value for trend among low, medium, and high categories=0.002) but not higher levels of LH (p=0.054). When adjusting for potential confounders, categories for all three metabolites, as well as their summed values, were positively associated with FSH (all p-values for trend <0.05). Statistically significant relationships with LH were also found. CDCCA and TDCCA were inversely associated with inhibin B (p for trend=0.03 and 0.02, respectively). Finally, there was evidence that TDCCA was inversely associated with testosterone and the free androgen index (the ratio of testosterone to sex hormone binding globulin; p-value for trend=0.09 and 0.05, respectively).|
|Reproductive Effects in Women|
|Hu et al. (2018)||615 women in Shanghai, China, recruited from two preconception care clinics between August 2013 and April 2015. 496 women left for study analysis.||PBA, TDCCA, and CDCCA measured in urine. OP exposure also measured. Women asked about pesticides exposure and perceived stress and lifestyle.||Women in the highest quartile for PBA had longer TTP (fecundability OR=0.72, 95% CI 0.53–0.98) and increased infertility (OR=2.03, 95% CI 1.10–3.74). Concentrations of TDCCA and CDCCA were below the level of detection and were not further examined.|
|Adverse Pregnancy Outcomes|
|Dabrowski et al. (2003)||Women who delivered their children in 25 maternity hospitals in the region of Lodz, central Poland. 389 women and 377 infants with birth weight ≥2,500 g. Infants delivered randomly on 70 selected days between January 31, 1998, and June 30, 2001.||Questionnaire on pesticide use and involvement in heavy physical work on the farm in each pregnancy trimesters was administered by a physician 1-2 days after delivery.
95 women (19.3%) reported pesticide exposure in first or second trimester of pregnancy.
|Infants born to women exposed to pesticides in first or second trimester had birth weight lower by 189 g than that of infants of the nonexposed women. When adjusted for pregnancy duration, the women exposed to pesticides delivered infants with birth weight lower by about 100 g (p=0.067) compared with nonexposed women.
Mothers exposed to pesticides on average delivered half a week earlier than those nonexposed (p=0.05).
|Ding et al. (2015)||Prospective cohort in rural northern China. 454 mother–infant pairs recruited between September 2010 and 2012.||Pyrethroid metabolites in maternal urine at delivery analyzed. Assessed birth outcomes, including birth weight, length, head circumference, and gestational duration. Creatinine-adjusted medians of pyrethroid metabolites in urine were 0.51 μg/g for CDCCA, 0.65 μg/g for TDCCA, and 0.68 μg/g for PBA.||Increase in total metabolites, but not any individual metabolite, was associated with a decrease in birth weight (adjusted β= –96.76 g per log10 unit increase, 95% CI –173.15– –20.37).|
|Zhang et al. (2014b)||147 mother–infant pairs recruited during early pregnancy at a Tokyo hospital in Japan 2009–2011. Blood and urine samples obtained in the first trimester; infant blood collected 5 days postpartum.||Concentrations of FT4, TSH, and TBG in maternal serum; PBA and iodine concentrations in maternal urine; and of FT4 and TSH in neonatal blood were determined.||No significant correlation of PBA concentration in maternal urine of the first trimester of gestation with neonatal thyroid hormone concentration (FT4, p=0.09; and TSH, p=0.69) or with body size at birth (p>0.05). No nonlinear relationship between prenatal pyrethroid exposure and neonatal outcomes.|
|Neta et al. (2011)||Baltimore Tracking Health Related to Environmental Exposures study; cord blood was collected from 341 births between November 26, 2004, and March 16, 2005.||185 (62%) cord blood samples were analyzed for cis- and trans-permethrin isomers and PBO.||Permethrin levels were (p-value <0.05) negatively associated with anti-inflammatory response (cytokine component 3) and positively associated with an IL-1β response (cytokine component 5). Associations were significant only for most highly exposed group versus lowest exposure group. A one-unit change in permethrin levels was associated with a nonsignificant 1-day decrease in gestational age (95% CI –2–0.2 days); most highly exposed group was significantly associated with a 2-day decrease versus lowest-exposure group.|
|Horton et al. (2011)||CCCEH cohort study. Subjects included 230 mother–child pairs selected from an ongoing prospective cohort study. Pregnant women ages 18–35 years who self-identified as either African American or Dominican.||Children assessed at age 36 months with BSID-II. Children’s home environments were evaluated at 3 years of age; maternal prenatal urine samples from third trimester and maternal personal air samples (48 hrs) were collected during the third trimester of pregnancy.||No significant association between permethrin concentration in air and 36-month MDI. However, there was a significant risk of lower MDI scores associated with the highest quartile PBO concentrations versus controls (OR= –4.57, 95% CI –0.30– –8.84). The highest category of PBO exposure was associated with a 3.9-point mean decrement in MDI score (95% CI –0.25– –7.49).|
TABLE 5-5 Continued
|Gunier et al. (2017)||CHAMACOS cohort, see Eskenazi et al. (2004) 283 mother–child pairs.||Children assessed at 7 yrs old for neurodevelopment using WISC-IV. Agricultural pesticide use within 1 km of mother’s residence during pregnancy estimated using California Pesticide Use Reporting data from 1999–2001.||Pyrethroid use was associated with lower scores on the full-scale IQ (β= –2, 95% CI –3.7– –0.3) and the verbal comprehension (β= –1.8, 95% CI –3.4– –0.3) and perceptual reasoning (β= –2.1, 95% CI –4.0– –0.2) subscales, but not working memory or processing speed scores. There were no significant changes in other WISC-IV domains.|
|Furlong et al. (2017b)||Mount Sinai Children’s Environmental Health Cohort assessed 162 mother–infant pairs during pregnancy in 1998–2001 in New York City.||Maternal urine samples in third trimester of pregnancy analyzed for PBA, TDCCA, and CDCCA. Children were assessed for neurodevelopment at 4, 6, and 7–9 years of age using parent reports on the BASC and BRIEF.||Detectable levels of PBA during pregnancy were associated with worse scores on internalizing (β= –4.50, 95% CI –8.05– –0.95), depression (β= –3.21, 95% CI –6.38– –0.05), somatization (β= –3.22, 95% CI –6.38– –0.06), behavioral regulation (β= –3.59, 95% CI –6.97– –0.21), emotional control (β= –3.35, 95% CI –6.58– –0.12), shifting (β= –3.42, 95% CI –6.73– –0.11), and monitoring (β= –4.08, 95% CI –7.07– –1.08) scales. Detectable levels of CDCCA were associated with worse scores on externalizing (β= –4.74, 95% CI –9.37– –0.10), conduct problems (β= –5.35, 95% CI –9.90– –0.81), behavioral regulation (β= –6.42, 95% CI –11.39– –1.45), and inhibitory control (β= –7.20, 95% CI –12.00– –2.39). TDCCA was not associated with adverse performance for any composite or subscale except for better performance on the BASC anxiety subscale (β=4.61, 95% CI 0.05–9.16). No effects of sex or race at 0.10 level.|
|Hisada et al. (2017)||See Zhang et al. (2014b) above.||Maternal urine analyzed for PBA; infant blood for TSH. Mothers completed Kinder Infant Development Scale, Index of Child Care Environment.||Infants whose mother had higher pyrethroid exposure during pregnancy had a higher development quotient score, and infants whose Index of Child Care Environment scores were higher had a higher development quotient. A positive canonical correlation coefficient (0.216, p=0.03) was found, indicating that a higher subscale score on “social relationships with children” was obtained for children whose mother had higher pyrethroid exposure during pregnancy.|
|Xue et al. (2013)||Prenatal urine of 497 pregnant women from Henan province, China, was collected in July–December 2010. Infants assessed at 1 year of age.||Mother answered questions about chemical exposure during pregnancy.||Increasing levels of pyrethroid metabolites in the urine of mothers was associated with a gradual downward trend (p=0.47) in the development quotient of the 1-year infants.|
|Watkins et al. (2016)||187 pregnant women enrolled between 1997 and 2001 into the Early Life Exposures in Mexico to Environmental Toxicants study, a longitudinal cohort study in Mexico City; subset of 21 women. Children assessed for developmental functioning at 24 and 36 months of age.||Measured maternal urinary PBA concentrations during each trimester of pregnancy. Used BSID-IIS to assess developmental functioning of infants and children using the MDI and PDI.||Lower MDI scores at 24 months were seen in the medium and high PBA categories compared with those with nondetectable urinary PBA (–3.5 and –3.8, respectively); p-value for trend=0.07). Findings were slightly stronger among girls, with MDI-24 scores among those in the medium and high PBA groups 6 points below scores of girls in the nondetectable category (95% CI= –12.3– –0.14 and –12.4–1.3, respectively; p-value for trend=0.08). At 36 months there was no statistical difference in scores among children in the high PBA category versus children with undetectable levels of PBA (p-value for trend=0.14). No associations between maternal PBA levels during the third trimester of pregnancy and PDI scores at 24 or 36 months of age among all participants or in sex-stratified analyses.|
|Eskenazi et al. (2018)||708 mother–child pairs living in Venda district of South Africa as part of the Venda Health Examination of Mothers, Babies and Their Environment (VHEMBE) longitudinal birth cohort assessed in 2012–2013.||Maternal urine and blood samples collected just before or after delivery and analyzed for PBA, CDCCA, DBCA, and TDCCA. Neurodevelopment assessed in children at 1 and/or 2 years of age using 7 or 8 scales of BSID-III.||At 1 year of age, each 10-fold increase in CDCCA, TDCCA, and PBA was significantly associated with a decrement in Social-Emotional scores (β= –0.63, 95% CI –1.14– –0.12; β= –0.48, 95% CI –0.92– –0.05; and β= –0.58, 95% CI –1.11– –0.06, respectively). At 2 years of age, each 10-fold increase in maternal DBCA levels was associated with significant decrements in Language Composite scores (β= –1.74, 95% CI –3.34– –0.13) and Expressive Communication scores (β= –0.40, 95% CI –0.77– –0.04). Significant differences by sex were estimated for pyrethroid metabolites and motor function scores at 2 years of age, with higher scores for boys and lower scores for girls.|
TABLE 5-5 Continued
|Viel et al. (2015)||PELAGIE cohort in France. 3,421 pregnant women from Brittany, France, were recruited January 2002 to February 2006 before the 19th week of gestation.||Urine samples were collected during early pregnancy (6–19 gestational weeks) for mothers and at the visit at 6 years of age for children. 287 children received cognitive assessments with the WISC at age 6 as did mothers. Urine analyzed for pyrethroid metabolites PBA, 4-F-PBA, CDCCA, TDCCA and DBCA.||Assessed association between maternal prenatal pyrethroid metabolite concentrations and neurocognitive scores, adjusting for potential confounders of the associations, urinary creatinine levels, DMP and DEP prenatal concentrations, and the corresponding childhood pyrethroid metabolite concentrations. No consistent association was found, although some evidence suggested a potentially negative trend between TDCCA concentrations and WISC-Working memory index (p-value for trend=0.18).|
|Viel et al. (2017)||See Viel et al. (2015), PELAGIE cohort.||The children’s behavior was further evaluated for three subscales of the Strengths and Difficulties Questionnaire: prosocial behavior, internalizing disorders, and externalizing disorders.||Only increasing concentrations of CDCCA in prenatal urine were associated with any abnormal or borderline scores on the Strengths and Difficulties Questionnaire and only for internalizing disorders (p=0.05).|
|Shelton et al. (2014)||Childhood Autism Risks from Genetics and Environment study is a population-based case-control study of 970 participants with ASD (n=486) or developmental disorder (n=168) with typically developing referents (n=316).||Commercial pesticide application data from the California Pesticide Use Report (1997–2008) were linked to the addresses during pregnancy. One-third of mothers lived within 1.5 km of an agricultural pesticide application.||Prenatal proximity (within 1.5 km) to pyrethroids was associated with an increased risk for ASD if the exposure was prior to conception (OR=1.82, 95% CI 1.00–3.31) or in the third trimester (OR=1.87, 95% CI 1.02–3.43). Exposure at 1.75 km was associated with an increased risk for ASD only in the third trimester (OR=1.83, 95% CI 1.04–3.23). For developmental disorder and residential proximity to agricultural pesticide application, there was a significant risk only for a distance of 1.75 km during the third trimester (OR=2.34, 95% CI 1.18–4.67).|
|Other Developmental Effects|
|Carmichael et al. (2014)||569 cases of children born during 1997–2006 with 1 of 8 congenital heart defects in 8 counties in San Joaquin Valley, CA, and 785 infants born without birth defects. Heart defects (heterotaxia; tetralogy of Fallot; hypoplastic left heart syndrome; coarctation of the aorta; pulmonary valve stenosis; ventricular septal defect, perimembranous; and atrial septal defect, secundum) based on California Birth Defects Monitoring Program.||Pesticide exposure was estimated for each mother on the basis of her residence from 3 months prior to conception to delivery using geospatial coding for eight San Joaquin counties and the Pesticide Use Reporting records for pesticides applied at greater than 100 lbs in any of the eight counties.||Prenatal exposure to permethrin was not significantly associated with any heart defect and only the ventricular septal defect, perimembranous showed a nonsignificant increase (OR=2.2, 95% CI 0.8–5.7).|
|Monge et al. (2007)||N=334 subjects with cases of leukemia ages 0–14 years old. Diagnosed in Costa Rica during 1995–2000. Population controls (N=579).||Parents were interviewed about risk factors for childhood leukemia, and their exposures were rated as to hazard value and by time of occurrence. 22 pesticides were identified, including deltamethrin.||Neither maternal nor paternal exposure to deltamethrin at any time prior to birth was significantly associated with childhood leukemia for either boys or girls.|
|Liu et al. (2012)||CCCEH birth cohort. Children 5–6 years of age.||PBO and permethrins were measured in personal air during the third trimester of pregnancy. Air samples collected for 2 consecutive days. Mothers reported on their children’s coughs.||Prenatal exposure to cis- and trans-permethrin was not significantly associated with noninfectious cough at age 5–6 years; however, there was a positive association between noninfectious cough and prenatal PBO and noninfectious cough in children ages 5–6 years (OR=1.27 for unit change in log PBO, 95% CI 1.09–1.48, p<0.01). Exposure to PBO or permethrins at ages 5–6 years was not associated with an increase in childhood cough.|
TABLE 5-5 Continued
|Coker et al. (2018)||708 mother–child pairs in VHEMBE longitudinal birth cohort.||Maternal urine and blood samples collected just before or after delivery and analyzed for PBA, DCCA, DBCA, and TDCCA. Anthropometric measures of children taken at least at one or both follow-up visits around age 1 or 2 years old.||DBCA and TDCCA were inversely related to BMI-for-age and weight-for-height overall; however, results suggested that weight-for-age and weight-for-height associations for TDCCA (sex interaction p-value weight-for-age=0.02; p-value weight-for-height=0.13) and CDCCA (sex interaction p-value weight for-age=0.02; p-value weight-for-height=0.08) were strongest and most consistent in boys relative to girls. Only boys had significant decreases in BMI-for-age (p=0.05) and weight-for-height (p=0.05) for DBCA exposure, and for weight-for-age (p=0.05), BMI-for-age (p=0.04), and weight-for-height (p=0.03) for exposure to TDCCA. None of the other exposures or measures were significantly related.|
NOTE: ASD=autism spectrum disorder; BASC=Behavior Assessment System for Children; BMI=body mass index; BSID=Bayley Scales of Infant Development II; BRIEF=Behavior Rating Inventory of Executive Functioning; CCCEH=Columbia Center for Children’s Environmental Health study; CDCCA=cis-3-2,2dichlorovinyl)-2,2-dimethylcyclopropane carboxylic acid; CHAMACOS=Center for the Health Assessment of Mothers and Children of Salinas study; CI=confidence interval; DBCA=cis-2,2-dibromovinyl-2,2-dimethylcyclopropane-1-carboxylic acid; DEP=diethyl phosphate; DFI=DNA fragmentation index; DMP=dimethyl phosphate; DNA=deoxyribonucleic acid; FSH=follicle-stimulating hormone; FT4=free thyroxine; IQ=intelligence quotient; IRR=incidence rate ratio; LH=luteinizing hormone; LOD=level of detection; MDI=mental development index; OP=organophosphate; OR=odds ratio; PBA=3-phenoxybenzoic acid; PBO=piperonyl butoxide; PDI=psychomotor development index; SHBG=sex hormone binding globulin; TBG=thyroxin-binding globulin; TDCCA=trans-3-2,2-dichlorovinyl)-2,2-dimethylcyclopropane carboxylic acid; TSH=thyroid-stimulating hormone; TTP=time to pregnancy; VHEMBE=Venda Health Examination of Mothers, Babies and Their Environment; WHO=World Health Organization; WISC-IV=Wechsler Intelligence Scale for Children, 4th edition.
