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Journal of Toxicology and
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Epidemiologic Evidence of Relationships Between
Reproductive and Child Health Outcomes and
Environmental Chemical Contaminants
Donald T. Wigle a; Tye E. Arbuckle b; Michelle C. Turner a; Annie Bérubé c;
Qiuying Yang d; Shiliang Liu e; Daniel Krewski a
a
McLaughlin Centre for Population Health Risk Assessment, University of Ottawa,
Ottawa, Ontario
b
Healthy Environments and Consumer Safety Branch, Health Canada, Ottawa,
Ontario
c
Vulnerable Populations Division, Safe Environments Program, Health Canada, Ottawa, Ontario
d
OMNI Research Group, Department of Obstetrics and Gynecology, Faculty of Medicine, University of Ottawa and
Ottawa Health Research Institute, Ottawa, Ontario
e
Centre for Healthy Human Development, Public Health Agency of Canada, Ottawa, Ontario, Canada
Online Publication Date: 01 May 2008
To cite this Article: Wigle, Donald T., Arbuckle, Tye E., Turner, Michelle C., Bérubé, Annie, Yang, Qiuying, Liu,
Shiliang and Krewski, Daniel (2008) 'Epidemiologic Evidence of Relationships Between Reproductive and Child Health
Outcomes and Environmental Chemical Contaminants', Journal of Toxicology and Environmental Health, Part B, 11:5,
373 — 517
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Journal of Toxicology and Environmental Health, Part B, 11:373–517, 2008
Copyright © Taylor & Francis Group, LLC
ISSN: 1093-7404 print / 1521-6950 online
DOI: 10.1080/10937400801921320
EPIDEMIOLOGIC EVIDENCE OF RELATIONSHIPS BETWEEN REPRODUCTIVE AND
CHILD HEALTH OUTCOMES AND ENVIRONMENTAL CHEMICAL CONTAMINANTS
Donald T. Wigle1, Tye E. Arbuckle2, Michelle C. Turner1, Annie Bérubé3, Qiuying Yang4,
Shiliang Liu5, Daniel Krewski1
1
McLaughlin Centre for Population Health Risk Assessment, University of Ottawa, Ottawa,
Ontario, 2Healthy Environments and Consumer Safety Branch, Health Canada, Ottawa, Ontario,
3
Vulnerable Populations Division, Safe Environments Program, Health Canada, Ottawa, Ontario,
4
OMNI Research Group, Department of Obstetrics and Gynecology, Faculty of Medicine,
University of Ottawa and Ottawa Health Research Institute, Ottawa, Ontario, and 5Centre for
Healthy Human Development, Public Health Agency of Canada, Ottawa, Ontario, Canada
This review summarizes the level of epidemiologic evidence for relationships between prenatal and/or early life
exposure to environmental chemical contaminants and fetal, child, and adult health. Discussion focuses on fetal
loss, intrauterine growth restriction, preterm birth, birth defects, respiratory and other childhood diseases, neuropsychological deficits, premature or delayed sexual maturation, and certain adult cancers linked to fetal or childhood
exposures. Environmental exposures considered here include chemical toxicants in air, water, soil/house dust and
foods (including human breast milk), and consumer products. Reports reviewed here included original epidemiologic studies (with at least basic descriptions of methods and results), literature reviews, expert group reports, metaanalyses, and pooled analyses. Levels of evidence for causal relationships were categorized as sufficient, limited, or
inadequate according to predefined criteria. There was sufficient epidemiological evidence for causal relationships
between several adverse pregnancy or child health outcomes and prenatal or childhood exposure to environmental
chemical contaminants. These included prenatal high-level methylmercury (CH3Hg) exposure (delayed developmental
milestones and cognitive, motor, auditory, and visual deficits), high-level prenatal exposure to polychlorinated
biphenyls (PCBs), polychlorinated dibenzofurans (PCDFs), and related toxicants (neonatal tooth abnormalities, cognitive and motor deficits), maternal active smoking (delayed conception, preterm birth, fetal growth deficit [FGD]
and sudden infant death syndrome [SIDS]) and prenatal environmental tobacco smoke (ETS) exposure (preterm
birth), low-level childhood lead exposure (cognitive deficits and renal tubular damage), high-level childhood CH3Hg
exposure (visual deficits), high-level childhood exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (chloracne),
childhood ETS exposure (SIDS, new-onset asthma, increased asthma severity, lung and middle ear infections, and
adult breast and lung cancer), childhood exposure to biomass smoke (lung infections), and childhood exposure to
outdoor air pollutants (increased asthma severity). Evidence for some proven relationships came from investigation
of relatively small numbers of children with high-dose prenatal or early childhood exposures, e.g., CH3Hg poisoning
episodes in Japan and Iraq. In contrast, consensus on a causal relationship between incident asthma and ETS exposure came only recently after many studies and prolonged debate. There were many relationships supported by
limited epidemiologic evidence, ranging from several studies with fairly consistent findings and evidence of doseresponse relationships to those where 20 or more studies provided inconsistent or otherwise less than convincing
evidence of an association. The latter included childhood cancer and parental or childhood exposures to pesticides.
In most cases, relationships supported by inadequate epidemiologic evidence reflect scarcity of evidence as
opposed to strong evidence of no effect. This summary points to three main needs: (1) Where relationships between
child health and environmental exposures are supported by sufficient evidence of causal relationships, there is a
need for (a) policies and programs to minimize population exposures and (b) population-based biomonitoring to
track exposure levels, i.e., through ongoing or periodic surveys with measurements of contaminant levels in blood,
urine and other samples. (2) For relationships supported by limited evidence, there is a need for targeted research
and policy options ranging from ongoing evaluation of evidence to proactive actions. (3) There is a great need for
population-based, multidisciplinary and collaborative research on the many relationships supported by inadequate
evidence, as these represent major knowledge gaps. Expert groups faced with evaluating epidemiologic evidence of
Funding for this review was provided by the National Collaborating Centre for Environmental Health. The authors acknowledge
Robert Cushman for providing helpful comments. Daniel Krewski is the NSERC/SSHRC/McLaughlin Chair in Population Health Risk
Assessment at the University of Ottawa.
This work does not necessarily reflect the views of Health Canada and the Public Health Agency of Canada and no official endorsement
should be inferred. The findings and conclusions of this report are those of the authors and do not necessarily represent the views of
Health Canada and the Public Health Agency of Canada.
Address correspondence to Donald T. Wigle, McLaughlin Centre for Population Health Risk Assessment, University of Ottawa,
Room 318B, One Stewart Street, Ottawa, ON K1N 6N5, Canada. E-mail: don.wigle@sympatico.ca
373
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374
D. T. WIGLE ET AL.
potential causal relationships repeatedly encounter problems in summarizing the available data. A major driver for
undertaking such summaries is the need to compensate for the limited sample sizes of individual epidemiologic
studies. Sample size limitations are major obstacles to exploration of prenatal, paternal, and childhood exposures
during specific time windows, exposure intensity, exposure–exposure or exposure–gene interactions, and relatively
rare health outcomes such as childhood cancer. Such research needs call for investments in research infrastructure,
including human resources and methods development (standardized protocols, biomarker research, validated exposure metrics, reference analytic laboratories). These are needed to generate research findings that can be compared
and subjected to pooled analyses aimed at knowledge synthesis.
Epidemiologic research conducted mainly since 1970 has demonstrated several causal relationships
and many possible associations between parental or childhood environmental exposures and
adverse pregnancy, childhood, and adult health outcomes. Toxicologic research supports assessment of biologic plausibility, one of the major criteria for evaluating cause-effect relationships in
humans since first applied by the U.S. Surgeon General in assessing human health effects of smoking
(U.S. Department of Health, 1964). Epidemiologic studies also draw on toxicologic research
for clues as to possible causes of human diseases and functional deficits and for biomarkers of
exposure, susceptibility, and adverse health effects.
Findings from epidemiologic and toxicologic studies have shown that the risk of adverse pregnancy, child, and delayed health outcomes depends not only on the dose and potency of a given
toxicant, but also on the occurrence of exposure during critical developmental time periods (Selevan
et al., 2000). Such evidence of critical exposure windows is congruent with biomolecular research
showing the dependence of fetal and child development on a complex orchestration of genes in
specific cell types at different times (National Academy of Sciences, 2000b). Disruption of prenatal
and early childhood developmental processes can produce permanent disability and functional
deficits, as well as delayed effects such as cancer later in life.
Children may have relatively high exposures to some environmental contaminants because of
their behavior, diet and metabolic and physiologic characteristics (Moya et al., 2004). Hand–mouth
behavior is common among young children and increases exposure to contaminants in soil, house
dust, and toys. Children eat relatively high amounts of certain foods that may contain pesticide
residues or other toxicants and take in more air, water, and food per unit body weight per day than
adults. They also have age-dependent differences in the absorption, distribution, metabolism, and
excretion of chemicals (National Academy of Sciences, 1993). For instance, breast-fed infants have
relatively high exposures to polychlorinated biphenyls (PCBs) and certain other lipid-soluble toxicants
at a time of rapid growth and development (Berlin et al., 2002). Children in disadvantaged households
(low-income and/or low parental education) are at increased risk of exposure to many environmental
hazards, including lead (Bernard & McGeehin 2003), tobacco smoke (Lund & Helgason 2005),
cockroach allergens (Cohn et al., 2006), and outdoor air pollutants (Chaix et al., 2006).
Several recent reports or books provide high-level summaries of current knowledge of relationships between child health and the environment (American Academy of Pediatrics Committee on
Environmental Health, 2003; European Environment Agency and the WHO Regional Office for
Europe [Copenhagen], 2002; Wigle, 2003; Wigle et al., 2006). In this review, we survey the level of
epidemiologic evidence for relationships between adverse reproductive and child health outcomes
and preconceptual, prenatal, and childhood exposures to environmental chemical contaminants.
Wherever possible, we rely on peer-reviewed expert group reports supplemented by subsequently
published original studies. We occasionally cite original studies included in an expert group report
to illustrate the strengths or weaknesses of supporting evidence. Space limitations precluded
comprehensive discussion of each of the many exposure–outcome relationships reviewed here.
Thus our review is a survey of the current state of knowledge in this field, as opposed to a collection of
exhaustive evaluations. Our goal was to categorize the level of evidence for many exposure–outcome
relationships to support public health research and policy planning (further discussed in the Conclusion).
We defined three basic categories for level of evidence (see Methods). Many relationships were
supported by limited evidence; it was beyond the scope of this review to define a fourth category
such as suggestive evidence, as this would have entailed substantial lengthening of the article.
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
375
Discussion is organized by health outcomes, including functional deficits, disability, and structural
abnormalities. These include fetal loss, intrauterine growth restriction, preterm birth, birth defects, cancer (including certain adult cancers linked to prenatal or childhood exposures—note that prenatal
exposure is defined throughout the text as prenatal maternal exposure), asthma, other childhood diseases, neuropsychological deficits, and premature or delayed sexual maturation. Environmental exposures considered here include chemical toxicants in air, water, soil, house dust, foods, and consumer
products. Evidence of health effects at high exposure levels is briefly summarized, but the major focus is
on potential health effects at relatively low background exposure levels to which the majority of the
population is likely exposed. The tables summarize key findings for relationships supported by limited
or sufficient epidemiologic evidence (see definitions in Methods). Additional information on the epidemiologic evidence published up to 2004 for these relationships is available elsewhere (Wigle, 2005).
METHODS
This review included peer-reviewed English-language publications and government reports identified from PubMed and TOXNET searches and other reports identified from bibliographies of retrieved
articles published by December 31, 2006. Search strategies included key words for various combinations of health outcome and environmental exposure. The PubMed “related articles” function was
used to search for other relevant articles not retrieved in initial key word searches. Key words for
adverse health outcomes included (but were not limited to) fertility, conception, pregnancy, spontaneous abortion, stillbirth, fetal death, birth weight, gestation length, preterm birth, birth defect, congenital anomaly, chromosomal abnormality, sudden infant death syndrome, asthma, otitis media,
bronchitis, bronchiolitis, pneumonia, allergy, growth, milestone, cognitive, psychomotor, auditory,
visual, hyperactivity, attention, childhood cancer, leukemia, lymphoma, brain cancer, Wilms, neuroblastoma, germ cell, bone cancer, sarcoma, retinoblastoma, testicular cancer, breast cancer, lung cancer, chloracne, teeth, enamel, dentin, renal, menarche, puberty, and Tanner stage. For exposure, key
words included (but were not limited to) environment, chemical, metal, lead, mercury, cadmium,
manganese, arsenic, polybrominated biphenyl (PBB), polychlorinated biphenyl (PCB), dioxin, pesticide, environmental tobacco smoke, air pollution, smoke, particulate matter, carbon monoxide, polycyclic aromatic hydrocarbon (PAH), ozone, volatile, water, disinfection by-product (DBP), nitrate,
nitrite, bisphenol, and phthalate. Other searches were done using names of authors of relevant articles. Terms such as case-control, cohort, review, and meta-analysis were used to narrow some
searches. Reviewed reports included original epidemiologic studies (with at least basic descriptions of
methods and results), literature reviews, expert group reports, meta-analyses, and pooled analyses.
When authoritative review articles were available, consideration of original reports was generally limited to those published since the year before the most recent review was published. Excluded papers
included case reports, analytic studies with fewer than five exposed cases or case parents, studies published more than a year before a recent authoritative review, and preliminary reports of subsequently
published studies. All included papers are cited in the text.
Levels of evidence for causal relationships were defined as: (i) sufficient—at least one expert
group has reviewed the available evidence and published a peer-reviewed report indicating a consensus view that there is a causal relationship, (ii) limited—evidence is suggestive of an association
between the agent and the outcome but is limited (and may or may not represent a causal relationship) because chance, bias and confounding cannot be ruled out with confidence, e.g., at least one
high-quality study shows a positive association but the results of other studies are inconsistent and,
(iii) inadequate—available studies are of insufficient quality (e.g., available studies have failed to
adequately control for confounding or have inadequate exposure assessment), consistency or statistical power to permit a conclusion regarding the presence or absence of an association or no studies
exist that examine the relationship. We did not attempt to identify exposure–outcome associations
for which there is limited or sufficient epidemiologic evidence of no causal relationship as the
limitations of available studies precluded firm conclusions about the absence of any risk. The definitions of limited and inadequate epidemiologic evidence are those used recently by the U.S.
National Academy of Sciences (National Academy of Sciences, 2000a)
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376
D. T. WIGLE ET AL.
RESULTS
Adverse Pregnancy Outcomes
The level of epidemiologic evidence for associations between delayed conception and fetal
growth and survival and environmental factors is summarized in Table 1.
Delayed Conception (Time to Pregnancy) Couple fecundability is defined as the probability
of conception in a menstrual cycle exposed to unprotected intercourse and can be an indicator of a
wide range of reproductive processes from gametogenesis to survival of the conceptus up to the
time of detection (Baird et al., 1986). The number of menstrual cycles it takes for a couple to conceive (time to pregnancy) can be used to assess fecundability. Delayed conception may be assessed
as a continuous (e.g., number of weeks of unprotected intercourse before conception) or categorical
variable (e.g., duration of unprotected intercourse before conception greater than 6 or 12 mo).
Time to pregnancy studies are most often retrospective; however, prospective studies that recruit
couples at the start of their attempt at pregnancy and incorporate regular testing and follow-up (as
in studies of early pregnancy loss described below) are becoming more feasible (Joffe et al., 2005).
Lead Male partner occupational exposure, inadequate evidence: While sperm concentration
has been found to be reduced among men with a blood lead concentration above 44 µg/dl (Bonde
et al., 2002), results of well-designed time to pregnancy studies have been inconsistent. In a large
multicountry European study, no clear pattern of association of time to pregnancy with short-term
lead exposure, with duration of lead exposure or with cumulative exposure to lead was observed
(Joffe et al., 2003). A more recent study of time to pregnancy among lead battery workers in Taiwan
reported a dose-response relationship with significant decreases in the likelihood of conception at
blood lead levels of 30–39 and ≥40 µg/dl (respective hazard ratios of 0.50, 95% CI 0.34–0.74, and
0.38, 95% CI 0.26–0.56) (Shiau et al., 2004).