Lindane is the only organochlorine pesticide considered by the committee in this report. It is the γ-isomer of 1,2,3,4,5,6-hexachlorocyclohexane (HCH), a persistent organochlorine pesticide commonly used to treat ectoparasites (e.g., scabies and lice). It is available as a 1% shampoo and lotion to topically treat lice infestations on humans. The Food and Drug Administration recommends that lindane be used only when other treatments have not worked or if the person cannot use other treatments (FDA, 2018). Service members are most likely to be exposed to lindane during deployment as a result of its use as an ectoparasiticide and ovicide. Fricker et al. (2000) reported that lindane was used to delouse prisoners of war in the 1990–1991 Gulf War, but U.S. service members were also issued small containers of lindane as a powder for personal use, and several soldiers indicated they used the powder more frequently than directed. Lindane was also reported to be used to treat uniforms (VA, 2017). The Volume 11 committee was unable to identify any information on the extent of lindane use in the Post-9/11 conflicts.
Lindane is a central nervous system stimulant and may cause neurologic effects, including seizures, even when used correctly. Chlorophenols are the primary urinary metabolites of lindane, with 2,3,5-, 2,4,6-, and 2,4,5-trichlorophenol accounting for more than 50% of the metabolites (ATSDR, 2005). Chlorophenols have also been detected in blood in animal studies following lindane exposure (ATSDR, 2005).
The reproductive and developmental effects of lindane are summarized in Table 5-6 at the end of this section and in the following section.
Reproductive Effects in Men and Women
The Volume 2 committee (IOM, 2003) and ATSDR (2005) both identified only one study that assessed lindane’s reproductive effects in men—an examination of workers at a lindane production facility for 8 years (Tomczak et al., 1981). Blood samples from the exposed workers showed elevated serum LH concentrations, somewhat elevated levels of FSH, and somewhat depressed testosterone levels.
The Volume 11 committee identified three studies on female reproductive effects from exposure to lindane. In a study of 94 couples at a fertility clinic in Egypt, follicular fluid from each woman was collected during in vitro fertilization–embryo transfer and analyzed for the presence of lindane (mean=418.6±171.4 [sd] μg/L) among other pesticides and polychlorinated biphenyls (Al-Hussaini et al., 2018). There were negative correlations between the follicular fluid concentration of lindane and endometrial thickness (p=0.0001), and a high concentration of lindane was associated with a lower implantation rate (p=0.006). Lindane did not reduce fertilization or early embryo cleavage rates.
In an earlier study of the association between exposure to lindane and endometriosis, Buck Louis et al. (2012) assessed 473 women 18–44 years of age scheduled for a laparoscopy or laparotomy (operative cohort) and an age and residence matched population cohort of 127 women in San Francisco, California, between 2007 and 2009. Serum and fat samples were analyzed for HCH and its β and γ isomers along with other persistent organic pollutants. Lindane in the fat of the operative cohort was significantly associated with endometriosis (OR=1.27, 95% CI 1.01–1.59). Serum β-HCH was significantly associated with endometriosis in the population cohort (OR=1.72, 95% CI 1.09–2.72), but not the operative cohort; fat samples were not analyzed for the population cohort, only sera. In another study in western Washington State, conducted from 1996 to 2001, Upson et al. (2013) compared levels of several endocrine-disrupting chemicals in the sera of 248 women with endometriosis and 538 population-based women. β-HCH—but not γ-HCH (lindane)—was associated with endometriosis
(third versus lowest quartile OR=1.7, 95% CI 1.0–2.8; highest versus lowest quartile OR=1.3, 95% CI 0.8–2.4), and the risk was even greater for ovarian endometriosis specifically (highest versus lowest quartile OR=2.5, 95% CI 1.1–5.3). A third study of 55 women in France with endometriosis (24 with deep infiltrating endometriosis only and 25 with deep infiltrating endometriosis and ovarian endometriosis) and 44 controls was conducted between 2013 and 2015 (Ploteau et al., 2017). Concentrations of β-HCH in adipose tissue collected during surgery were not associated with an increased risk of deep infiltrating endometriosis (OR=1.58, 95% CI 0.94–2.80), but there was a significant association with deep infiltrating endometriosis combined with ovarian endometriosis (OR=3.64, 95% CI 1.52–11.15). The researchers analyzed for γ-HCH in the fat tissue, but the detection rate was so low that further analysis was not performed (Ploteau et al., 2016).
Adverse Pregnancy Outcomes
Volume 2 discussed two studies that examined the effects of lindane on fetal outcomes; however, lindane was not a unique exposure in either study. Willis et al. (1993) found that among 535 pregnant women who resided in an agricultural area in San Diego, California, only those who were not exposed to pesticides—which included carbaryl and lindane—had spontaneous abortion or preterm delivery; furthermore, the unexposed women also had a greater incidence of low-birth-weight infants. The Volume 2 committee found this study hard to interpret because critical information such as the number of exposed women in the cohort was not presented.
The second study assessed paternal insecticide exposure and pregnancy outcomes among 32 Italian pesticide applicators and their wives (Petrelli et al., 2000). Although there was an increased risk of spontaneous abortion among the exposed wives (OR=3.8, 95% CI 1.2–12.0), there was little information on the selection and recruitment of study subjects, and a crude exposure measurement was used.
The 2005 ATSDR review Toxicological Profile for Alpha-, Beta-, Gamma-, and Delta-Hexachlorocyclohexane had no reports on fetal outcomes following inhalation exposure in humans. One study found that babies born to pregnant Indian women with higher blood levels of HCH isomers, including the gamma isomer, also had higher levels of lindane and were more likely to have intrauterine fetal growth retardation (Siddiqui et al., 2003); however, mothers also had exposure to other organochlorine pesticides.
The Volume 11 committee identified one new study that examined the effects of prenatal exposure to organochlorine pesticides in infants born in an urban area of Northern India (Tyagi et al., 2015). The levels of γ-HCH in maternal blood and placental tissue were measured in 50 women who were at less than 37 weeks of gestation and 50 women who were at 37 weeks or more; γ-HCH residues were detected in the blood of all the women. There was a negative correlation between the blood levels of γ-HCH and placental weight (r= –0.401, p=0.006) but no correlation with the weight of the infant or length of gestation.
The Volume 2 committee found mixed results from studies of the reproductive effects of lindane in animals. Lindane was found to produce reversible decreases in the total number of sperm in mice (Smith, 1991), testicular degeneration in rats (Chowdhury et al., 1987), and reduced “reproductive behavior” in young rams (Beard et al., 1999). Lindane also had adverse effects on reproduction in female animals. Low doses (10 mg/kg/day for 138 days) of lindane reduced fecundity and litter size in rats (Trifonova et al., 1970); increased the number of resorbed fetuses in hamsters, rabbits, and rats (Dzierzawski, 1977);
and increased the duration of diestrus, shortened the duration of estrus, lengthened the gestation period, decreased the number of fetuses, increased the number of dead fetuses, and decreased the growth of the young in rats (Nayshteyn and Leybovich, 1971). However, another study found lindane to have no effect on the numbers of pregnancies, abortions, and dead or resorbed fetuses when given to rats on GDs 6-15 (Khera et al., 1979).
The 2005 ATSDR review Toxicological Profile for Alpha-, Beta-, Gamma-, and Delta-Hexachlorocyclohexane found no studies on reproductive effects in animals following inhalation exposure to any HCH isomer. Anti-estrogenic properties were found in female rats given γ-HCH by the oral route (Chadwick et al., 1988), and female rabbits treated orally with γ-HCH had a reduced ovulation rate (Lindenau et al., 1994). The results of single and multigeneration reproduction studies in rats and mink indicate that exposure to γ-HCH reduces receptivity to mating and reduces whelping rate, decreases numbers of offspring at birth, reduces neonatal viability, and delays the maturation of pups. Those effects were primarily the result of prenatal or postnatal developmental toxicity (Beard and Rawlings, 1998; Beard et al., 1997; King, 1991).
The Volume 11 committee identified five new animal studies on the reproductive effects of lindane. In a two-generation study in rats, lindane administered in the diet to male and females for 10 weeks through mating was found to have no reproductive effects in males; in females, a lack of maternal behavior was seen in the F0 and F1 generations, but no effects were observed on mating, fertility, pregnancy, or parturition in F1 females (Matsuura et al., 2005).
Oral administration of lindane to male rats at 5 mg/kg for 30 days resulted in significantly reduced weights of the reproductive organs and in reductions of testicular DNA, RNA, protein, and enzymes (Sujatha et al., 2001). Damage to Leydig cells was seen in rats treated dermally with a single 1% dose of lindane (the medical dose for scabies) (Suwalsky et al., 2000).
Two in vitro animal studies were considered by the committee. Walsh and Stocco (2000) found that alpha-, delta-, and gamma-HCH inhibited steroidogenesis by reducing StAR protein expression, an action that may contribute to the pathogenesis of lindane-induced reproductive dysfunction. Ronco et al. (2001) also found that lindane produced a dose-dependent inhibition of testosterone production in hCG-stimulated Leydig cells.
The Volume 2 committee found no studies of developmental effects in humans following either maternal or paternal exposure. Furthermore, the 2005 ATSDR review Toxicological Profile for Alpha-, Beta-, Gamma-, and Delta-Hexachlorocyclohexane also found no studies of developmental effects of lindane in humans following either maternal or paternal exposure by any route of exposure.
The Volume 11 committee identified only one new study of developmental effects in infants following maternal exposure to lindane. Lindane levels in placentas were analyzed in infants with and without neural tube defects who had been born to 130 women in an agricultural area of China. Higher levels of γ-HCH in the placenta were associated with an increased risk for any neural tube defects (OR=3.36, 95% CI 1.10–6.44) (Ren et al., 2011).
In Volume 2, a number of studies indicated that lindane is not teratogenic. A three-generation study of rats (Palmer et al., 1978) and experiments with mice (Chernoff and Kavlock, 1983; Gray and Kavlock, 1984) did not demonstrate any teratogenic effects.
The 2005 ATSDR review Toxicological Profile for Alpha-, Beta-, Gamma-, and Delta-Hexachlorocyclohexane reported that there were no inhalation data on the developmental effects of lindane in animals. Adverse effects on testicular histology and sperm numbers occurred in the adult male offspring of mice that were exposed to γ-HCH during gestation (Traina et al., 2003).
No adverse prenatal developmental effects of γ-HCH from oral exposure have been found in rats or rabbits (Khera et al., 1979; Palmer et al., 1978; Seiler et al., 1994). Decreases in fetal weight, fetal thymic weight, and placental weight were reported in mice exposed to a single oral dose of γ-HCH on day 12 of gestation (Hassoun and Stohs, 1996). No effects on embryonic development were seen in rabbits treated orally with γ-HCH (Seiler et al., 1994). Due to the lack of developmental toxicity studies in humans as well as the lack of inhalation and dermal data in animals, insufficient information is available to indicate whether HCH affects development via all three routes of exposure. Pharmacokinetic data suggest that HCH isomers might have the potential to affect development across all routes of exposure.
The Volume 11 committee identified three studies of oral exposure to lindane in animals. In a series of experiments, pregnant mice were treated on GDs 9 to 16 with 15 or 25 mg/kg lindane, and male and female offspring were assessed for developmental effects. There were no significant effects on gross malformations in pups, the average number of live pups per litter, the incidence of stillbirths or runts, sex ratio, survival to weaning, or developmental landmarks. Adult F1 males had changes in testicular enzyme activity, germ cell distribution, and chromatic abnormalities of the sperm at PND 60 (i.e., sexual maturity) that were not significant by PND 100 (Di Consiglio et al., 2009; Traina et al., 2003). Female F1 mice showed early increases in uterus weight that were not evident at PND 100; only the diameter of follicular oocytes was significantly reduced in treated females (Maranghi et al., 2007). Preimplantation embryotoxicity with degenerating two-cell embryos was induced by exposure of preovulatory oocytes in mice at the highest lindane dose tested (25 mg/kg); no other signs of embryotoxicity were observed (e.g., apoptosis or micronucleus induction). In one of the few studies to assess preconception exposure only, female mice were given up to 64.4 mg/kg/day lindane in drinking water for 2 weeks prior to mating. There were no cytopathological effects on the male offspring germ cells (Lopez-Casas et al., 2012).