Unspecified heavy metals Male partner occupational exposure, inadequate evidence: Studies
of occupational exposures to other heavy metals are limited. A study of workers in a smelter
reported a nonsignificant dose response relation with shorter median waiting times if only one parent
worked in the smelter compared to couples where both were employed in the smelter (Wulff et al.,
1999). Other studies have reported that male welders had lower fecundability ratios than nonwelding metal workers (Hjollund et al., 1998), and suggest that male exposure to metal fumes and
solvents in a mint is associated with an increase in time to pregnancy (Figa-Talamanca et al., 2000).
PCBs Female partner exposure, inadequate evidence:PCB exposure via fish consumption has
been examined as a risk factor for longer time to pregnancy in several cohorts. A recent review of
these studies has concluded that no firm conclusions can be drawn due to uncertainties in the
exposure estimation and inconsistencies in the results (Toft et al., 2004). One study has reported a
weak and inconclusive association between serum levels of PCBs and time to pregnancy (Law et al.,
2005), while another provided no evidence of an adverse effect related to serum PCB–153 levels
(Axmon et al., 2004).
Pesticides Male partner exposure, specific pesticides, inadequate evidence: Studies from
Finland and Canada have examined risks associated with specific pesticides. The only statistically
significant finding was for unprotected use of pyrethroids (fecundability density ratio [FDR]=0.40,
95% CI=0.19–0.85) (Sallmen et al., 2003); however, both studies suggested that fecundability
might be reduced for fungicides (Curtis et al., 1999; Sallmen et al., 2003). One study has suggested
that dichlorodiphenyltrichloroethane (DDT) exposure may be associated with reduced fecundability
(Cocco et al., 2005).
Female partner exposure, specific pesticides, inadequate evidence: Although none of the
results were statistically significant, one study has reported that preconceptual pesticide use was
associated with reduced fecundability for dicamba, glyphosate, 2,4-dichlorophenoxyacetic acid
(2,4-D), organophosphates, and thiocarbamates (Curtis et al., 1999). Two studies have used
biomonitoring data to establish exposure. One study reported that DDT exposure of mothers was
associated with decreased fecundability of their daughters but dichlorodiphenyldichloroethylene
(DDE) exposure was linked with increased fecundability of their daughters (Cohn et al., 2003).
Another study found that only at the highest concentration (≥60 µg/L) was preconceptual serum
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TABLE 1. Role of Environmental Toxicants in Delayed Conception and Fetal Growth and Survival
Toxicant
Exposure
Lead
Mercury
Arsenic
Prenatal
Paternal
Prenatal
Prenatal
Cadmium
Other and unspecified metals
PCBs
Paternal
Prenatal
Paternal
Prenatal
PBBs
TCDD
2,4,5-T, chlorophenate wood
preservatives
Other chlorophenoxy herbicides
377
Other or unspecified herbicides
Delayed concepa
I
I
I
Prenatal
Prenatal
Paternal
Paternal
Prenatal
Paternal
Prenatal
I (2,4-D, dicamba)
I (glyphosate)
Paternal
DDT/DDE
Organophosphate insecticides
Other or unspecified
insecticides, repellents
Fungicides (any)
Ethylene oxide
Unspecified pesticides
Prenatal
Paternal
Prenatal
Paternal
Prenatal
I
I
I
Paternal
Prenatal
Paternal
Prenatal
Paternal
Prenatal
Paternal
I (pyrethroids)
I
I
I
Spont. aborb
Stillbirthc
Preterm birthd
FGDe
L
L
I
L
I
I
I
Drinking water—I
Airborne—I
Airborne—I
I
L
I
L
I
Drinking water—I
Airborne—I
Airborne—I
I
Drinking water—I
Airborne—I
Airborne—I
I
High-level—I
Low-level—I
I
I
I
I
High-level—I
Low-level—I
Environ—I
Occup—I
Environ—I
Occup—I
I
I
I
I
I
I
I
I (2,4-D, 2,4-DB, MCPA)
L (2,4-D, MCPA, dicamba)
I (atrazine, glyphosate)
I (carbaryl, thiocarbamate,
atrazine, glyphosate)
L
I
I
I
I (DEET, organochlorine)
I (carbaryl, unspecified insecticdes)
I
I
L
I
I
I
I
I
I
I
I
I (atrazine,
metolachlor,
cyanazine)
I (atrazine)
L
L
L
I
I
I
I (pyrethroids, organochlorine,
unspecified)
I (organochlorine,
DEET, pyrethroids)
I (DEET, propoxur,
organochlorine)
I
I
I
I
I
I
I
I
I
(Continued)
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TABLE 1. (Continued)
Toxicant
Exposure
Delayed concepa
Spont. aborb
Stillbirthc
Preterm birthd
FGDe
Active smoking
Prenatal
Paternal
Prenatal
S
I
I
L
L
S
S
L
I
S
L
Prenatal
I
I
L
L
L
I
I
I
I
I
I
I
I
L
L
I
I
I
I
I
Environmental tobacco
smoke
Outdoor air pollutionf
near residence
Airborne industrial emissions
Drinking water DBPs
Drinking water nitrate
Hazardous waste disposal sites
Chlorinated solvents
Glycol ethers
378
Other or unspecified solvents
Bisphenol A
Oil, oil products
Plastics
Prenatal
Prenatal
Prenatal
Prenatal
Prenatal
Prenatal
Paternal
Prenatal
Paternal
Prenatal
Paternal
Paternal
L
I
I
I
I
L
L
I
L
L
I
I
I
I
I
Note. TCDD = 2,3,7,8-tetrachlorodibenzo-p-dioxin. S, Sufficient evidence—based on peer-reviewed reports of expert groups or authoritative reviews that concluded that a causal relationship existed. L, Limited evidence—includes relationships for which several epidemiologic studies, including at least one case-control or cohort study, found fairly consistent associations
and evidence of exposure-risk relationships after control for potential confounders. I, Inadequate evidence—relationships for which epidemiologic studies were limited in number and quality
(e.g., small studies, ecologic studies, limited control of potential confounders), had inconsistent results, or found little or no evidence of exposure-risk relationships.
a
For delayed conception, prenatal or paternal exposure, respectively, refer to preconceptual female or male partner exposure.
b
Clinically apparent pregnancy loss before gestation week 20.
c
Currently defined as fetal death after gestation week 20; previously defined as fetal death after gestation week 28.
d
Gestation length <37 wk.
e
Fetal growth deficit comprises small for gestational age (birth weight below 10th percentile based on gestation length) and term low birth weight (birth weight <2500 g for infants born at
37 or more weeks gestation).
f
Major pollutants from fossil fuel combustion.
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
379
DDE concentration associated with reduced fecundability (FDR=0.65, 95% CI 0.32–1.31) (Law
et al., 2005).
Female partner occupations, unspecified pesticides, inadequate evidence: Two studies have
reported that work in flower production was associated with impaired fecundability, but did not
attempt to determine risks for specific pesticides (Abell et al., 2000; Idrovo et al., 2005).
Male partner occupations, unspecified pesticides, inadequate evidence: A number of European
studies have investigated the role of occupations involving pesticide use on time to pregnancy with
conflicting results (de Cock et al., 1994; Larsen et al., 1998; Petrelli & Figa-Talamanca 2001;
Thonneau et al., 1999). While none of these studies examined risk associated with specific pesticides, some studies attempted to qualify exposures by employing crude measures of intensity of
exposure (e.g., use of protective equipment, duration of exposure) with higher exposures associated
with longer time to pregnancy (de Cock et al., 1994; Petrelli & Figa-Talamanca 2001; Sallmen
et al., 2003).
Tobacco smoke Female partner, active smoking, sufficient evidence: Female smokers have a
dose-related increased risk of delayed conception (Bolumar et al., 1996; Curtis et al., 1997; Hassan &
Killick 2004). The U.S. Surgeon General concluded there was sufficient evidence of a causal relationship between female partner active smoking and reduced fertility (U.S. Department of Health
and Human Services, 2004).