Synthesis and Conclusions
There is little epidemiologic information available on lindane. Both the Volume 2 committee and ATSDR found little information on the effects of lindane exposure on male reproduction in humans and none for lindane’s effects on female reproduction. Three studies of endometriosis found that serum levels of β-HCH were associated with an increased risk of endometriosis, but only one of the studies found an increased risk specifically with lindane exposure (Buck Louis et al., 2012). Lindane was also associated with reduced endometrial thickness but not embryo implantation in one study of in vitro fertilization (Al-Hussaini et al., 2018). Thus, all three studies suggest that hexachlorocyclohexane may affect endometrial tissue but the association is stronger for the β-isomer and not the γ-isomer (lindane). Thus, the committee finds that more research on the association between the γ-isomer and female reproductive effects is warranted as lindane is still available to the general public.
The studies of the effects of lindane on birth outcomes are difficult to interpret because lindane was not the only pesticide to which the mothers were exposed. The single new study identified by the Volume 11 committee, Tyagi et al. (2015), found only a negative correlation with lindane and placental weight and no other adverse effects.
In general, the animal studies examined by ATSDR and the Volume 2 and Volume 11 committees found that oral exposure to lindane was associated with numerous reproductive outcomes in both male
and female animals, including reduced mating behavior in both sexes and poor fetal outcomes. This body of animal evidence suggests that additional studies of the effects of lindane in humans is warranted to ascertain whether effects seen in animals may be occurring in humans. Furthermore, animals studies with gestational exposures that are equivalent to human organogenesis would be helpful.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between exposure to lindane and reproductive effects in men, or with adverse pregnancy outcomes.
The Volume 11 committee also concludes that there is limited/suggestive evidence of an association between exposure to lindane and reproductive effects in women.
There is a lack of information on developmental effects in humans following prenatal exposures to lindane; only one study of women in China found an association—in this case, between the women’s lindane exposure and neural tube defects in their children (Ren et al., 2011).
Lindane has not been shown to be teratogenic in animals, and the results of studies to ascertain adverse developmental outcomes following oral exposure prenatally are mixed. Developmental effects were seen in two-generation studies of mice; adult offspring of both sexes displayed altered reproductive parameters at sexual maturity, although the effects appeared to resolve with age. ATSDR (2005) indicated that there was insufficient information to determine if lindane affects development in animals or humans. Although the animal studies indicate that lindane may be both a reproductive and a developmental toxicant, confirmation of the effects in humans is lacking. The available human studies are generally small and may be confounded by exposure to other pesticides or organochlorines; and exposure may have occurred at any point prior to or during pregnancy.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between prenatal exposure to lindane and developmental effects.
TABLE 5-6 Summary of Reproductive and Developmental Effects of Lindane
|Reproductive Effects in Men and Women|
|Al-Hussaini et al. (2018)||94 couples (women ages 20 to 38 years) at a university-affiliated fertility center in Assiut, Egypt; cross-sectional observation study.||Follicular fluid samples collected during in vitro fertilization–embryo transfer and analyzed for pesticides, including lindane.||Higher lindane concentrations were significantly associated with lower implantation rate (p=0.006) and reduced endometrial thickness (p<0.01) but there was no effect on the number of eggs retrieved, number of fertilized oocytes, or number of cleaved embryos.|
|Buck Louis et al. (2012)||473 menstruating women 18–44 years of age scheduled for a laparoscopy or laparotomy (190 with a diagnosis of endometriosis and 293 without) and an age and residence matched population cohort of 127 women (14 with endometriosis and 113 without).||Blood and urine collected from all participants and omental (abdominal) fat collected from operative cohort only.||Lindane levels in the fat of women in the operative cohort were significantly associated with endometriosis (OR=1.27, 95% CI 1.01–1.59). β-HCH in the sera of women in the population cohort was significantly associated with endometriosis (OR=1.72, 95% CI 1.09–2.72) but not in the operative cohort (OR=0.77, 95% CI 0.54–1.14).|
|Upson et al. (2013)||248 women ages 18–49 years with surgically confirmed endometriosis and 538 population based controls enrolled in a large health care system in western Washington State in 1996–2001.||Organochlorine pesticides were determine in serum of all participants, including β- and γ-HCH (lindane).||Serum concentrations of β-HCH were significantly associated with an increased risk of endometriosis (third versus lowest quartile OR=1.7, 95% CI 1.0–2.8; highest versus lowest quartile OR=1.3, 95% CI 0.8–2.4); the association was stronger when analysis restricted to cases of ovarian endometriosis (third versus lowest quartile OR=2.5, 95% CI 1.5–5.2; highest versus lowest quartile OR=2.5, 95% CI 1.1–5.3). γ-HCH was not significantly associated with any endometriosis or ovarian endometriosis.|
|Ploteau et al. (2017)||55 women, ages 18–45, in France with endometriosis (24 with deep infiltrating endometriosis only and 25 with deep infiltrating endometriosis and ovarian endometriosis) and 44 controls; assessed between 2013 and 2015.||Subcutaneous or abdominal adipose tissue taken during surgery was analyzed for 76 persistent organic pollutants, including the α-, β-, and γ-isomers of HCH.||Concentrations of β-HCH in adipose tissue were not associated with an increased risk of deep infiltrating endometriosis (OR=1.58, 95% CI 0.94–2.80), but there was a significant association with deep infiltrating endometriosis combined with ovarian endometriosis (OR=3.64, 95% CI 1.52–11.15). Low detection of γ-HCH in the adipose tissue precluded further analysis.|
|Adverse Pregnancy Outcomes|
|Tyagi et al. (2015)||50 women who had below 37 weeks of gestation and 50 women who had a gestation period of 37 weeks or above term labor in Delhi, India, between November 2011 and April 2013.||Maternal blood collected at time of labor and placenta at delivery. Analyzed for residues of organochlorine pesticides, including lindane.||All samples contained lindane residue. Increased levels of γ-HCH were associated with reduced placental weight (r= –0.401, p=0.006). Lindane was not significantly associated with reduced birth weight or length of gestation.|
|Ren et al. (2011)||130 women in a rural area of China; 80 women who had pregnancies with neural tube defects (cases) and 50 women who delivered healthy infants (controls).||Placental samples of cases and controls analyzed for persistent organic pollutants; concentrations of lindane were significantly higher in case than in control placentas.||Higher levels of γ-HCH were associated with increased risk for any NTDs (OR=3.36, 95% CI 1.10–6.44). Median concentration γ-HCH was higher in the placental samples of cases with anencephaly and spina bifida than those of controls, although differences in the concentration of γ-HCH for the two subtypes of NTDs were not statistically significant, and there was no apparent dose–response relationship between levels of γ-HCH and the risk of any subtype of NTDs.|
NOTE: β-HCH=beta-hexachlorocyclohexane; γ-HCH=gamma-hexachlorocyclohexane; CI=confidence interval; NTD=neural tube defects; OR=odds ratio.
DEET is a widely used liquid insect repellant. It is the active ingredient in a number of commercially available products such as Off® and Cutter® insect repellants. DEET may be present in these repellants at a concentration of up to 30% and, according to the label, may be used on children as young as 2 years old (ATSDR, 2017).
During the 1990–1991 Gulf War, DoD shipped approximately two 2-oz tubes of 33% DEET per service member (PAC, 1996). It was unclear to the committee if this allotment per service members was the only supply of DEET they received for the duration of their deployment.
There were no epidemiologic or other human studies of the possible reproductive or developmental effects of DEET considered by the Volume 2 committee. One animal study found an increased permeability of the blood–brain barrier and the blood–testis barrier for male rats exposed dermally to DEET and permethrin for 60 days (Abou-Donia et al., 2001). Milder effects were reported for DEET alone, but not for permethrin alone. There were no other reports on any animal reproductive or developmental effects of DEET.
The 2017 ATSDR review Toxicological Profile for DEET (N,N-diethyl-meta-toluamide) reported that there were no reproductive studies of DEET in humans and that there were no developmental effects of DEET found in two epidemiological studies of women and their children. In animal studies, DEET was not observed to affect fertility in male or female rats in a two-generation continuous feeding study, nor did it affect reproductive organs in intermediate- or chronic-duration oral studies in animals. One 90-day study found that the dermal application of a high dose of DEET (≥624 mg/kg/day) increased the incidence of tubular degeneration in hamster testes. One relevant animal inhalation study of male and female Sprague-Dawley rats given up to 1,511 mg/m3 of aerosolized DEET for 13 weeks found no gross or microscopic alterations in the reproductive organs.
The Volume 11 committee found little new information on possible reproductive or developmental effects from exposure to DEET. Because there are only two new studies to be considered, there is no table of reproductive or developmental effects provided here.
In one study of the effect of DEET on human fertility, Segal et al. (2017) evaluated urine and semen samples from 90 men participating in a prospective cohort study at the Massachusetts General Hospital Fertility Center. Urinary concentrations of the DEET metabolite 3-(diethylcarbamoyl)benzoic acid were not associated with any semen parameters, and there were no differences in the parameters between the highest and lowest metabolite concentrations.
A review by Koren et al. (2003) examined the potential for increased risks to children and pregnant women from DEET use. The researchers found that the available evidence did not support an increased risk for adverse reproductive or developmental outcomes in mothers who used DEET during their second and third trimesters or in their children. There were no studies on first-trimester exposures. In the era of Zika virus, Wylie et al. (2016) also summarized available reproductive and developmental toxicity literature on DEET and concluded that DEET did not pose a health risk when used as directed.
Only one study was identified that assessed the level of DEET in cord blood and birth outcomes in Fuyang, China (Wickerham et al., 2012). Of the 112 women recruited at 36 weeks gestation who reported a healthy, uncomplicated singleton pregnancy, 73% had detectable levels of DEET in cord blood collected at delivery, the most of any pesticide. However, there was no association between DEET levels and a change in birth weights (β=110, 95% CI –49.2–269).
Synthesis and Conclusions
DEET has been used for more than 50 years as an effective insect repellant around the world. Unlike most pesticides, however, there is a dearth of information on its long-term toxicity, including its possible reproductive or developmental effects, in both humans or animals. Nonetheless, the few available data indicate that, when used as directed, DEET appears to pose no risks for reproductive or developmental effects in humans, although there do not appear to be any long-term studies of any potential reproductive or developmental effects in humans or animals to confirm this. Given the widespread use of this pesticide by men, women, and children, studies on its long-term reproductive and developmental effects in animals are warranted.
The Volume 11 committee concludes that there is inadequate/insufficient evidence to determine whether an association exists between exposure to DEET and reproductive or developmental effects.
Abd El-Aziz, M.I., A.M. Sahlab, and M. Abd El-Khalik. 1994. Influence of diazinon and deltamethrine on reproductive organs and fertility of male rats. Deutsche Tieräerztliche Wochenshrift 101:213–248.
Abdallah, F.B., K. Hamden, I. Galeraud-Denis, A. El Feki, and L. Keskes-Ammar. 2010. An in vitro study on reproductive toxicology of deltamethrin on rat spermatozoa. Andrologia 42(4):254–259.
Abou-Donia, M. B., L.B. Goldstein, A. Dechovskaia, S. Bullman, K.H. Jones, E.A. Herrick, A.A. Abdel-Rahman, and W.A. Khan. 2001. Effects of daily dermal application of DEET and epermethrin, alone and in combination, on sensorimotor performance, blood-brain barrier, and blood-testis barrier in rats. Journal of Toxicology and Environmental Health Part A 62(7):523–541.
Aitken, R.J. 2017. DNA damage in human spermatozoa; Important contributor to mutagenesis in the offspring. Translational Andrology and Urology 6(Suppl 4):S761–S764.
Akbarsha, M. A., P.N. Latha, and P. Murugaian. 2000. Retention of cytoplasmic droplet by rat cauda epididymal spermatozoa after treatment with cytotoxic and xenobiotic agents. Journal of Reproduction & Fertility 120(2):385–390.
Akhtar, N., M.K. Srivastava, and R.B. Raizada. 2006. Transplacental disposition and teratogenic effects of chlorpyrifos in rats. Journal of Toxicological Science 31(5):521–527.
Alaa-Eldin, E.A., D.A. El-Shafei, and N.S. Abouhashem. 2017. Individual and combined effect of chlorpyrifos and cypermethrin on reproductive system of adult male albino rats. Environmental Science and Pollution Research International 24(2):1532–1543.
Aldridge, J.E., F.J. Seidler, A. Meyer, I. Thillai, and T.A. Slotkin. 2003. Serotonergic systems targeted by developmental exposure to chlorpyrifos: Effects during different critical periods. Environmental Health Perspectives 111(14):1736–1743.
Aldridge, J.E., F.J. Seidler, and T.A. Slotkin. 2004. Developmental exposure to chlorpyrifos elicits sex-selective alterations of serotonergic synaptic function in adulthood: Critical periods and regional selectivity for effects on the serotonin transporter, receptor subtypes, and cell signaling. Environmental Health Perspectives 112(2):148–155.
Aldridge, J.E., A. Meyer, F.J. Seidler, and T.A. Slotkin. 2005. Alterations in central nervous system serotonergic and dopaminergic synaptic activity in adulthood after prenatal or neonatal chlorpyrifos exposure. Environmental Health Perspectives 113(8):1027–1031.
Al-Hamdani, N.M., and H.N. Yajurvedi. 2010. Cypermethrin reversibly alters sperm count without altering fertility in mice. Ecotoxicology and Environmental Safety 73(5):1092–1097.
Al-Hussaini, T.K., A.A. Abdelaleem, I. Elnashar, O.M. Shabaan, R. Mostafa, M.A.H. El-Baz, S.E.M. El-Deek, and T.A. Farghaly. 2018. The effect of follicullar fluid pesticides and polychlorinated biphenyls concentrations on intracytoplasmic sperm injection (ICSI) embryological and clinical outcome. European Journal of Obstetrics & Gynecology and Reproductive Biology 220:39–43.
Arbuckle, T.E., Z. Lin, and L.S. Mery. 2001. An exploratory analysis of the effect of pesticide exposure on the risk of spontaneous abortion in an Ontario farm population. Environmental Health Perspectives 109(8):851–857.
ATSDR (Agency for Toxic Substances and Disease Registry). 1997a. Toxicological profile for chlorpyrifos. Atlanta, GA: U.S. Department of Health and Human Services.
ATSDR. 1997b. Toxicological profile for dichlorvos. Atlanta, GA: U.S. Department of Health and Human Services.
ATSDR. 2003a. Toxicological profile for malathion. Atlanta, GA: U.S. Department of Health and Human Services.
ATSDR. 2003b. Toxicological profile for pyrethrins and pyrethroids. Atlanta, GA: U.S. Department of Health and Human Services.
ATSDR. 2005. Toxicological profile for alpha-, beta-, gamma-, and delta-hexachlorocyclohexane. Atlanta, GA: U.S. Department of Health and Human Services.
ATSDR. 2008. Toxicological profile for diazinon. Atlanta, GA: U.S. Department of Health and Human Services.
ATSDR. 2017. Toxicological profile for DEET (N,N-diethyl-meta-toluamide). Atlanta, GA: U.S. Department of Health and Human Services.
Beard, A.P., and N.C. Rawlings. 1998. Reproductive effects in mink (Mustela vison) exposed to the pesticides lindane, carbofuran and pentachlorophenol in a multigenerational study. Journal of Reproduction & Fertility 113:93–104.
Beard, A.P., A.C. McRae, and N.C. Rawlings. 1997. Reproductive efficiency in mink (Mustela vison) treated with the pesticides lindane, carbofuran and pentachlorophenol. Journal of Reproduction & Fertility 111:21–28.
Beard, A.P., P.M. Bartlewski, R.K. Chandolia, A. Honaramooz, and N.C. Rawlings. 1999. Reproductive and endocrine function in rams exposed to the organochlorine pesticides lindane and pentachlorophenol from conception. Journal of Reproduction & Fertility 115(2):303–314.
Bell, E., I. Hertz-Picciotto, and J.J. Beaumont. 2001. Case-cohort analysis of agricultural pesticide applications near maternal residence and selected causes of fetal death. American Journal of Epidemiology 154:702–710.
Ben Slima, A., F. Ben Abdallah, L. Keskes-Ammar, Z. Mallek, A. El Feki, and R. Gdoura. 2012. Embryonic exposure to dimethoate and/or deltamethrin impairs sexual development and programs reproductive success in adult male offspring mice. Andrologia 44(Suppl 1):661–666.
Ben Slima, A., Y. Chtourou, M. Barkallah, H. Fetoui, T. Boudawara, and R. Gdoura. 2017. Endocrine disrupting potential and reproductive dysfunction in male mice exposed to deltamethrin. Human & Experimental Toxicology 36(3):218–226.
Berkowitz, G.S., J.G. Wetmur, E. Birman-Deych, J. Obel, R.H. Lapinski, J.H. Godbold, I.R. Holzman, and M.S. Wolff. 2004. In utero pesticide exposure, maternal paraoxonase activity, and head circumference. Environmental Health Perspectives 112(3):388–391.
Bicker, W., M. Lämmerhofer, and W. Lindner. 2005. Determination of chlorpyrifos metabolites in human urine by reversed-phase/weak anion exchange liquid chromatography-electrospray ionisation-tandem mass spectrometry. Journal of Chromatography B: Analytical Technologies in the Biomedical and Life Sciences 822(1–2):160–169.
Bouchard, M.F., J. Chevrier, K.G. Harley, K. Kogut, M. Vedar, N. Calderon, C. Trujillo, C. Johnson, A. Bradman, D.B. Barr, and B. Eskenazi. 2011. Prenatal exposure to organophosphate pesticides and IQ in 7-year-old children. Environmental Health Perspectives 119(8):1189–1195.
Bradman, A., L. Quirós-Alcalá, R. Castorina, R. Aguilar Schall, J. Camacho, N.T. Holland, D.B. Barr, and B. Eskenazi. 2015. Effect of organic diet intervention on pesticide exposures in young children living in low-income urban and agricultural communities. Environmental Health Perspectives 123(10):1086–1093.
Buck Louis, G.M., Z. Chen, C.M. Peterson, M.L. Hediger, M.S. Croughan, R. Sundaram, J.B. Stanford, M.W. Varner, V.Y. Fujimoto, L.C. Giudice, A. Trumble, P.J. Parsons, and K. Kannan. 2012. Persistent lipophilic environmental chemicals and endometriosis: The ENDO Study. Environmental Health Perspectives 120:811–816.
Burke, R.D., S.W. Todd, E. Lumsden, R.J. Mullins, J. Mamczarz, W.P. Fawcett, R.P. Gullapalli, W.R. Randall, E.F.R. Pereira, and E.X. Albuquerque. 2017. Developmental neurotoxicity of the organophosphorus insecticide chlorpyrifos: From clinical findings to preclinical models and potential mechanisms. Journal of Neurochemistry 142(Suppl 2):162–177.
Buscail, C., C. Chevrier, T. Serrano, P. Fabienne, C. Monfort, S. Cordier, and J. Viel. 2015. Prenatal pesticide exposure and otitis media during early childhood in the PELAGIE mother–child cohort. Occupational and Environmental Medicine 72:837–844.
Bustos-Obregón, E., and P. González-Hormazabal. 2003. Effect of a single dose of malathion on spermatogenesis in mice. Asian Journal of Andrology 5(2):105–107.
Carmichael, S.L., W. Yang, E. Roberts, S.E. Kegley, A.M. Padula, P.B. English, E.J. Lammer, and G.M. Shaw. 2014. Residential agricultural pesticide exposures and risk of selected congenital heart defects among offspring in the San Joaquin Valley of California. Environmental Research 135:133–138.
Carmichael, S.L., W. Yang, E. Roberts, S.E. Kegley, T.J. Brown, P.B. English, E.J. Lammer, and G.M. Shaw. 2016. Residential agricultural pesticide exposures and risks of selected birth defects among offspring in the San Joaquin Valley of California. Birth Defects Research (Part A) 106:27–35.
Casida, J.E., D.W. Gammon, A.H. Glickman, and L.J. Lawrence. 1983. Mechanisms of selective action of pyrethroid insecticides. Annual Review of Pharmacology and Toxicology 23:413–438.
Cecchine, G., B.A. Golomb, L.H. Hilborne, D. Spektor, and C.R. Anthony. 2000. A review of the scientific literature as it pertains to Gulf War illnesses: Pesticides. Volume 8. Arlington, VA: National Defense Research Institute (RAND).
Chadwick, R.W., R.L. Cooper, J. Chang, G.L. Rehnberg, and W.K. McElroy. 1988. Possible antiestrogenic activity of lindane in female rats. Journal of Biochemistry and Toxicology 3:147–158.
Chauhan, L.K., M. Varshney, V. Pandey, P. Sharma, V.K. Verma, P. Kumar, and S.K. Goel. 2016. ROS-dependent genotoxicity, cell cycle perturbations and apoptosis in mouse bone marrow cells exposed to formulated mixture of cypermethrin and chlorpyrifos. Mutagenesis. 31(6):635–642.
Chebab, S., F. Mekircha, and E. Leghouchi. 2017. Potential protective effect of pistacia lentiscus oil against chlorpyrifos-induced hormonal changes and oxidative damage in ovaries and thyroid of female rats. Biomedicine & Pharmacotherapy 96:1310–1316.
Chen, X.P., Y.S. Chao, W.Z. Chen, and J.Y. Dong. 2017. Mother gestational exposure to organophosphorus pesticide induces neuron and glia loss in daughter adult brain. Journal of Environmental Science and Health, Part B 52(2):77–83.
Chernoff, N., and R.J. Kavlock. 1983. A teratology test system which utilizes postnatal growth and viability in the mouse. Environmental Science Research 27:417–427.
Choudhary, N., R. Goyal, and S.C. Joshi. 2008. Effect of malathion on reproductive system of male rats. Journal of Environmental Biology 29(2):259–262.
Chowdhury, A.R., H. Venkatakrishna-Bhatt, and A.K. Guatam. 1987. Testicular changes of rats under lindane treatment. Bulletin of Environmental Contamination and Toxicology 38:154–156.
Coker, E., J. Chevrier, S. Rauch, A. Bradman, M. Obida, M. Crause, R. Bornman, and B. Eskenazi. 2018. Association between prenatal exposure to multiple insecticides and child body weight and body composition in the VHEMBE South African birth cohort. Environment International 113C:122–132.
Coleman, R. E., D.A. Burkett, J.L. Putnam, V. Sherwood, J.B. Caci, B.T. Jennings, L.P. Hochberg, S.L. Spradling, E.D. Rowton, K. Blount, J. Ploch, G. Hopkins, J.W. Raymond, M.L. O’Guinn, J.S. Lee, and P.J. Weina. 2006. Impact of phlebotomine sand flies on U.S. Military operations at Tallil Air Base, Iraq: 1. Background, military situation, and development of a “leishmaniasis control program.” Journal of Medical Entomology 43(4):647–662.
Collins, T.F.X., W.H. Hansen, and H.V. Keeler. 1971. The effect of carbaryl (Sevin) on reproduction of the rat and gerbil. Toxicology and Applied Pharmacology 19(2):202–216.
Curtis, K.M., D.A, Savitz, C.R. Weinberg, and T.E. Arbuckle. 1999. The effect of pesticide exposure on time to pregnancy. Epidemiology 10(2):112–117.
Dabrowski, S., W. Hanke, K. Polanska, T. Makowiec-Dabrowska, and W. Sobala. 2003. Pesticide exposure and birthweight: An epidemiological study in Central Poland. International Journal of Occupational Medicine and Environmental Health 16(1):31–39.
Dahamna, S., F. Bencheikh, D. Harzallah, S. Boussahel, A. Belgeit, M. Merghem, and H. Bouriche. 2010. Cypermetherin toxic effects on spermatogenesis and male mouse reproductive organs. Communications in Agricultural and Applied Biological Sciences 75(2):209–216.
Dean, J.B., and D. Blair. 1976. Dominant lethal assay in female mice after oral dosing with dichlorvos or exposure to atmospheres containing dichlorovos. Mutation Research 40(1):67–72.
Declerck, K., S. Remy, C. Wohlfahrt-Veje, K.M. Main, G. Van Camp, G. Schoeters, W. Vanden Berghe, and H.R. Andersen. 2017. Interaction between prenatal pesticide exposure and a common polymorphism in the PON1 gene on DNA methylation in genes associated with cardio-metabolic disease risk—an exploratory study. Clinical Epigenetics 9:35.
Di Consiglio, E., G. De Angelis, M.E. Traina, E. Urbani, and E. Testai. 2009. Effect of lindane on CYP-mediated steroid hormone metabolism in male mice following in utero exposure. Journal of Applied Toxicology 29(8):648–655.
Ding, G., C. Cui, L. Chen, Y. Gao, Y. Zhou, R. Shi, and Y. Tian. 2015. Prenatal exposure to pyrethroid insecticides and birth outcomes in rural northern China. Journal of Exposure Science and Environmental Epidemiology 25:264–270.
Dirican, E. K., and Y. Kalender. 2012. Dichlorvos-induced testicular toxicity in male rats and the protective role of vitamins C and E. Experimental and Toxicologic Pathology 64(7–8):821–830.
Donauer, S., M. Altaye, Y. Xu, H. Sucharew, P. Succop, A.M. Calafat, J.C. Khoury, B. Lanphear, and K. Yolton. 2016. An observational study to evaluate associations between low-level gestational exposure to organophosphate pesticides and cognition during early childhood. American Journal of Epidemiology 184(5):410–418.
Dutta, A.L., and C.R. Sahu. 2013. Protective effect of emblica officinalis on chlorpyrifos (an organophosphate insecticide) induced male reproductive system in rats. International Journal of Pharma and Bio Sciences 4(1):B49–B58.
Dzierzawski, A. 1977. Embryotoxicity studies of lindane in the golden hamster, rat and rabbit. Bulletin of the Veterinary Institute in Pulawy 21:85–93.
Ecobichon, D.J. 2001. Pesticide use in developing countries. Toxicology 160(1–3):27–33.
Elbetieha, A., S. I. Da’as, W. Khamas, and H. Darmani. 2001. Evaluation of the toxic potentials of cypermethrin pesticide on some reproductive and fertility parameters in the male rats. Archives of Environmental Contamination and Toxicology 41(4):522–528.
Elmazoudy, R.H., A.A. Attia, and H.S. Abdelgawad. 2011. Evaluation of developmental toxicity induced by anticholinesterase insecticide, diazinon in female rats. Birth Defects Research. Part B, Developmental and Reproductive Toxicology 92(6):534–542.
Elsharkawy, E.E., D. Yahia, and N.A. El-Nisr. 2014. Chlorpyrifos induced testicular damage in rats: Ameliorative effect of glutathione antioxidant. Environmental Toxicology 29(9):1011–1019.
Engel, S.M., G.S. Berkowitz, D.B. Barr, S.L. Teitelbaum, J. Siskind, S.J. Meisel, J.G. Wetmur, and M.S. Wolff. 2007. Prenatal organophosphate metabolite and organochlorine levels and performance on the Brazelton Neonatal Behavioral Assessment Scale in a multiethnic pregnancy cohort. American Journal of Epidemiology 165(12):1397–1404.
Engel, S.M., J. Wetmur, J. Chen, C. Zhu, D.B. Barr, R. L. Canfield, and M.S. Wolff. 2011. Prenatal exposure to organophosphates, paraoxonase 1, and cognitive development in childhood. Environmental Health Perspectives 119(8):1182–1188.
Engel, S.M., A. Bradman, M.S. Wolff, V.A. Rauh, K.G. Harley, J.H. Yang, L.A. Hoepner, D.B. Barr, K. Yolton, M.G. Vedar, Y. Xu, R.W. Hornung, J.G. Wetmur, J. Chen, N.T. Holland, F.P. Perera, R.M. Whyatt, B.P. Lanphear, and B. Eskenazi. 2016. Prenatal organophosphorus pesticide exposure and child neurodevelopment at 24 months: An analysis of four birth cohorts. Environmental Health Perspectives 124(6):822–830.
EPA (U.S. Environmental Protection Agency). 1983. Tolerances in food administered by EPA: Pemethrin. Federal Register 48:36246–36247.
EPA. 1987. Integrated Risk Information System (IRIS) Chemical Assessment Summary: Methomyl; CASRN 16752-77-5.https://cfpub.epa.gov/ncea/iris/iris_documents/documents/subst/0069_summary.pdf (accessed June 15, 2018).