Female partner ETS exposure, inadequate evidence: There have been few epidemiologic studies
of the role of environmental tobacco smoke (ETS) exposure in delayed conception. The U.S.
Surgeon General reviewed the four available studies and found inadequate evidence for an association between reduced female fertility and female partner ETS exposure alone or in combination
with active smoking (U.S. Department of Health and Human Services, 2006). In the Avon Longitudinal Study of Pregnancy and Childhood Study, however, delayed conception was associated with
prenatal active smoking (OR=1.23, 95% CI 0.98–1.49, for delay of over 6 mo; OR=1.54, 95% CI
1.19–2.01, for delay of over 12 mo) and, independently, with prenatal ETS exposure (OR=1.17,
95% CI 1.02–1.37 and OR=1.14, 95% CI 0.92–1.42, respectively) (Hull et al., 2000).
No consistent pattern has been observed between male partner smoking and fecundability;
it appears that no published studies have assessed male partner ETS exposure and delayed
conception.
Outdoor air pollution Male or female partner exposure, inadequate evidence: One study has
reported an association between ambient sulfur dioxide (SO2) levels and fecundability in the first
unprotected menstrual cycle (Dejmek et al., 2000).
Indoor air pollution Female partner exposure, inadequate evidence: Frequent occupational
exposure to nitrous oxide for midwives has been associated with longer time to pregnancy (FDR = 0.64,
95% CI 0.44–0.95) (Ahlborg et al., 1996).
Solvents Female partner exposure, glycol ethers, inadequate evidence: There are some data
showing that females occupationally exposed to ethylene glycol ethers have longer time to pregnancy (FDR = 0.59, 95% CI 0.37–0.94) (Chen et al., 2002b).
Female partner exposure, organic solvents, inadequate evidence: Decreased fecundity was associated with female partner solvent exposure (FDR = 0.79, 95% CI 0.68–0.93), particularly those using
acetone (FDR=0.72, 95% CI 0.53–0.97) (Wennborg et al., 2001). One study reported that daily toluene exposure was associated with reduced fecundity in women (FDR = 0.47, 95% CI 0.29–0.77)
(Plenge-Bonig & Karmaus 1999). Similarly, female exposure to formaldehyde at work may have an
adverse effect on fecundity (FDR = 0.64, 95% CI 0.43–0.92) (Taskinen et al., 1999). The Agricultural
Health Study (AHS) cohort, a large study of licensed pesticide applicators in Iowa and North Carolina,
reported that the likelihood of not becoming pregnant after 12 mo of unprotected intercourse was elevated for women (OR=1.42, 95% CI 1.15–1.75) with at least monthly exposure to solvents (Sallmen
et al., 2006). Stronger associations were apparent when solvent exposure was defined as either partner
(OR=1.62, 95% CI 1.20–2.17) or both partners (OR=2.10, 95% CI 1.22–3.60).
Male partner exposure, organic solvents, inadequate evidence: Men with frequent occupational
solvent exposure experienced decreased fecundity (FDR = 0.80, 95% CI 0.57–1.11) (Sallmen et al.,
1998), but studies reported no effect (Luderer et al., 2004; Spinelli et al., 1997). Toluene exposure
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D. T. WIGLE ET AL.
was not associated with reduced fecundity in men (Plenge-Bonig & Karmaus 1999). The Agricultural
Health Study (AHS) cohort, a large study of licensed pesticide applicators in Iowa and North Carolina,
reported that the likelihood of not becoming pregnant after 12 mo of unprotected intercourse
was elevated for men (OR=1.21, 95% CI 0.93–1.57) with at least monthly exposure to solvents
(Sallmen et al., 2006). Stronger associations were apparent when solvent exposure was defined as
either partner (OR=1.62, 95% CI 1.20–2.17) or both partners (OR=2.10, 95% CI 1.22–3.60).
Other toxicants Male partner exposure, plastics, inadequate evidence: One study reported no
association between workers highly exposed to di(2-ethylhexyl)phthalate and time to pregnancy
(Modigh et al., 2002). While reduced fecundity (FDR = 0.79, 95% CI 0.59–1.05) was observed in
styrene-exposed workers, no dose response was observed in relation to tasks indicating higher
exposure (Kolstad et al., 2000).
Male partner exposure, oil and oil products, inadequate evidence: One study has reported no
major influence of exposure to hydrocarbons on time to conception (Bull et al., 1999).
Summary Except for female partner active smoking (sufficient evidence), epidemiologic
evidence for the role of environmental toxicants in delayed conception is inadequate.
Early Pregnancy Loss Early pregnancy loss is defined as pregnancy detected by daily urinary
human chorionic gonadotrophin (hCG) monitoring with loss occurring less than 6 wk after onset of
the last normal menstrual period (LNMP). Early pregnancy loss occurs in about 10–20% of conceptions and is usually not recognized or reported (Ellish et al., 1996; Hjollund et al., 2000; Wang et al.,
2003b). As there have been very few studies of early pregnancy loss and environmental exposures,
this field remains largely unexplored.
Pesticides Maternal exposure, inadequate evidence: A cohort study of newly married Chinese
female textile workers monitored for conception using daily urinary hCG measurements revealed a
dose-response relationship between early pregnancy loss (gestation length <6 wk, confirmed by
hCG) and maternal preconceptual serum DDT levels (2nd vs. 1st tertile, OR=1.07, 95% CI 0.58–1.99;
3rd vs. 1st tertile, OR=1.71, 95% CI 0.93–3.12, p-trend=.06); there were similar results when
serum DDE was modeled (Venners et al., 2005).
Spontaneous Abortion (Gestation Week <20) Spontaneous abortion is defined as unintentional fetal loss before gestation week (GW) 20. This is the definition used in most recent epidemiologic
studies, but some studies have included fetal deaths up to GW 24 or 28. Because early pregnancy loss is
rarely recognized and reported, clinically recognized spontaneous abortions virtually all occur after GW 6.
Lead Maternal exposure, limited evidence: A recent review concluded that high-level prenatal
occupational lead exposure during the 19th and early 20th centuries likely increased the risk of
spontaneous abortion and that limited evidence supports an association at prenatal blood lead
levels below approximately 30 µg/dl (Hertz-Picciotto 2000). In particular, a birth cohort study in
Mexico City revealed a monotonic dose-response relationship between spontaneous abortion and
prenatal blood lead (respective odds ratios for 5–9, 10–14, and ≥15 µg/dl were 2.3, 5.4, and 12.2,
p-trend=.03) (Borja-Aburto et al., 1999).
Paternal occupational exposure, limited evidence: Reviewers noted limited evidence for an
association between spontaneous abortion and paternal occupational lead exposure (Anttila &
Sallmen, 1995; Bellinger, 2005). Spontaneous abortion was associated with preconceptual blood
lead levels of at least 39 µg/dl in a Finnish cohort of occupationally exposed men (OR=3.0, 95% CI
1.0–8.7) (Lindbohm et al., 1991b). A similar study in British Columbia found no association with
paternal blood lead levels; selection bias is possible as only 38% of workers participated in this
study (Alexander et al., 1996). A retrospective cohort study of Norwegian men reported an elevated
risk of 2nd trimester spontaneous abortion and likely lead exposure based on job titles (OR=2.4,
95% CI 0.8–6.9) (Kristensen et al., 1993).
Inorganic and elemental mercury Maternal exposure, inadequate evidence: Case-control
studies in Massachusetts revealed an association of borderline statistical significance between spontaneous abortion and detectable drinking water mercury levels in the community of prenatal residence (OR=1.5, 95% CI 1.0–2.3) (Aschengrau et al., 1989). In a small retrospective cohort study,
spontaneous abortion risk was not elevated among women occupationally exposed to elemental
mercury (OR=1.07, 95% CI 0.27–4.56, calculated from data in paper) (Elghany et al., 1997).
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Inorganic arsenic Maternal exposure, limited evidence: A case-control study In Massachusetts
observed no association between spontaneous abortion and maternal residence in communities
with detectable arsenic in drinking water supplies (≥1.4 µg/L vs. undetectable, OR=1.5, 95% CI
0.4–4.7); the highest level detected was 1.9 µg/L, well below the current U.S. Environmental
Protection Agency (EPA) drinking water arsenic maximum contaminant level (10 µg/L) (Aschengrau
et al., 1989). A retrospective cohort study in Bangladesh revealed an increased risk of spontaneous
abortion in a town with average drinking-water arsenic levels of 240 µg/L, relative to a comparison
town with arsenic levels below 20 µg/L (RR=2.82, 95% CI 1.12–7.36, calculated from data in
report) (Ahmad et al., 2001). Spontaneous abortion risk was slightly elevated among women working
in a Swedish copper smelter (OR=1.33, 95% CI 0.94–2.08) (Wulff et al., 2002). Reviewers found
limited epidemiologic evidence and sufficient toxicologic evidence of fetal deaths after prenatal
arsenic exposure (Golub et al., 1998).
PCBs PCB congeners have half-lives in humans or monkeys of about 3–20 yr (Masuda, 2001;
Mes et al., 1995). As their concentrations per unit weight of lipid in maternal or cord blood, breast
milk, or adipose tissue samples are highly correlated, PCB levels in any of these samples provide an
index of prenatal and fetal exposure. For breast-fed infants, they also reflect lactational exposure,
especially when combined with duration of breast feeding. Although most epidemiologic studies
report serum or plasma PCB concentrations adjusted for lipid content (based on total cholesterol
and triglycerides), a substantial proportion of serum organochlorines are not associated with lipid
(Longnecker et al., 2002). Thus use of lipid-adjusted serum PCB levels may contribute to misclassification of exposure levels and reduce the chance of observing true associations with health outcomes.
Maternal high-level exposure, inadequate evidence: A retrospective cohort study of women
who consumed cooking oil contaminated by high levels of PCBs, polychlorinated dibenzofurans
(PCDFs) and related toxicants during the Yucheng incident in Taiwan revealed no association
between spontaneous abortion and maternal preconceptual serum PCB levels (baseline maternal
serum PCB >46 vs. ≤46 µg/g lipid, crude OR=1.12, 95% CI 0.34–3.70) (Yu et al., 2000).
Maternal low-level exposure, inadequate evidence: Reviewers found inadequate evidence
for an association between spontaneous abortion and background environmental PCB exposure
(Longnecker et al., 1997). In a Swedish retrospective cohort study of fishing families, spontaneous
abortion was not associated with residence in a region with fish contaminated by relatively high
PCB concentrations (1st trimester fetal death, OR=0.51, 95% CI 0.27–0.96; 2nd trimester,
OR=0.90, 95% CI 0.44–1.83) (Axmon et al., 2000). In the absence of body-burden data, interpretation of these results is difficult. A case-control study nested within a cohort of Chinese textile workers
revealed no association between spontaneous abortion and prenatal serum PCB (per 1 ng/100 g serum
increment, OR=0.96, 95% CI 0.87–1.05) (Korrick et al., 2001). In a small Japanese case-control study,
spontaneous abortion was not associated with prenatal serum PCB concentration (mean serum PCB,
cases vs. controls, 263.7 ± 136.9 vs. 319.9 ± 189.7 ng/g lipid) (Sugiura-Ogasawara et al., 2003). In an
Australian birth cohort study, pregnancy loss (spontaneous abortion or stillbirth) was not associated with
breast milk PCB levels (≤50 µg/kg lipid vs. undetectable, OR=0.60, 95% CI 0.17–2.14; >50 µg/kg,
OR=1.07, 95% CI 0.34–3.35, p-trend=0.65) (Khanjani & Sim, 2007).
PBBs Maternal exposure, inadequate evidence: A retrospective cohort study of women in
Michigan who ate meat from livestock accidentally poisoned by polybrominated biphenyls (PBBs)
revealed no association between spontaneous abortion and maternal serum PBBs at baseline soon
after exposure (>2 ppb vs. <LD, OR=0.73, 95% CI 0.47–1.13) (Small et al., 2007).
TCDD Maternal environmental exposure, inadequate evidence: A cohort study of women
living in Seveso at the time of the 1976 factory explosion that released substantial amounts of
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) found no association between spontaneous abortion
and maternal preconceptual serum TCDD levels (per 10-fold increment of maternal serum TCDD,
OR=1.0, 95% CI 0.6–1.6) (Eskenazi et al., 2003).
Paternal occupational exposure, inadequate evidence: Reviewers found insufficient evidence
for an association between spontaneous abortion and paternal occupational exposure to phenoxy
herbicides potentially contaminated with TCDD (National Academy of Sciences, 2003). This review
focused mainly on health risks for Vietnam veterans potentially exposed to Agent Orange (a 50:50
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D. T. WIGLE ET AL.
mixture of 2,4-D and 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) known to be contaminated with
TCDD). A retrospective cohort study of veterans revealed no association between spontaneous
abortion and paternal serum TCDD levels defined as background (current serum TCDD < 10 pg/g
lipid), low (current serum TCDD ≥ 10 and baseline level 10–109 pg/g lipid) and high (current serum
TCDD ≥ 10 and baseline level ≥ 110 pg/g lipid) (Wolfe et al., 1995). The respective odds ratios for
spontaneous abortion, calculated from data in the paper, were background (OR=1.13, 95% CI
0.81–1.59), low (OR=1.32, 95% CI 0.94–1.86) and high (OR=0.99, 95% CI 0.68–1.43). A subsequent study of wives of men highly exposed to TCDD during production of Agent Orange revealed
no association between spontaneous abortion and paternal serum TCDD (per 10-fold increment of
paternal serum TCDD at conception, OR=0.97, 95% CI 0.88–1.09) (Schnorr et al., 2001).
Pesticides Reviewers noted limited but somewhat inconsistent evidence for associations
between fetal deaths (spontaneous abortions or stillbirths) and maternal or paternal pesticide
exposure indices (Arbuckle & Sever, 1998; Sever et al., 1997). They noted methodologic issues,
especially inadequate exposure assessment and limited statistical power of epidemiologic studies.
A more recent review examined evidence from epidemiologic studies of pregnancy outcome that
assessed pesticide class, family, and/or active ingredient (Weselak et al., 2006). These reviewers
found limited epidemiologic evidence for associations between fetal death and DDT and inadequate evidence for associations with other pesticide categories including chlorophenoxy herbicides,
triazine herbicides, or thiocarbamate fungicides.
Maternal exposure, chlorophenoxy herbicides, inadequate evidence: Chlorophenoxy herbicides
comprise many closely related chemical entities including 2,4-dichlorophenoxyacetic acid (2,4-D1),
2,4-dichlorophenoxybutyric acid (2,4-DB), 2-methyl–4-chlorophenoxyacetic acid (MCPA), and 2methyl–4-chlorophenoxypropionic acid (MCPP, mecoprop) (Wood, 2007). The Ontario farm
family study reported no association between spontaneous abortion at GW <20 and prenatal farm
use of any chlorophenoxy herbicide (OR=1.1, 95% CI 0.6–2.1) or subtypes including
2,4-D (OR=1.0, 95% CI 0.5–2.0), 2,4-DB (OR=0.7, 95% CI 0.3–1.7) or MCPA (OR=0.9, 95%
CI 0.4–2.0) (Arbuckle et al., 1999). The Ontario study also found no association between these
herbicides and spontaneous abortions before GW 12 or during GW 12–19. A subsequent report of
this study confirmed no association between prenatal farm use of chlorophenoxy herbicides and
spontaneous abortion at GW <12 (OR=0.6, 95% CI 0.4–1.0) or GW 12–19 (OR=1.3, 95% CI
0.8–2.0) (Arbuckle et al., 2001). However, there were associations of borderline statistical significance
between late spontaneous abortion (GW 12–19) and prenatal farm use of 2,4-D (crude OR=1.6,
95% CI 0.9–2.7) or dicamba (crude OR=1.6, 95% CI 0.8–3.2). A case-only analysis within this study
showed that early spontaneous abortion (GW <12) was more likely after preconceptual compared to
postconceptual (prenatal) chlorophenoxy herbicide use (OR=1.9, 95% CI 1.1–3.3). In the Ontario
study, pesticide exposure reflected both pesticide use by the farm operator spouse (80% males) and
indirect exposure of the other spouse since the study was limited to couples on operating farms.