EPA 1992. Baygon; CASRN 114-26-1. Integrated Risk Information System (IRIS) Chemical Assessment Summary. National Center for Environmental Assessment.
EPA. 1994. Memorandum: Resmethrin: Evaluation of 2-generation reproduction study in rat and reevaluation of the following previously submitted studies: 3-generation reproduction study in rat, six month dog, rat. Toxicological Review 011329. Washington, DC: EPA, Office of Pesticides and Toxic Substances.
EPA. 2000. Propoxur (Baygon). https://www.epa.gov/sites/production/files/2016-09/documents/propoxur.pdf (accessed August 16, 2018).
EPA. 2016. Chlorpyrifos: Revised human health risk assessment for registration review. Memorandum. Washington, DC: EPA Office of Chemical Safety and Pollution Prevention. November 3, 2016.
Eskenazi, B., K. Harley, A. Bradman, E. Weltzien, N.P. Jewell, D.B. Barr, C.E. Furlong, and N.T. Holland. 2004. Association of in utero organophosphate pesticide exposure and fetal growth and length of gestation in an agricultural population. Environmental Health Perspectives 112(10):1116–1124.
Eskenazi, B., A.R. Marks, A. Bradman, K. Harley, D.B. Barr, C. Johnson, N. Morga, and N.P. Jewell. 2007. Organophosphate pesticide exposure and neurodevelopment in young Mexican-American children. Environmental Health Perspectives 115(5):792–798.
Eskenazi, B., K. Kogut, K. Huen, K.G. Harley, M. Bouchard, A. Bradman, D. Boyd-Barr, C. Johnson, and N. Holland. 2014. Organophosphate pesticide exposure, PON1, and neurodevelopment in school-age children from the CHAMACOS study. Environmental Research 134:149–157.
Eskenazi, B., A. Sookee, S.A. Rauch, E.S.Coker, A. Maphula, M. Obida, M. Crause, K.R. Kogut, R. Bornman, and J. Chevrier. 2018. Prenatal exposure to DDT and pyrethroids for malaria control and child neurodevelopment: The VHEMBE cohort, South Africa. Environmental Health Perspectives 126(4):047004.
Extoxnet. 1994. Pyrethrins and pyrethroids. http://ace.orst.edu/info/extoxnet/pips/pyrethri.htm (accessed August 12, 2018).
Farag, A.T., N.F. Goda, A.H. Mansee, and N.A. Shaaban. 2006. Effects of permethrin given before mating on the behavior of F1-generation in mice. NeuroToxicology 27(3):421–428.
Farag, A.T., A. Radwan, F. Sorour, A. El Okazy, E.-S. El-Agamy, and A. El-Khaliek El-Sebae. 2010. Chlorpyrifos induced reproductive toxicity in male mice. Reproductive Toxicology 29:80–85.
FDA (U.S. Food and Drug Administration). 2018. Lindane shampoo.https://www.accessdata.fda.gov/drugsatfda_docs/label/2003/006309shampoolbl.pdf (accessed June 26, 2018).
Figueroa, Z.I., H.A. Young, J.D. Meeker, S.E. Martenies, D. Boyd Barr, G. Gray, and M.J. Perry. 2015. Dialkyl phosphate urinary metabolites and chromosomal abnormalities in human sperm. Environmental Research 143(0):256–265.
Flores, D., V. Souza, M. Betancourt, M. Teteltitla, H. Gonzalez-Marquez, E. Casas, E. Bonilla, P. Ramírez-Noguera, M.C. Gutierrez-Ruíz, and Y. Ducolomb. 2017. Oxidative stress as a damage mechanism in porcine cumulus–oocyte complexes exposed to malathion during in vitro maturation. Environmental Toxicology 32:1669–1678.
Fortenberry, G., J.D. Meeker, B.N. Sánchez, D. Boyd Barr, P. Panuwet, D. Bellinger, L. Schnaas, M. Solano-González, A.S. Ettinger, M. Hernandez-Avila, H. Hu, and M.M. Tellez-Rojo. 2014. Urinary 3,5,6-trichloro-2-pyridinol (TCPy) in pregnant women from Mexico City: Distribution, temporal variability, and relationship with child attention and hyperactivity. International Journal of Hygiene and Environmental Health 217(0):405–412.
Fricker, R.D., E. Reardon, D.M. Spektor, S.K. Cotton, J. Hawes-Dawson, J.E. Pace, and S.D. Hosek. 2000. Pesticide use during the Gulf War: A survey of Gulf War veterans. Santa Monica, CA: RAND; National Defense Research Institute.
Furlong, M.A., S.M. Engel, D.B. Barr, and M.S. Wolff. 2014. Prenatal exposure to organophosphate pesticides and reciprocal social behavior in childhood. Environment International 70:125–131.
Furlong, M.A., A. Herring, J.P. Buckley, B.D. Goldman, J.L. Daniels, L.S. Engel, M.S. Wolff, J. Chen, J. Wetmur, D. B. Barr, S.M. Engel. 2017a. Prenatal exposure to organophosphorus pesticides and childhood neurodevelopmental phenotypes. Environmental Research 158:737–747.
Furlong, M.A., D.B. Barr, M.S. Wolff, and S.M. Engel. 2017b. Prenatal exposure to pyrethroid pesticides and childhood behavior and executive functioning. Neurotoxicology 62:231–238.
Geng, X., H. Shao, Z. Zhang, J.C. Ng, and C. Peng. 2015. Malathion-induced testicular toxicity is associated with spermatogenic apoptosis and alterations in testicular enzymes and hormone levels in male Wistar rats. Environmental Toxicology and Pharmacology 39(2):659–667.
Giri, S., S.B. Prasad, A. Giri, and G.D. Sharma. 2002. Genotoxic effects of malathion: An organophosphorus insecticide, using three mammalian bioassays in vivo. Mutation Research 514(1–2):223–231.
Gomes, J., and O.L. Lloyd. 2009. Oral exposure of mice to formulations of organophosphorous pesticides: Gestational and litter outcomes. International Journal of Environmental Health Research 19(2):125–137.
Gomes, J., O.L. Lloyd, and Z. Hong. 2008. Oral exposure of male and female mice to formulations of organophosphorous pesticides: Congenital malformations. Human & Experimental Toxicology 27(3):231–240.
González-Alzaga, B., A.F. Hernandez, M. Rodríguez-Barranco, I. Gómez, C. Aguilar-Garduño, I. López-Flores, T. Parrón, and M. Lacasaña. 2015. Pre- and postnatal exposures to pesticides and neurodevelopmental effects in children living in agricultural communities from South-Eastern Spain. Environment International 85:229–237.
Gray, L.E., Jr., and R.J. Kavlock. 1984. An extended evaluation of an in vivo teratology screen utilizing postnatal growth and viability in the mouse. Teratogenesis, Carcinogenesis, and Mutagenesis 4(5):403–426.
Gunier, R., A. Bradman, K.G. Harley, K. Kogut, and B. Eskenazi. 2017. Prenatal residential proximity to agricultural pesticide use and IQ in 7-year-old children. Environmental Health Perspectives 125(5):057002.
Hallenbeck, W.H., and K.M. Cunningham-Burns. 1985. Pesticides and human health. New York: Springer-Verlag.
Harley, K.G., K. Huen, R.A. Schall, N.T. Holland, A. Bradman, D.B. Barr, and B. Eskenazi. 2011. Association of organophosphate pesticide exposure and paraoxonase with birth outcome in Mexican-American women. PLOS ONE 6(8):e23923.
Harley, K.G., S.M. Engel, M.G. Vedar, B. Eskenazi, R.M. Whyatt, B.P. Lanphear, A. Bradman, V.A. Rauh, K. Yolton, R.W. Hornung, J.G. Wetmur, J. Chen, N.T. Holland, D.B. Barr, F.P. Perera, and M.S. Wolff. 2016. Prenatal exposure to organophosphorous pesticides and fetal growth: Pooled results from four longitudinal birth cohort studies. Environmental Health Perspectives 124(7):1084–1092.
Hassoun, E.A., and S.J. Stohs. 1996. Comparative teratological studies on TCDD, endrin, and lindane in C57BL/6J and DBA/2J mice. Comparative Biochemistry and Physiology 113C(3):393–398.
Hisada, A., J. Yoshinaga, J. Zhang, T. Kato, H. Shiraishi, K. Shimodaira, T. Okai, N. Ariki, Y. Komine, M. Shirakawa, Y. Noda, and N. Kato. 2017. Maternal exposure to pyrethroid insecticides during pregnancy and infant development at 18 months of age. International Journal of Environmental Research and Public Health 14(1):E52.
Horton, M.K., A. Rundle, D.E. Camann, D. Boyd Barr, V.A. Rauh, and R.M. Whyatt. 2011. Impact of prenatal exposure to piperonyl butoxide and permethrin on 36-month neurodevelopment. Pediatrics 127(3):e699–e706.
Horton, M.K., L.G. Kahn, F. Perera, D.B. Barr, and V. Rauh. 2012. Does the home environment and the sex of the child modify the adverse effects of prenatal exposure to chlorpyrifos on child working memory? Neurotoxicology and Teratology 34(5):534–541.
Hossain, F., O. Ali, U.J.A. D’Souza, and D.K. Saw Naing. 2010. Effects of pesticide use on semen quality among farmers in rural areas of Sabah, Malaysia. Journal of Occupational Health 52(6):353–360.
Hu, J.X., Y.F. Li, J. Li, C. Pan, Z. He, H.Y. Dong, and L.C. Xu. 2013. Toxic effects of cypermethrin on the male reproductive system: With emphasis on the androgen receptor. Journal of Applied Toxicology 33(7):576–585.
Hu, Y., L. Ji, Y. Zhang, R. Shi, W. Han, L. A. Tse, R. Pan, Y. Wang, G. Ding, J. Xu, Q. Zhang, Y. Gao, and Y. Tian. 2018. Organophosphate and pyrethroid pesticide exposures measured before conception and associations with time to pregnancy in Chinese couples enrolled in the Shanghai Birth Cohort. Environmental Health Perspectives 126(7):077001.
Huen, K., O. Solomon, K. Kogut, B. Eskenazi, and N. Holland. 2018. PON1 DNA methylation and neurobehavior in Mexican-American children with prenatal organophosphate exposure. Environment International 121:312–340.
IARC (International Agency for Research on Cancer). 1991. Deltamethrin. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans 53. Lyon, France: IARC Press.
IARC. 2016a. Malathion. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans 112. Lyon, France: IARC Press.
IARC. 2016b. Diazinon. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans 112. Lyon, France: IARC Press.
Icenogle, L.M., N.C. Christopher, W.P. Blackwelder, D.P. Caldwell, D. Qiao, F.J. Seidler, T.A. Slotkin, and E.D. Levin. 2004. Behavioral alterations in adolescent and adult rats caused by a brief subtoxic exposure to chlorpyrifos during neurulation. Neurotoxicology and Teratology 26(1):95–101.
Imai, K., J. Yoshinaga, M. Yoshikane, H. Shiraishi, M.N. Mieno, M. Yoshiike, S. Nozawa, and T. Iwamoto. 2014. Pyrethroid insecticide exposure and semen quality of young Japanese men. Reproductive Toxicology 43:38–44.
Imanishi, S., M. Okura, H. Zaha, T. Yamamoto, H. Akanuma, R. Nagano, H. Shiraishi, H. Fujimaki, and H. Sone. 2013. Prenatal exposure to permethrin influences vascular development of fetal brain and adult behavior in mice offspring. Environmental Toxicology 28(11):617–629.
Imming R.J., B.C. Shaffer, and G. Woodward. 1969. Sevin®. Safety evaluation by feeding to female beagles from day one of gestation through weaning of the offspring. Report from Woodward Research Corporation to Union Carbide Agricultural Products Company, Inc. Research Triangle Park, NC (as cited in R. L. Baron. 1991. Carbamate insecticides. In: W.J. Hayes, Jr., and E.R. Laws, Jr. (eds.). Handbook of pesticide toxicology. Vol. 3. Classes of pesticides. San Diego: Academic Press Inc. Pp. 1125–1189).
IOM (Institute of Medicine). 2003. Gulf War and health, volume 2: Insecticides and solvents. Washington, DC: The National Academies Press.
Ismail, M.F., and H.M. Mohamed. 2012. Deltamethrin-induced genotoxicity and testicular injury in rats: Comparison with biopesticide. Food and Chemical Toxicology 50(10):3421–3425.
Issam, C., H. Samir, H. Zohra, Z. Monia, and B.C. Hassen. 2009. Toxic responses to deltamethrin (DM) low doses on gonads, sex hormones and lipoperoxidation in male rats following subcutaneous treatments. Journal of Toxicological Sciences 34(6):663–670.
Ji, G., Y. Xia, A. Gu, X. Shi, Y. Long., L. Song, S. Wang, and X. Wang. 2011. Effects of non-occupational environmental exposure to pyrethroids on semen quality and sperm DNA integrity in Chinese men. Reproductive Toxicology 31:171–176.
Joshi, S.C., R. Mathur, and N. Gulati. 2007. Testicular toxicity of chlorpyrifos (an organophosphate pesticide) in albino rat. Toxicology and Industrial Health 23:439–444.
Jurewicz, J., M. Radwan, B. Wielgomas, W. Sobala, M. Piskunowicz, P. Radwan, M. Bochenek, and W. Hanke. 2015. The effect of environmental exposure to pyrethroids and DNA damage in human sperm. Systems Biology in Reproductive Medicine 61(1):37–43.
Kaloianova, F.P., and M.A. El Batawi. 1991. Human toxicology of pesticides. Boca Raton, FL: CRC Press.
Khaki, A., A.A. Khaki, and A. Rajabzadeh. 2017. The effects of permethrin and antioxidant properties of Allium cepa (onion) on testicles parameters of male rats. Toxin Reviews 36(1):1–6.
Khera, K.S., C. Whalen, G. Trivett, and G. Angers. 1979. Teratogenicity studies on pesticidal formulations of dimethoate, diuron, and lindane in rats. Bulletin of Environmental Contamination and Toxicology 22(4–5):522–529.
Kidd, H., and D.R. James (eds.). 1991. The agrochemicals handbook. 3rd ed. Cambridge, UK: Royal Society of Chemistry Information Services.
King, V. 1991. Lindane: Reproductive performance study in rats treated continuously through two successive generations: Final Report. (Addendum to MRID 422210). Lab Project Number 91/0948: 91CIL004-0948. CIL-004-LIND. Unpublished study prepared by Life Science Research, Ltd. MRID Number 42246101.
Kitagawa, K., M. Wakakura, and S. Ishikawa. 1977. Light microscopic study of endocrine organs of rats treated by carbamate pesticide. Journal of Toxicological Sciences 2:53–60.
Koren, G., D. Matsui, and B. Bailey. 2003. DEET-based insect repellents: Safety implications for children and pregnant and lactating women. CMAJ 169(3):209–212.
Kotil, T., and N.D. Yon. 2015. The effects of permethrin on rat ovarian tissue morphology. Experimental and Toxicologic Pathology 67(3):279–285.
Kumar, A., and M. Nagar. 2014. Histomorphometric study of testis in deltamethrin treated albino rats. Toxicology Reports 1:401–410.
Lafiura, K.M., D.M. Bielawski, N.C. Posecion, Jr., E.M. Ostrea, Jr., L.H. Matherly, J.W. Taub, and Y. Ge. 2007. Association between prenatal pesticide exposures and the generation of leukemia-associated T(8;21). Pediatric Blood & Cancer 49(5):624–628.
Lan, A., M. Kalimian, B. Amram, and O. Kofman. 2017. Prenatal chlorpyrifos leads to autism-like deficits in C57Bl6/J mice. Environmental Health 16(43):1–10.
Lazarini, C.A., R.Y. Lima, A.P. Guedes, and M.M. Bernardi. 2004. Prenatal exposure to dichlorvos: Physical and behavioral effects on rat offspring. Neurotoxicology and Teratology 26(4):607–614.
Leong, C. T., U.J.A. D’Souza, M. Iqbal, and Z.A. Mustapha. 2013. Lipid peroxidation and decline in antioxidant status as one of the toxicity measures of diazinon in the testis. Redox Report 18(4):155–164.
Lindenau, A., B. Fischer, P. Seiler, and H.M. Beier. 1994. Effects of persistent chlorinated hydrocarbons on reproductive tissues in female rabbits. Human Reproduction 9:772–780.
Litchfield, M.H. 1985. Toxicity to mammals. In J.P. Leahey (ed.). The pyrethroid insecticides. London: Taylor & Francis. Pp. 99–150.
Liu, B., K.H. Jung, M.K. Horton, D.E. Camann, X. Liu, A.M. Reardon, M.S. Perzanowski, H. Zhang, F.P. Perera, R.M. Whyatt, and R.L. Miller. 2012. Prenatal exposure to pesticide ingredient piperonyl butoxide and childhood cough in an urban cohort. Environment International 48:156–161.
Lopez-Casas, P.P., S.C. Mizrak, L.A. Lopez-Fernandez, M. Paz, D.G. de Rooij, and J. del Mazo. 2012. The effects of different endocrine disruptors defining compound-specific alterations of gene expression profiles in the developing testis. Reproductive Toxicology 33(1):106–115.
Luca, D., and M. Balan. 1987. Sperm abnormality assay in the evaluation of the genotoxic potential of carbaryl in rats. Morphologie et Embryologie 33(1):19–22.
Mahgoub, A.A., and A.H. El-Medany. 2001. Evaluation of chronic exposure of the male rat reproductive system to the insecticide methomyl. Pharmacological Research 44(2):73–80.
Malaviya, M., R. Husain, and P.K. Seth. 1993. Perinatal effect of two pyrethroid insecticides on brain neurotransmitter function in the neonatal rat. Veterinary and Human Toxicology 35(2):119–122.
Mandal, T.K., and N.S. Das. 2011. Correlation of testicular toxicity and oxidative stress induced by chlorpyrifos in rats. Human and Experimental Toxicology 30(10):1529–1539.
Mandal, T.K., and N.S. Das. 2012. Testicular gametogenic and steroidogenic activities in chlorpyrifos insecticide-treated rats: A correlation study with testicular oxidative stress and role of antioxidant enzyme defence systems in Sprague-Dawley rats. Andrologia 44(2):102–115.
Manikkam, M., C. Guerrero-Bosagna, R. Tracey, M.M. Haque, and M.K. Skinner. 2012a. Transgenerational actions of environmental compounds on reproductive disease and identification of epigenetic biomarkers of ancestral exposures. PLOS ONE 7(2):e31901.
Manikkam, M., R. Tracey, C. Guerrero-Bosagna, and M.K. Skinner. 2012b. Pesticide and insect repellent mixture (permethrin and DEET) induces epigenetic transgenerational inheritance of disease and sperm epimutations. Reproductive Toxicology 34(4):708–719.
Maranghi, F., M. Rescia, C. Macri, E. Di Consiglio, G. De Angelis, E. Testai, D. Farini, M. De Felici, S. Lorenzetti, and A. Mantovani. 2007. Lindane may modulate the female reproductive development through the interaction with ER-beta: An in vivo–in vitro approach. Chemico-Biological Interactions 169(1):1–14.
Marks, A.R., K. Harley, A. Bradman, K. Kogut, D.B. Barr, C. Johnson, N. Calderon, and B. Eskenazi. 2010. Organophosphate pesticide exposure and attention in young Mexican-American children: The CHAMACOS study. Environmental Health Perspectives 118(12):1768–1774.
Martenies, S.E., and M.J. Perry. 2013. Environmental and occupational pesticide exposure and human sperm parameters: A systematic review. Toxicology 307:66–73.
Matsuura, I., T. Saitoh, E. Tani, Y. Wako, H. Iwata, N. Toyota, Y. Ishizuka, M. Namiki, N. Hoshino, M. Tsuchitani, and Y. Ikeda. 2005. Evaluation of a two-generation reproduction toxicity study adding endpoints to detect endocrine disrupting activity using lindane. Journal of Toxicological Sciences 30(Spec Iss.):135–161.
Meeker, J.D., L. Ryan, D.B. Barr, R.F. Herrick, D.H. Bennett, R. Bravo, and R. Hauser. 2004a. The relationship of urinary metabolites of carbaryl/naphthalene and chlorpyrifos with human semen quality. Environmental Health Perspectives 112(17):1665–1670.
Meeker, J.D., N.P. Singh, L. Ryan, S.M. Duty, D.B. Barr, R.F. Herrick, D.H. Bennett, and R. Hauser. 2004b. Urinary levels of insecticide metabolites and DNA damage in human sperm. Human Reproduction 19(11):2573–2580.
Meeker, J.D., L. Ryan, D.B. Barr, and R. Hauser. 2006. Exposure to nonpersistent insecticides and male reproductive hormones. Epidemiology 17(1):61–68.
Meeker, J. D., S.R. Ravi, D.B. Barr, and R. Hauser. 2008a. Circulating estradiol in men is inversely related to urinary metabolites of nonpersistent insecticides. Reproductive Toxicology 25(2):184–191.
Meeker, J.D., D.B. Barr, and R. Hauser. 2008b. Human semen quality and sperm DNA damage in relation to urinary metabolites of pyrethroid insecticides. Human Reproduction 23(8):1932–1940.
Meeker, J.D., D.B. Barr, and R. Hauser. 2009. Pyrethroid insecticide metabolites are associated with serum hormone levels in adult men. Reproductive Toxicology 27(2):155–160.
Melgarejo, M., J. Mendiola, H.M. Koch, M. Moñino-García, J.A. Noguera-Velasco, and A. Torres-Cantero. 2015. Associations between urinary organophosphate pesticide metabolite levels and reproductive parameters in men from an infertility clinic. Environmental Research 137:292–298.
Meyer, A., F.J. Seidler, and T.A. Slotkin. 2004a. Developmental effects of chlorpyrifos extend beyond neurotoxicity: critical periods for immediate and delayed-onset effects on cardiac and hepatic cell signaling. Environmental Health Perspectives 112(2):170–178.
Meyer, A., F.J. Seidler, J.E. Aldridge, C.A. Tate, M.M. Cousins, and T.A. Slotkin. 2004b. Critical periods for chlorpyrifos-induced developmental neurotoxicity: Alterations in adenylyl cyclase signaling in adult rat brain regions after gestational or neonatal exposure. Environmental Health Perspectives 112(3):295–301.
Meyer, K.J., J.S. Reif, D.N. Rao Veeramachaneni, T.J. Luben, B.S. Mosley, and J.R. Nuckols. 2006. Agricultural pesticide use and hypospadias in Eastern Arkansas. Environmental Health Perspectives 114(10):1589–1595.
Millenson, M.E., J.M. Braun, A.M. Calafat, D.B. Barr, Y.T. Huang, A. Chen, B.P. Lanphear, and K. Yolton. 2017. Urinary organophosphate insecticide metabolite concentrations during pregnancy and children’s interpersonal, communication, repetitive, and stereotypic behaviors at 8 years of age: The HOME study. Environmental Research 157:9–16.
Miranda-Contreras, L., R. Gomez-Perez, G. Rojas, I. Cruz, L. Berrueta, S. Salmen, M. Colmenares, S. Barreto, A. Balza, L. Zavala, Y. Morales, Y. Molina, L. Valeri, C. A. Contreras, and J. A. Osuna. 2013. Occupational exposure to organophosphate and carbamate pesticides affects sperm chromatin integrity and reproductive hormone levels among Venezuelan farm workers. Journal of Occupational Health 55(3):195–203.
Miyamoto J. 1976. Degradation, metabolism and toxicity of synthetic pyrethroids. Environmental Health Perspectives 14:15–28.
Monge, P., C. Wesseling, J. Guardado, I. Lundberg, A. Ahlbom, K.P. Cantor, E. Weiderpass, and T. Partanen. 2007. Parental occupational exposure to pesticides and the risk of childhood leukemia in Costa Rica. Scandinavian Journal of Work, Environment & Health 33(4):293–303.
Mosbah, R., M.I. Yousef, F. Maranghi, and A. Mantovani. 2016. Protective role of Nigella sativa oil against reproductive toxicity, hormonal alterations, and oxidative damage induced by chlorpyrifos in male rats. Toxicology and Industrial Health 32(7):1266–1277.
Nayshteyn, S.Y., and D.L. Leybovich. 1971. Low doses of DDT, γ-HCCH and mixtures of these: Effect on sexual function and embryogenesis in rats. Gigiena i Sanitariia 36(5):19–22 [In Russian]. As cited in: W.J. Hayes, Jr., and E.J. Laws, Jr. (eds.). 1991. Handbook of pesticide toxicology. Vol. 2, Classes of pesticides. San Diego: Academic Press, Inc.
Neta, G., L.R. Goldman, D. Barr, B.J. Apelberg, F.R. Witter, and R.U. Halden. 2011. Fetal exposure to chlordane and permethrin mixtures in relation to inflammatory cytokines and birth outcomes. Environmental Science and Technology 45(4):1680–1687.
Ngoula, F., P. Watcho, T.S. Bouseko, A. Kenfack, J. Tchoumboue, and P. Kamtchouing. 2007. Effects of propoxur on the reproductive system of male rats. African Journal of Reproductive Health 11(1):125–132.
Nilsson, E., G. Larsen, M. Manikkam, C. Guerrero-Bosagna, M.I. Savenkova, M.K. Skinner. 2012. Environmentally induced epigenetic transgenerational inheritance of ovarian disease. PLOS ONE 7:e36129.
OEHHA (California Office of Environmental Health Hazard Assessment). 2017. Meeting synopsis and slide presentations from the Developmental and Reproductive Toxicant Identification Committee Meeting held on November 29, 2017. Sacramento, CA: California Environmental Protection Agency.
OEHHA. 2018. OEHHA chemical database. https://oehha.ca.gov/chemicals/carbaryl (accessed July 2, 2018).
Ojha, A., and Y.K. Gupta. 2015. Evaluation of genotoxic potential of commonly used organophosphate pesticides in peripheral blood lymphocytes of rats. Human and Experimental Toxicology 34(4):390–400.
Okamura, A., M. Kamijima, E. Shibata, K. Ohtani, K. Takagi, J. Ueyama, Y. Watanabe, M. Omura, H. Wang, G. Ichihara, T. Kondo, and T. Nakajima. 2005. A comprehensive evaluation of the testicular toxicity of dichlorvos in Wistar rats. Toxicology 213(1–2):129–137.
Okamura, A., M. Kamijima, K. Ohtani, O. Yamanoshita, D. Nakamura, Y. Ito, M. Miyata, J. Ueyama, T. Suzuki, R. Imai, K. Takagi, and T. Nakajima. 2009. Broken sperm, cytoplasmic droplets and reduced sperm motility are principal markers of decreased sperm quality due to organophosphorus pesticides in rats. Journal of Occupational Health 51(6):478–487.
Ostrea, Jr., E.M., A. Reyes, E. Villanueva-Uy, R. Pacifico, B. Benitez, E. Ramos, R.C. Bernardo, D.M. Bielawski, V. Delaney-Black, L. Chiodo, J.J. Janisse, and J.W. Ager. 2012. Fetal exposure to propoxur and abnormal child neurodevelopment at 2 years of age. NeuroToxicology 33(4):669–675.
PAC (Presidential Advisory Committee on Gulf War Veterans’ Illnesses). 1996. Presidential Advisory Committee on Gulf War Veterans’ Illnesses: Final report. Washington, DC: US Government Printing Office.
Palmer, A.K., D.D. Cozens, E.J.F. Spicer, and A.N. Worden. 1978. Effects of lindane upon reproductive function in a 3-generation study in rats. Toxicology 10(1):45–54.
Pant, N., S.C. Srivastava, A.K. Prasad, R. Shankar, and S.P. Srivastava. 1995. Effects of carbaryl on the rat’s male reproductive system. Veterinary and Human Toxicology 37(5):421–425.
Pant. N., R. Shankar, and S.P. Srivastava. 1996. Spermatotoxic effects of carbaryl in rats. Human and Experimental Toxicology 15(9):736–738.
Perera, F.P., V. Rauh, W.Y. Tsai, P. Kinney, D. Camann, D. Barr, T. Bernert, R. Garfinkel, Y.H. Tu, D. Diaz, J. Dietrich, and R.M. Whyatt. 2003. Effects of transplacental exposure to environmental pollutants on birth outcomes in a multiethnic population. Environmental Health Perspectives 111(2):201–205.