Maternal exposure, other herbicides, inadequate evidence: The Ontario farm family study
reported no association between spontaneous abortion before GW 20 and prenatal farm use of
atrazine (OR= 0.8, 95% CI 0.5–1.2) or glyphosate (OR=1.1, 95% CI 0.7–1.7) (Arbuckle et al.,
2001). There was also no association between these pesticides and spontaneous abortions before
GW 12 or during GW 12–19.
Maternal exposure, DDT/DDE, limited evidence: A recent review found limited evidence
based on five studies that reported associations between fetal death and biomarkers of prenatal
DDT/DDE exposure (Weselak et al., 2006). Among recent studies cited in their review, a large U.S.
nation-wide study found a nonmonotonic dose-response relationship between fetal death at any
gestation length and prenatal serum DDE (maternal serum DDE increment of 60 µg/L, OR=1.4,
95% CI 1.1–1.6) but not DDT levels (Longnecker et al., 2005). Among newly married Chinese
female textile workers monitored for conception using daily urinary hCG measurements, there was
no association between clinical spontaneous abortion (GW 6–19) and maternal preconceptual
serum DDT levels (2nd vs. 1st tertile, OR=1.22, 95% CI 0.51–2.92; 3rd vs. 1st tertile, OR=1.28,
1
First used in 1948.
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383
95% CI 0.53–3.10, p-trend=.61); there were similar results when serum DDE was modeled
(Venners et al., 2005). In an Australian retrospective cohort study, fetal death (any gestation length)
was not associated with breast milk DDT or DDE levels (2nd vs. 1st tertile DDE, OR=0.81, 95% CI
0.0.47–1.42; 3rd vs. 1st tertile, OR=0.76, 95% CI 0.41–1.39) (Khanjani & Sim, 2006).
Maternal exposure, organophosphate insecticides, inadequate evidence: A nested case-control
study in the San Francisco Bay Area revealed a weak association between spontaneous abortion
before GW 28 and prenatal residence in areas treated with aerially applied malathion less than a
week before the outcome (OR=1.20, 95% CI 0.94–1.52) but not for such use 1–4 wk before the outcome (OR=0.91, 95% CI 0.75–1.12) (Thomas et al., 1992). A retrospective cohort study of Ontario
farm families reported no association between spontaneous abortion (GW <20) and postconceptual
farm use of organophosphate insecticides (OR=0.6, 95% CI 0.4–1.0) (Arbuckle et al., 2001).
Maternal exposure, other specified insecticides or repellents, inadequate evidence: In a
randomized controlled trial of N,N-diethyl-m-toluamide (DEET) to prevent malaria during pregnancy,
spontaneous abortion risk was not increased among exposed women (OR=1.52, 95% CI 0.49–4.85)
(McGready et al., 2001). In an Australian retrospective cohort study, fetal death (any gestation
length) was not associated with breast milk dieldrin, heptachlor epoxide or oxychlordane levels
(Khanjani & Sim. 2006).
Maternal exposure, fungicides, inadequate evidence: Hexachlorobenzene (HCB) is an organochlorine fungicide used as a seed treatment, especially on wheat. A small case-control pilot study
nested within a cohort of Chinese textile workers revealed an association of borderline statistical
significance between spontaneous abortion and prenatal serum HCB (per 1 ng/100 g serum increment, OR=1.06, 95% CI 1.00–1.14) (Korrick et al., 2001). A small study of Turkish women
poisoned as children (average age 6 yr) by consumption of hexachlorobenzene-contaminated
wheat seed grain (or by breastfeeding if mothers were exposed) revealed an association between spontaneous abortion and maternal serum hexachlorobenzene levels (per log HCB (µg/L), β=2.88 ± 0.91,
p< .001) but did not assess or adjust for potential confounders (Jarrell et al., 1998). There was no association between repeated spontaneous abortion (3 or more) and maternal serum HCB levels (mean
HCB ± SD, cases vs. controls, 17.6 ± 10.2 vs. 21.2 ± 10.0 ng/g lipid, p > .05) in a small Japanese casecontrol study (Sugiura-Ogasawara et al., 2003). In an Australian retrospective cohort study, fetal death
(any gestation length) was not associated with breast milk HCB levels (2nd vs. 1st tertile, OR=1.03,
95% CI 0.61–1.75; 3rd vs. 1st tertile, OR=0.49, 95% CI 0.26–0.90) (Khanjani & Sim, 2006). The
Ontario farm family study reported no association between spontaneous abortion (GW <20) and postconceptual farm use of the broad class of fungicides (OR=0.8, 95% CI 0.5–1.1) (Arbuckle et al., 2001).
Ethylene oxide Maternal exposure, limited evidence: Ethylene oxide is used as a grain fumigant
but epidemiological studies have assessed occupational exposures in other settings. Among Finnish
nurses, spontaneous abortion was not associated with exposure to anaesthetic gases (Hemminki
et al., 1985). In a retrospective cohort study of most recent pregnancies among female dental assistants in California who conceived while working full-time, spontaneous abortion was associated
with self-reported occupational ethylene oxide exposure (OR=2.5, 95% CI 1.0–6.3) and was independent of age and exposure to nitrous oxide or preparation of mercury amalgams (Rowland et al.,
1996). A South African retrospective cohort study revealed a strong association between spontaneous abortion and occupational exposure to ethylene oxide in hospital sterilizing units while working
full-time during the relevant pregnancy (high vs. low exposure, OR=20.8, 95% CI 2.1–199); this
estimate was based on only 4 spontaneous abortions among 19 highly exposed pregnancies and 1
among 78 pregnancies of women with relatively low exposure (Gresie-Brusin et al., 2006).
Maternal exposure, unspecified pesticides, inadequate evidence: Reviewers found limited
evidence for associations between spontaneous abortion and maternal employment in agriculture
or other occupational exposure to unspecified pesticides, including dose-response relationships
with number of pesticides used and nonuse of protective equipment (Arbuckle & Sever, 1998).
Recently, a Chinese retrospective cohort study reported that spontaneous abortion was associated
with maternal use of drinking water from sources subject to runoff from pesticide-treated cotton
fields (OR=1.63, 95% CI 1.11–2.39) but not with occupational pesticide exposure (OR=0.60, 95%
CI 0.43–0.84) (Cho et al., 1999). In a retrospective cohort study of licensed pesticide applicators in
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Minnesota, spontaneous abortion was associated with prenatal mixing, loading or application of
pesticides (OR=1.81, 95% CI 1.04–3.12) (Garry et al., 2002). Although suggestive, the heterogeneity of exposure indices precludes firm conclusions and more research is needed to examine specific
pesticides, critical exposure windows and dose-response relationships.
Paternal occupational exposure, TCDD-free, inadequate evidence: Reviewers found insufficient evidence for an association between spontaneous abortion and paternal occupational exposure to chlorophenoxy herbicides potentially contaminated with TCDD (National Academy of
Sciences, 2003). See also discussion of paternal occupational TCDD exposure above.
Paternal occupational exposure, TCDD-contaminated chlorophenoxy herbicides, chlorophenoxy
herbicides, limited evidence: An initial report of the Ontario farm family cohort showed no association between spontaneous abortions before GW 20 and preconceptual use (mainly paternal) of 2,4D (OR=1.3, 95% CI 0.8–2.1), MCPA (OR=1.1, 95% CI 0.6–1.8) or dicamba (OR=1.1, 95% CI
0.6–2.1) (Savitz et al., 1997a). Further analysis revealed an association between preconceptual agricultural use (mainly paternal) of chlorophenoxy herbicides and spontaneous abortion before GW
12 (OR=2.5, 95% CI 1.0–6.4) but not with those at GW 12–19 (OR=0.4, 95% CI 0.2–1.0)
(Arbuckle et al., 1999). The Ontario study also revealed statistically nonsignificant elevated risks of
early spontaneous abortion (GW < 12) related to preconceptual paternal use of 2,4-D (OR=1.9,
95% CI 0.7–4.8) or MCPA (OR=2.3, 95% CI 0.8–6.5). A subsequent report indicated that spontaneous abortion at GW < 12 was associated with preconceptual chlorophenoxy herbicide use
(crude OR=1.5, 95% CI 1.1–2.1) and weakly with 2,4-D use (crude OR=1.3, 95% CI 0.9–2.0)
(Arbuckle et al., 2001). Although there was no association between spontaneous abortion at GW
<20 and preconceptual farm use of dicamba (OR=1.1, 95% CI 0.7–1.9, 95% CI 0.7–1.7), there
was a statistically nonsignificant elevated risk of early spontaneous abortion (GW < 12) related to
such exposure (OR=1.6, 95% CI 0.8–3.2). In a case-only analysis, early spontaneous abortions
(GW < 12) were more likely after preconceptual compared to postconceptual exposure chlorophenoxy herbicide use (OR=3.1, 95% CI 1.4–6.4) (Arbuckle et al., 2001). A retrospective cohort study
in Minnesota revealed a statistically nonsignificant association between spontaneous abortion
before GW 13 among the subset of pregnancies conceived during the spring spray season and use
of any chlorophenoxy herbicide (relative to use of any other pesticides, OR=1.59, 95% CI 0.77–
3.27) (Garry et al., 2002). However, there was an association between spontaneous abortions conceived during spring and paternal combined use of chlorophenoxy, sulfonylurea, and benzothiodazole herbicides (relative to use of any other pesticides, OR=2.94, 95% CI 1.40–6.16). In
Minnesota, herbicides are generally applied to crops in spring, and insecticides in summer and fungicides as needed in summer and fall (Garry et al., 2002).
Paternal occupational exposure, other herbicides, inadequate evidence: The Ontario farm
family study reported that spontaneous abortion before GW 20 was associated with preconceptual
use (mainly paternal) of carbaryl (OR=1.9, 95% CI 1.1–3.1) and thiocarbamate crop herbicides
(OR=1.9, 95% CI 1.1–3.3); there were also associations of borderline statistical significance with
atrazine (OR=1.5, 95% CI 0.9–2.4) and glyphosates (OR=1.5, 95% CI 0.8–2.7) (Savitz et al.,
1997a). A subsequent report of the Ontario study revealed no association between spontaneous
abortion at GW <20 and preconceptual farm use of atrazine (OR=0.8, 95% CI 0.5–1.2) or glyphosate (OR=1.1, 95% CI 0.7–1.7) (Arbuckle et al., 2001). The latter report indicated a statistically
nonsignificant elevated risk of early spontaneous abortion (GW <12) related to preconceptual farm
use of glyphosate (OR=1.4, 95% CI 0.8–2.5). The Minnesota retrospective cohort study revealed
associations between spontaneous abortion before GW 13 among the subset of pregnancies
conceived during the spring spray season and use of herbicides including sulfonylurea (relative to
use of any other pesticides, OR=2.11, 95% CI 1.09–4.09) and imidizolinone (OR=2.56, 95%
CI 1.11–5.87) (Garry et al., 2002). Further studies are needed to assess specific herbicides, critical
exposure windows, and dose-response relationships.
Paternal occupational exposure, DDT/DDE, inadequate evidence: Spontaneous abortion
before GW 20 was associated with paternal employment using backpacks to apply DDT and other
insecticides on cotton crops in India (exposed vs. unexposed men, RR=2.00, 95% CI 1.77–2.26)
(Rupa et al., 1991). A retrospective cohort study of male malaria control workers in Mexico
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reported a weak and statistically nonsignificant association between spontaneous abortion and
paternal DDE body burden estimated from self-reported information on timing, intensity, and duration of DDT exposure (4th vs. 1st quartile estimated paternal DDE, OR=1.24, 95% CI 0.91–1.70)
(Salazar-Garcia et al., 2004).
Paternal occupational exposure, organophosphate insecticides, inadequate evidence: The
Ontario farm family study reported that spontaneous abortion before GW 20 was not associated
with preconceptual use (mainly paternal) of organophosphate crop insecticides (OR=1.3, 95%
CI 0.7–2.3) (Savitz et al., 1997a). A more recent report of this study also revealed no association
between preconceptual organophosphate insecticide use and early (GW <12) (OR=1.0, 95%
CI 0.6–1.6) or late (GW 12–19) spontaneous abortion (OR=1.0, 95% CI 0.6–1.7) (Arbuckle et al.,
2001). In a case-only analysis, the Ontario study showed that early spontaneous abortion was associated with preconceptual organophosphate insecticide use (compared to postconceptual use,
OR=3.8, 95% CI 1.1–13.4). The significance of the latter finding is not clear but supports the need
for research to clarify the role of preconceptual organophosphate insecticide exposure.
Paternal occupational exposure, other or unspecified insecticides, inadequate evidence: The
Ontario farm family study reported that spontaneous abortion before GW 20 was associated with preconceptual carbaryl use (OR=2.1, 95% CI 1.1–4.1) (Savitz et al., 1997a). In a large retrospective cohort
of licensed pesticide applicators in Minnesota, spontaneous abortion was not associated with insecticide
use (insecticide and herbicide vs. herbicide only use, OR=1.27, 95% CI 0.68–2.36) (Garry et al., 2002).
Paternal occupational exposure, fungicides, inadequate evidence: The Ontario farm family
study reported that spontaneous abortion before GW 20 was not associated with preconceptual use
(mainly paternal) of fungicides (OR=1.2, 95% CI 0.7–2.1) (Savitz et al., 1997a). In a subsequent
report, there were statistically nonsignificant associations between preconceptual fungicide use and
spontaneous abortion before GW 12 (OR=1.3, 95% CI 0.9–1.9) or during GW 12–19 (OR=1.4,
95% CI 0.9–2.1) (Arbuckle et al., 2001). In a case-only analysis, spontaneous abortion before GW
12 (but not those at GW 12–19) was associated with preconceptual farm fungicide use (compared
to postconceptual use, OR=3.9, 95% CI 1.4–10.3). The Minnesota study revealed associations
between spontaneous abortion and application of the fungicides mancozeb and/or maneb
(compared to unexposed men, OR=1.77, 95% CI 1.11–2.83) and organotin fungicides
(OR=1.55, 95% CI 1.01–2.37) (Garry et al., 2002).
Ethylene oxide Paternal exposure, inadequate evidence: In a Finnish retrospective cohort
study based on linkage of national databases on pregnancy outcome and census information, spontaneous abortion was associated with paternal occupations likely exposed to ethylene oxide (based
on job-exposure matrix) (OR=4.7, 95% CI 1.2–18.4) (Lindbohm et al., 1991a).
Paternal occupational exposure, unspecified pesticides, inadequate evidence: Retrospective
cohort studies reported associations between spontaneous abortion and paternal occupational
exposure to unspecified pesticides in semi-enclosed Columbian greenhouses (exposure vs. preexposure period, OR=1.79, 95% CI 1.16–2.77) (Restrepo et al., 1990b) and backpack application of
multiple pesticides in India (OR=2.00, 95% CI 1.82–2.66) (Rupa et al., 1991). In another
retrospective cohort study, Norwegian farm families had an elevated risk of spontaneous abortion
(GW 16–27) compared to nonfarm families living in agricultural communities (OR=2.18, 95% CI
1.67–2.85) (Kristensen et al., 1997a). This study could not distinguish between maternal and paternal pesticide exposure; 57% of men and 34% of women worked at least 500 h/yr on their farms. A
small retrospective cohort study in Italy reported an association between spontaneous abortion and
paternal occupation as a pesticide applicator (OR=3.8, 95% CI 1.2–12.0) (Petrelli et al., 2000). In
a cohort of Italian greenhouse workers, there was a strong association between spontaneous abortion and paternal occupational pesticide exposure (compared to unexposed greenhouse workers,
crude OR=3.70, 95% CI 1.04–13.3, calculated from data in paper) (Petrelli et al., 2003). A retrospective cohort study of wives of agricultural pesticide aerial applicator pilots revealed no increased
risk of spontaneous abortion (crude OR=0.65, 95% CI 0.41–1.03, calculated from data in paper)
(Roan et al., 1984). Although the collective evidence from these studies is suggestive, the heterogeneity of exposure indices precludes firm conclusions.