Pérez-Herrera, N., H. Polanco-Minaya, E. Salazar-Arredondo, M.J. Solis-Heredia, I. Hernandez-Ochoa, E. Rojas-Garcia, J. Alvarado-Mejia, V.H. Borja-Aburto, and B. Quintanilla-Vega. 2008. PON1Q192R genetic polymorphism modifies organophosphorous pesticide effects on semen quality and DNA integrity in agricultural workers from southern Mexico. Toxicology and Applied Pharmacology 230(2):261–268.
Perry, M.J., S.A. Venners, D. Barr, and X. Xu. 2007. Environmental pyrethroid and organophosphorus insecticide exposures and sperm concentration. Reproductive Toxicology 23:113–118.
Perry, M.J., S.A. Venners, X. Chen, and X. Liu. 2011. Organophosphorous pesticide exposures and sperm quality. Reproductive Toxicology 31:75–79.
Petrelli, G., G. Siepi, L. Miligi, and P. Vineis. 1993. Solvents in pesticides. Scandanavian Journal of Work, Environment & Health 19(1):64–65.
Petrelli, G., I. Figa-Talamanca, R. Tropeano, M. Tangucci, C. Cini, S. Aquilani, L. Gasperini, and P. Meli. 2000. Reproductive male-mediated risk: Spontaneous abortion among wives of pesticide applicators. European Journal of Epidemiology 16(4):391–393.
Pina-Guzman, B., M.J. Solis-Heredia, and B. Quintanilla-Vega. 2005. Diazinon alters sperm chromatin structure in mice by phosphorylating nuclear protamines. Toxicology and Applied Pharmacology 202(2):189–198.
Ploteau, S., J.-P. Antignac, C. Volteau, P. Marchand, A. Vénisseau, V. Vacher, and B. Le Bizec. 2016. Distribution of persistent organic pollutants in serum, omental, and parietal adipose tissue of French women with deep infiltrating endometriosis and circulating versus stored ratio as new marker of exposure. Environment International 97:125–136.
Ploteau, S., G. Cano-Sancho, C. Volteau, A. Legrand, A. Vénisseau, V. Vacher, P. Marchand, B. Le Bizec, and J.-P. Antignac. 2017. Associations between internal exposure levels of persistent organic pollutants in adipose tissue and deep infiltrating endometriosis with or without concurrent ovarian endometrioma. Environment International 108:195–203.
Polakova, H., and M. Vargova. 1983. Evaluation of the mutagenic effects of decamethrin: Cytogenetic analysis of bone marrow. Mutation Research 120(2–3):167–171.
Qiao, D., F.J. Seidler, S. Padilla, and T.A. Slotkin. 2002. Developmental neurotoxicity of chlorpyrifos: What is the vulnerable period? Environmental Health Perspectives 110(11):1097–1103.
Qiao, D., F.J. Seidler, C.A. Tate, M.M. Cousins, and T.A. Slotkin. 2003. Fetal chlorpyrifos exposure: Adverse effects on brain cell development and cholinergic biomarkers emerge postnatally and continue into adolescence and adulthood. Environmental Health Perspectives 111(4):536–544.
Qiao, D., F.J. Seidler, Y. Abreu-Villaça, C.A. Tate, M.M. Cousins, and T.A. Slotkin. 2004. Chlorpyrifos exposure during neurulation: Cholinergic synaptic dysfunction and cellular alterations in brain regions at adolescence and adulthood. Brain Research Developmental Brain Research 148(1):43–52.
Quirós-Alcalá, L., A. Bradman, K. Smith, G. Weerasekera, M. Odetokun, D.B. Barr, M. Nishioka, R. Castorina, A.E. Hubbard, M. Nicas, S.K. Hammond, T.E. McKone, and B. Eskenazi. 2012. Organophosphorous pesticide breakdown products in house dust and children’s urine. Journal of Exposure Science and Environmental Epidemiology 22(6):559–568.
Raanan, R., K.G. Harley, J.R. Balmes, A. Bradman, M. Lipsett, and B. Eskenazi. 2015. Early-life exposure to organophosphate pesticides and pediatric respiratory symptoms in the CHAMACOS cohort. Environmental Health Perspectives 123(2):179–185.
Raanan, R., J.R. Balmes, K.G. Harley, R.B. Gunier, S. Magzamen, A. Bradman, and B. Eskenazi. 2016. Decreased lung function in 7-year-old children with early-life organophosphate exposure. Thorax 71:148–153.
Radwan, M., J. Jurewicz, B. Wielgomas, W. Sobala, M. Piskunowicz, P. Radwan, and W. Hanke. 2014. Semen quality and the level of reproductive hormones after environmental exposure to pyrethroids. Journal of Occupational and Environmental Medicine 56(11):1113–1119.
Radwan, M., J. Jurewicz, B. Wielgomas, M. Piskunowicz, W. Sobala, P. Radwan, L. Jakubowski, W. Hawula, and W. Hanke. 2015. The association between environmental exposure to pyrethroids and sperm aneuploidy. Chemosphere 128:42–48.
Rauch, S.A., J.M. Braun, D.B. Barr, A.M. Calafat, J. Khoury, M.A. Montesano, K. Yolton, and B.P. Lanphear. 2012. Associations of prenatal exposure to organophosphate pesticide metabolites with gestational age and birth weight. Environmental Health Perspectives 120(7):1055–1060.
Rauh, V.A., R. Garfinkel, F.P. Perera, H.F. Andrews, L. Hoepner, D.B. Barr, R. Whitehead, D. Tang, and R.W. Whyatt. 2006. Impact of prenatal chlorpyrifos exposure on neurodevelopment in the first 3 years of life among inner-city children. Pediatrics 118(6):e1845–e1859.
Rauh, V., S. Arunajadai, M. Horton, F. Perera, L. Hoepner, D.B. Barr, and R. Whyatt. 2011. Seven-year neurodevelopmental scores and prenatal exposure to chlorpyrifos, a common agricultural pesticide. Environmental Health Perspectives 119(8):1196–1201.
Rauh, V.A., F.P. Perera, M.K. Horton, R.M. Whyatt, R. Bansal, X. Hao, J. Liu, D.B. Barr, T.A. Slotkin, and B.S. Peterson. 2012. Brain anomalies in children exposed prenatally to a common organophosphate pesticide. Proceedings of the National Academy of Sciences 109(20):7871–7876.
Rauh, V.A., W.E. Garcia, R.M. Whyatt, M.K. Horton, D.B. Barr, and E.D. Louis. 2015. Prenatal exposure to the organophosphate pesticide chlorpyrifos and childhood tremor. NeuroToxicology 51:80–86.
Ray, D. E.1991. Pesticides derived from plants and other organisms. In W.J. Hayes, Jr., and E. R. Laws, Jr. (eds.). Handbook of pesticide toxicology. San Diego: Academic Press. Pp. 585–636.
Recio-Vega, R., G. Ocampo-Gomez, V.H. Borja-Aburto, J. Moran-Martinez, and M.E. Cebrian-Garcia. 2008. Organophosphorus pesticide exposure decreases sperm quality: Association between sperm parameters and urinary pesticide levels. Journal of Applied Toxicology 28(5):674–680.
Ren, A., X. Qiu, L. Jin, J. Ma, Z. Li, L. Zhang, H. Zhu, R.H. Finnell, and T. Zhu. 2011. Association of selected persistent organic pollutants in the placenta with the risk of neural tube defects. Proceedings of the National Academy of Sciences 108(31):12770–12775.
Ridano, M.E., A.C. Racca, J.B. Flores-Martin, R. Fretes, C.L. Bandeira, L. Reyna, E. Bevilacqua, and S. Genti-Raimondi. 2017. Impact of chlorpyrifos on human villous trophoblasts and chorionic villi. Toxicology and Applied Pharmacology 329:26–39.
Roberts, J.R., and J.R. Reigart. 2013. Recognition and management of pesticide poisonings, sixth edition. EPA/735/R-98/003. Washington, DC: U.S. Environmental Protection Agency, Office of Pesticide Programs.
Rodriguez, H., H. Jara, S. Legua, D. Campos, J. Morales, and O. Espinoza-Navarro. 2017. Effects of cypermethrin on cytokeratin 8/18 and androgen receptor expression in the adult mouse Sertoli cell. Revista Internacional de Andrologia 15(2):51–57.
Ronco, A.M., K. Valdes, D. Marcus, and M. Llanos. 2001. The mechanism for lindane-induced inhibition of steroidogenesis in cultured rat Leydig cells. Toxicology 159(1–2):99–106.
Rowe, C., R. Gunier, A. Bradman, K.G. Harley, K. Kogut, K. Parra, and B. Eskenazi. 2016. Residential proximity to organophosphate and carbamate pesticide use during pregnancy, poverty during childhood, and cognitive functioning in 10-year-old children. Environmental Research 150:128–137.
Rupa, D.S., P.P Reddy, and O.S. Reddi. 1991. Reproductive performance in population exposed to pesticides in cotton fields in India. Environmental Research 55:123-128.
Sagiv, S.K., M.H. Harris, R.B. Gunier, K.R. Kogut, K.G. Harley, J. Deardorff, A. Bradman, N. Holland, and B. Eskenazi. 2018. Prenatal organophosphate pesticide exposure and traits related to autism spectrum disorders in a population living in proximity to agriculture. Environmental Health Perspectives 126(4):047012.
Sai, L., X. Li, Y. Liu, Q. Guo, L. Xie, G. Yu, C. Bo, Z. Zhang, and L. Li. 2014. Effects of chlorpyrifos on reproductive toxicology of male rats. Environmental Toxicology 29(9):1083–1088.
Saillenfait, A.M., J.P. Sabate, F. Denis, G. Antoine, A. Robert, A.C. Roudot, D. Ndiaye, and E. Eljarrat. 2017. Evaluation of the effects of cypermethrin on fetal rat testicular steroidogenesis. Reproductive Toxicology 72:106–114.
Saito, K, Y. Tomigahara, N. Ohe, N. Isobe, I. Nakatsuka, and H. Kaneko. 2000. Lack of significant estrogenic or antiestrogenic activity of pyrethroid insecticides in three in vitro assays based on classic estrogen receptor alpha-mediated mechanisms. Toxicological Sciences 57(1):54–60.
Salazar-Arredondo, E., M.D. Solis-Heredia, E. Rojas-Garcia, I. Hernandez-Ochoa, and B. Quintanilla-Vega. 2008. Sperm chromatin alteration and DNA damage by methyl-parathion, chlorpyrifos, and diazinon and their oxon metabolites in human spermatozoa. Reproductive Toxicology 25(4):455–460.
Sams, C. 2017. Urinary naphthol as a biomarker of exposure: Results from an oral exposure to carbaryl and workers occupationally exposed to naphthalene. Toxics 5(3):1–7.
Sánchez-Peña, L.C., B.E. Reyes, L. Lopez-Carrillo, R. Recio, J. Moran-Martinez, M.E. Cebrian, and B. Quintanilla-Vega. 2004. Organophosphorous pesticide exposure alters sperm chromatin structure in Mexican agricultural workers. Toxicology and Applied Pharmacology 196:108–113.
Santoni, G., F. Cantalamessa, L. Mazzucca, S. Ramagnoli, and M. Piccoli. 1997. Prenatal exposure to cypermethrin modulates rat NK cell cytotoxic functions. Toxicology 120:231–242.
Santoni, G.,F. Cantalamessa, R. Cavagna, S. Romagnoli, E. Spreghini, and M. Piccoli. 1998. Cypermethrin-induced alteration of thymocyte distribution and functions in prenatally-exposed rats. Toxicology 125:67–78.
Santoni, G., F. Cantalamessa, E. Spreghini, O. Sagretti, M. Staffolani, and M. Piccoli. 1999. Alterations of T cell distribution and functions in prenatally cypermethrin-exposed rats: Possible involvement of catecholamines. Toxicology 138:175–187.
Sarabia, L., I. Maurer, and E. Bustos-Obregón. 2009. Melatonin prevents damage elicited by the organophosphorous pesticide diazinon on mouse sperm DNA. Ecotoxicology and Environmental Safety 72(2):663–668.
Savitz, D.A., T. Arbuckle, D. Kaczor, and K.M. Curtis. 1997. Male pesticide exposure and pregnancy outcome. American Journal of Epidemiology 146(12):1025–1036.
Schleier, J.J., R.S. Davis, L.M. Barber, P.A. Macedo, and R.K.D. Peterson. 2009. A probabilistic risk assessment for deployed military personnel after the implementation of the “Leishmaniasis Control Program” at Tallil Air Base, Iraq. Journal of Medical Entomology 46(3):693–702.
Segal, T., L. Minguez-Alarcon, Y. Chiu, P. Williams, F. Nassan, R. Dadd, M. Ospina, A. Calafat, and R. Hauser. 2017. Urinary concentrations of DEET metabolites and semen parameters among men attending a fertility center. Fertility and Sterility 108(3 Suppl 1):e28–e29.
Seiler, P., B. Fischer, A. Lindenau, and H.M. Beier. 1994. Effects of persistent chlorinated hydrocarbons on fertility and embryonic development in the rabbit. Human Reproduction 9:1920–1926.
Shafer, T.J., D.A. Meyer, and K.M. Crofton. 2005. Developmental neurotoxicity of pyrethroid insecticides: Critical review and future research needs. Environmental Health Perspectives 113(2):123–136.
Sharma, P., I.A. Khan, and R. Singh. 2018. Curcumin and quercetin ameliorated cypermethrin and deltamethrin-induced reproductive system impairment in male Wistar rats by upregulating the activity of pituitary-gonadal hormones and steroidogenic enzymes. International Journal of Fertility & Sterility 12(1):72–80.
Shelton, J.F., E.M. Geraghty, D.J. Tancredi, L.D. Delwiche, R.J. Schmidt, B. Ritz, R.L. Hansen, and I. Hertz-Picciotto. 2014. Neurodevelopmental disorders and prenatal residential proximity to agricultural pesticides: The CHARGE study. Environmental Health Perspectives 122(10):1103–1109.
Siddiqui, M.K.J., S. Srivastava, S.P. Scrivastava, P.K. Mehrotra, N. Mathur, and I. Tandon. 2003. Persistent chlorinated pesticides and intrauterine foetal growth retardation: A possible solution. International Archives of Occupational and Environmental Health 76(1):75–80.
Silva, J.G., A.C. Boareto, A.K. Schreiber, D.D.B. Redivo, E. Gambeta, F. Vergara, H. Morais, J.M. Zanoveli, and P.R. Dalsenter. 2017. Chlorpyrifos induces anxiety-like behavior in offspring rats exposed during pregnancy. Neuroscience Letters 641:94–100.
Silver, M.K., J. Shao, B. Zhu, M. Chen, Y. Xia, N. Kaciroti, B. Lozoff, and J. D. Meeker. 2017. Prenatal naled and chlopyrifos exposure is associated with deficits in infant motor function in a cohort of Chinese infants. Environment International 106:248–256.
Smalley, H.E., J.M. Curtis, and F.L. Earl. 1968. Teratogenic action of carbaryl in beagle dogs. Toxicology and Applied Pharmacology 13(3):392–403.
Smith, A.G. 1991. Chlorinated hydrocarbon insecticides. In: W.J. Hayes, Jr., and E. R. Laws, Jr. (eds). Handbook of pesticide toxicology, vol. 2. New York: Academic Press. Pp. 731–915.
Sujatha, R., K.C. Chitra, C. Latchoumycandane, and P.P. Mathur. 2001. Effect of lindane on testicular antioxidant system and steroidogenic enzymes in adult rats. Asian Journal of Andrology 3(2):135–138.
Sumida, K., K. Saito, N. Ooe, N. Isobe, H. Kaneko, and I. Nakatsuka. 2001. Evaluation of in vitro methods for detecting the effects of various chemicals on the human progesterone receptor, with a focus on pyrethroid insecticides. Toxicology Letters 118(3):147–155.
Suwalsky, M., F. Villena, D. Marcus, and A.M. Ronco. 2000. Plasma absorption and ultrastructural changes of rat testicular cells induced by lindane. Human and Experimental Toxicology 19(9):529–533.
Taylor, J.S., B.M. Thomson, C.N. Lang, F.Y. Sin, and E. Podivinsky. 2010. Estrogenic pyrethroid pesticides regulate expression of estrogen receptor transcripts in mouse sertoli cells differently from 17beta-estradiol. Journal of Toxicology and Environmental Health, Part A 73(16):1075–1089.
Tomczak, S., K. Baumann, and G. Lehnert. 1981. Occupational exposure to hexachlorocyclohexane. IV. Sex hormone alterations in HCH-exposed workers. International Archives of Occupational and Environmental Health 48(3):283–287.
Traina, M.E., M. Rescia, E. Urbani, A. Mantovani, C. Macri, C. Ricciardi, A.V. Stazi, P. Fazzi, E. Cordelli, P. Eleuteri, G. Leter, and M. Spano. 2003. Long-lasting effects of lindane on mouse spermatogenesis induced by in utero exposure. Reproductive Toxicology 17(1):25–35.
Trifonova, T.K., I.N. Gladenko, and V.D. Shulyak. 1970. Effect of hexachlorocyclohexane gamma-isomer and Sevin on sexual function. Veterinariia (Kiev) 6:91–93 [In Russian]. As cited in: W.J. Hayes, Jr., and E.R. Laws, Jr. (eds.). 1991. Handbook of pesticide toxicology, vol. 2, Classes of Pesticides. San Diego: Academic Press, Inc.
Tyagi, V., N. Garg, M.D. Mustafa, B.D. Banerjee, and K.Guleria. 2015. Organochlorine pesticide levels in maternal blood and placental tissue with reference to preterm birth: A recent trend in North Indian population. Environmental Monitoring and Assessment 187(7):471.
Umosen, A.J., S.F. Ambali, J.O. Ayo, B. Mohammed, and C. Uchendu. 2012. Alleviating effects of melatonin on oxidative changes in the testes and pituitary glands evoked by subacute chlorpyrifos administration in Wistar rats. Asian Pacific Journal of Tropical Biomedicine 2(8):645–650.
Upson, K., A.J. De Roos, M.L. Thompson, S. Sathyanarayana, D. Scholes, D.B. Barr, and V.L. Holt. 2013. Organochlorine pesticides and risk of endometriosis: Findings from a population-based case–control study. Environmental Health Perspectives 121:1319–1324.
Uzun, F.G., S. Kalender, D. Durak, F. Demir, and Y. Kalender. 2009. Malathion-induced testicular toxicity in male rats and the protective effect of vitamins C and E. Food and Chemical Toxicology 47(8):1903–1908.
VA (Department of Veterans Affairs). 2017. Pesticides and Gulf War veterans.https://www.publichealth.va.gov/exposures/gulfwar/sources/pesticides.asp (accessed December 4, 2017).
Venerosi, A., G. Calamandrei, and L. Ricceri. 2006. A social recognition test for female mice reveals behavioral effects of developmental chlorpyrifos exposure. Neurotoxicology and Teratology 28:466–471.
Venerosi, A., L. Ricceri, M.L. Scattoni, and G. Calamandrei. 2009. Prenatal chlorpyrifos exposure alters motor behavior and ultrasonic vocalization in CD-1 mouse pups. Environmental Health 8:12.
Venerosi, A., L. Ricceri, A. Rungi, V. Sanghez, and G. Calamandrei. 2010. Gestational exposure to the organophosphate chlorpyrifos alters social-emotional behaviour and impairs responsiveness to the serotonin transporter inhibitor fluvoxamine in mice. Psychopharmacology (Berl) 208(1):99–107.
Venerosi, A., S. Tait, L. Stecca, F. Chiarotti, A. De Felice, M.F. Cometa, M.T. Volpe, G. Calamandrei, and L. Ricceri. 2015. Effects of maternal chlorpyrifos diet on social investigation and brain neuroendocrine markers in the offspring—A mouse study. Environmental Health 14:32.
Viel, J.F., C. Warembourg, G. Le Maner-Idrissi, A. Lacroix, G. Limon, F. Rouget, C. Monfort, G. Durand, S. Cordier, and C. Chevrier. 2015. Pyrethroid insecticide exposure and cognitive developmental disabilities in children: The PELAGIE mother–child cohort. Environment International 82:69–75.
Viel, J.F., F. Rouget, C. Warembourg, C. Monfort, G. Limon, S. Cordier, and C. Chevrier. 2017. Behavioural disorders in 6-year-old children and pyrethroid insecticide exposure: The PELAGIE mother–child cohort. Occupational and Environmental Medicine 74(4):275–281.
Walker, V.R., A.L. Boyles, K.E. Pelch, S.D. Holmgren, A.J. Shapiro, C.R. Blystone, M.J. Devito, R.R. Newbold, R. Blain, P. Hartman, K.A. Thayer, and A.A. Rooney. 2018. Human and animal evidence of potential transgenerational inheritance of health effects: An evidence map and state-of-the-science evaluation. Environment International 115:48–69.
Walsh, L.P., and D. M. Stocco. 2000. Effects of lindane on steroidogenesis and steroidogenic acute regulatory protein expression. Biology of Reproduction 63(4):1024–1033.
Wang, P., Y. Tian, X.J. Wang, Y. Gao, R. Shi, G.Q. Wang, G.H. Hu, and X.M. Shen. 2012. Organophosphate pesticide exposure and perinatal outcomes in Shanghai, China. Environment International 42:100–104.
Wang, Y., Y. Zhang, L. Ji, Y. Hu, J. Zhang, C. Wang, G. Ding, L. Chen, M. Kamijima, J. Ueyama, Y. Gao, and Y. Tian 2017. Prenatal and postnatal exposure to organophosphate pesticides and childhood neurodevelopment in Shandong, China. Environment International 108:119–126.
Ware, G.W. 1989. The pesticide book, 3rd ed. Fresno, CA: Thomson Publications.
Watkins, D.J., G.Z. Fortenberry, B.N. Sanchez, D.B. Barr, P. Panuwet, L. Schnaas, E. Osorio-Valencia, M. Solano-Gonzalez, A.S. Ettinger, M. Hernandez-Avila, H. Hu, M.M. Tellez-Rojo, and J.D. Meeker. 2016. Urinary 3-phenoxybenzoic acid (3-PBA) levels among pregnant women in Mexico City: Distribution and relationships with child neurodevelopment. Environmental Research 147:307–313.
Wauchope, R.D., T.M. Buttler, A.G. Hornsby, P.W.M. Augustijn-Beckers, and J.P. Burt. 1992. SCS/ARS/CES pesticide properties database for environmental decision making. Reviews of Environmental Contamination and Toxicology 123:1–155.
Wessels, D., D.B. Barr, and P. Mendola. 2003. Use of biomarkers to indicate exposure of children to organophosphate pesticides: Implications for a longitudinal study of children’s environmental health. Environmental Health Perspectives 111(16):1939–1946.
Whorton, M.D., T.H. Milby, H.A. Stubbs, B.H. Avashia, and E.Q. Hull. 1979. Testicular function among carbaryl-exposed employees. Journal of Toxicology and Environmental Health 5(5):929–941.
Whyatt, R.M., V. Rauh, D.B. Barr, D.E. Camann, H.F. Andrews, R. Garfinkel, L.A. Hoepner, D. Diaz, J. Dietrich, A. Reyes, D. Tang, P.L. Kinney, and F.P. Perera. 2004. Prenatal insecticide exposures and birth weight and length among an urban minority cohort. Environmental Health Perspectives 112(10):1125–1132.
Whyatt, R.M., D. Camann, F.P. Perera, V.A. Rauh, D. Tang, P.L. Kinney, R. Garfinkel, H. Andrews, L. Hoepner, and D.B. Barr. 2005. Biomarkers in assessing residential insecticide exposures during pregnancy and effects on fetal growth. Toxicology and Applied Pharmacology 206(2):246–254.
Wickerham, E.L., B. Lozoff, J. Shao, N. Kaciroti, Y. Xia, and J.D. Meeker. 2012. Reduced birth weight in relation to pesticide mixtures detected in cord blood of full-term infants. Environment International 47:80–85.
Willis, W.O., A. de Peyster, C.A. Molgaard, C. Walker, and T. MacKendrick. 1993. Pregnancy outcome among women exposed to pesticides through work or residence in an agricultural area. Journal of Occupational Medicine 35(9):943–949.
Wylie, B.J., M. Hauptman, A.D. Woolf, and R.H. Goldman. 2016. Insect repellants during pregnancy in the era of the Zika virus. Obstetrics and Gynecology 128(5):1111–1115.
Wyrobek, A.J., G. Watchmaker, L. Gordon, K. Wong, D. Moore, II, and D. Whorton. 1981. Sperm shape abnormalities in carbaryl-exposed employees. Environmental Health Perspectives 40:255–265.
Xia, Y., S. Cheng, Q. Bian, L. Xu, M.D. Collins, H.C. Chang, L. Song, J. Liu, S. Wang, and X. Wang. 2005. Genotoxic effects on spermatozoa of carbaryl-exposed workers. Toxicological Sciences 85(1):615–623.
Xia, Y., Y. Han, B. Wu, S. Wang, A. Gu, N. Lu, J. Bo, L. Song, N. Jin, and X. Wang. 2008. The relation between urinary metabolite of pyrethroid insecticides and semen quality in humans. Fertility and Sterility 89(6):1743–1750.
Xue, Z., X. Li, Q. Su, L. Xu, P. Zhang, Z. Kong, J. Xu, and J. Teng. 2013. Effect of synthetic pyrethroid pesticide exposure during pregnancy on the growth and development of infants. Asia-Pacific Journal of Public Health 25(4 Suppl):72S–79S.
Yoshinaga, J., K. Imai, H. Shiraishi, S. Nozawa, M. Yoshiike, M.N. Mieno, A.M. Andersson, and T. Iwamoto. 2014. Pyrethroid insecticide exposure and reproductive hormone levels in healthy Japanese male subjects. Andrology 2(3):416–420.
Young, H.A., J.D. Meeker, S.E. Martenies, Z.I. Figueroa, D.B. Barr, and M.J. Perry. 2013. Environmental exposure to pyrethroids and sperm sex chromosome disomy: A cross-sectional study. Environmental Health 12:111.
Young, J.G., B. Eskenazi, E.A. Gladstone, A. Bradman, L. Pedersen, C. Johnson, D.B. Barr, C.E. Furlong, and N.T. Holland. 2005. Association between in utero organophosphate pesticide exposure and abnormal reflexes in neonates. NeuroToxicology 26(2):199–209.
Yu, Y., Y. Yang, X. Zhao, X. Liu, J. Xue, J. Zhang, and A. Yang. 2017. Exposure to the mixture of organophosphorus pesticides is embryotoxic and teratogenic on gestational rats during the sensitive period. Environmental Toxicology 32(1):139–146.
Yucra, S., M. Gasco, J. Rubio, and G.F. Gonzales. 2008. Semen quality in Peruvian pesticide applicators: Association between urinary organophosphate metabolites and semen parameters. Environmental Health 7:59.
Zhang, S.Y., Y. Ito, O. Yamanoshita, Y. Yanagiba, M. Kobayashi, K. Taya, C. Li, A. Okamura, M. Miyata, J. Ueyama, C.H. Lee, M. Kamijima, and T. Nakajima. 2007. Permethrin may disrupt testosterone biosynthesis via mitochondrial membrane damage of Leydig cells in adult male mouse. Endocrinology 148(8):3941–3949.
Zhang, S.Y., J. Ueyama, Y. Ito, Y. Yanagiba, A. Okamura, M. Kamijima, and T. Nakajima. 2008. Permethrin may induce adult male mouse reproductive toxicity due to cis isomer not trans isomer. Toxicology 248(2–3):136–141.
Zhang, Y., S. Han, D. Liang, X. Shi, F. Wang, W. Liu, L. Zhang, L. Chen, Y. Gu, and Y. Tian. 2014a. Prenatal exposure to organophosphate pesticides and neurobehavioral development of neonates: A birth cohort study in Shenyang, China. PLOS ONE 9(2):e88491.
Zhang, J., J. Yoshinaga, A. Hisada, H. Shiraishi, K. Shimodaira, T. Okai, M. Koyama, N. Watanabe, E. Suzuki, M. Shirakawa, Y. Noda, Y. Komine, N. Ariki, and N. Kato. 2014b. Prenatal pyrethroid insecticide exposure and thyroid hormone levels and birth sizes of neonates. Science of the Total Environment 488–489:275–279.
Zidan, N.E.A. 2009. Evaluation of the reproductive toxicity of chlorpyrifos methyl, diazinon, and profenofos pesticides in male rats. International Journal of Pharmacology 5(1):51–57.
This page intentionally left blank.