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Tobacco smoke Prenatal active smoking, limited evidence: The U.S. Surgeon General
concluded that there is suggestive evidence of a causal relationship between spontaneous abortion
and maternal active smoking (U.S. Department of Health and Human Services, 2001).
Prenatal ETS exposure, limited evidence: Reviewers concluded that the three available epidemiologic studies provided limited evidence of an association between spontaneous abortion and
prenatal ETS exposure (Lindbohm et al., 2002). An expert panel found limited evidence of a causal
association between spontaneous abortion and prenatal ETS exposure independent of paternal
smoking (California Environmental Protection Agency, 2005). The panel noted limited evidence of
a causal association between spontaneous abortion and paternal active smoking (an important source
of prenatal ETS exposure) but noted that this could arise from preconceptual paternal germ-cell mutations rather than postconceptual maternal ETS exposure. The U.S. Surgeon General reviewed five
available studies and concluded that there was inconsistent evidence of an association between
spontaneous abortion and prenatal ETS exposure (U.S. Department of Health and Human Services,
2006). A subsequently published original case-control study of spontaneous abortions at GW 6–12
revealed an association with maternal ETS exposure (OR=1.67, 95% CI 1.17–2.38) (George et al.,
2006). This study was unusually robust in that both the outcome and exposure were confirmed
using biomarkers (hCG to confirm pregnancy and cotinine to verify exposure status).
Drinking-water disinfection by-products Most epidemiologic studies of DBPs and potential
adverse health outcomes have used drinking-water trihalomethane (THM) concentrations as an
indicator of total disinfection by-product (DBP) concentrations in chlorinated drinking water. DBP
concentrations tend to be highest in chlorinated surface waters that contain relatively high amounts
of natural organic material that reacts with chlorine to generate THMs and many other by-products.
The mutagenic activity of raw drinking water is generally very low but may increase substantially
after chlorination.
Maternal exposure, limited evidence: Reviewers noted limited and fairly consistent evidence for
an association between spontaneous abortion and drinking water disinfection by-products (Bove et al.,
2002; Graves et al., 2001; Nieuwenhuijsen et al., 2000). In experimental animals, high-dose prenatal
exposure to chloroform, bromodichloromethane (BDCM), haloacetonitriles, or haloacetic acids
(HAAs) caused fetal toxicity, including fetal resorptions and reduced fetal weight and survival
(Graves et al., 2001; Nieuwenhuijsen et al., 2000).
Hazardous waste disposal sites Maternal exposure, inadequate evidence: Reviewers found
inadequate evidence for an association between fetal deaths and prenatal residential proximity to
hazardous waste landfill sites or incinerators; there were few extant studies of spontaneous abortion
and such exposure (Johnson & DeRosa, 1995; Vrijheid, 2000). Subsequently published studies
revealed no association between spontaneous abortion and maternal residential proximity to
high-dioxin emission incinerators (OR=0.82, 95% CI 0.54–1.20, p-trend=.97 (over ten 1-km
increments)) (Tango et al., 2004) or hazardous waste disposal sites (high vs. low/moderate hazard
sites, OR=0.75, 95% CI 0.28–1.99) (Gilbreath & Kass, 2006a).
Solvents Prenatal occupational exposure, chlorinated solvents, limited evidence: Spontaneous
abortions were associated with prenatal occupational exposure to tetrachloroethylene in Scandinavia
(≥1 h/d, OR=2.88, 95% CI 0.98–8.44) (Olsen et al., 1990), California (OR=4.7, 95% CI 1.1–21.1)
(Windham et al., 1991), and the United Kingdom (dry-cleaning operator, OR=1.63, 95% CI 1.01–2.66)
(Doyle et al., 1997), trichloroethylene in California (OR=3.1, 95% CI 0.92–10.4) (Windham et al.,
1991), and with occupational chloroform use in Swedish laboratories (preconceptual exposure,
OR=2.3, 95% CI 0.9–5.9) (Wennborg et al., 2000).
Prenatal occupational exposure, glycol ethers, limited evidence: Employment in U.S. semiconductor
industry fabrication rooms during early pregnancy was associated with medical-record-confirmed
spontaneous abortion in a national cohort study (clinical spontaneous abortion, OR=1.43, 95% CI
0.95–2.09; clinical spontaneous abortion plus early pregnancy loss confirmed by daily urinary hCG
tests, OR=1.25, 95% CI 0.63–1.76) (Schenker et al., 1995). This study showed that spontaneous
abortion risk increased with level of exposure to glycol ethers and other photoresist and developer
chemicals (highly exposed vs. unexposed, OR=2.70, 95% CI 1.40–4.55) (Swan et al., 1995). In a
cohort study in two eastern U.S. semiconductor plants, spontaneous abortion was associated with
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high exposure to ethylene glycol ether exposure during the month of conception (compared to
unexposed women, OR=2.8, 95% CI 1.4–5.6) (Correa et al., 1996). A small case-control study
nested within a cohort of UK semiconductor industry female employees revealed no association
between spontaneous abortion and prenatal work in semiconductor fabrication (OR=0.64, 95% CI
0.27–1.51); there were too few cases to assess specific solvent exposures (Elliott et al., 1999).
Prenatal occupational exposure, other or unspecified solvents, limited evidence: A meta-analysis
of 5 studies published during 1988–1992 indicated a modest association between spontaneous abortion and maternal occupational exposure to organic solvents (summary OR=1.25, 95% CI 0.99–1.58)
(McMartin et al., 1998). In subsequently published studies, spontaneous abortion was associated with
prenatal occupational exposure to unspecified solvents in Finland (high vs. low 1st trimester exposure,
OR=2.3, 95% CI 1.1–4.3) (Taskinen et al., 1994) and Italy (high vs. low exposure during shoe manufacturing, OR=3.85, 95% CI 1.24–11.9) (Agnesi et al., 1997). Although the latter study did not assess
specific solvents, those commonly used in the local shoe industry at that time included ethylacetate,
cyclohexane, methylethylketone, and hexane. Spontaneous abortion was not associated with prenatal
occupational exposure to unspecified solvents in studies in Toronto (any solvent, crude OR=1.4, 95%
CI 0.4–4.9, calculated from data in paper) (Khattak et al., 1999) and Sweden (solvents other than
chloroform, OR=0.9, 95% CI 0.5–1.9) (Wennborg et al., 2000).
Paternal preconceptual exposure, glycol ethers, inadequate evidence: In a cohort study in two
eastern U.S. semiconductor plants, spontaneous abortion was not associated with high paternal
exposure to ethylene glycol ether exposure during the month of conception (compared to unexposed, OR=0.7, 95% CI 0.3–1.6) (Correa et al., 1996).
Paternal preconceptual exposure, unspecified solvents, limited evidence: A recent meta-analysis
of five epidemiologic studies published during 1984–1998 reported a statistically nonsignificant
elevated risk of spontaneous abortion related to paternal occupational solvent exposure (overall
RR=1.30, 95% CI = 0.81–2.11) (Logman et al., 2005). Among studies not included in this
meta-analysis, spontaneous abortion was associated with organic solvent exposure among men
monitored for occupational solvent exposure (OR=2.3, 95% CI 1.1–5.0) with somewhat higher
risks for painters (OR=3.3, 95% CI 1.6–6.8) and woodworkers (OR=3.8, 95% CI 1.2–12) (Taskinen
et al., 1989). In a cohort of Norwegian men in the printing industry, late spontaneous abortion (GW
20–27) was also associated with solvent exposure (with or without lead exposure, OR=5.5, 95% CI
1.8–17.2) (Kristensen et al., 1993).
Bisphenol A Bisphenol A (BPA) is a monomer used to produce polycarbonate plastics and
resins, while brominated BPA analogues are used as flame retardants.
Maternal exposure, inadequate evidence: A small case-control study reported an association
between recurrent spontaneous abortion and maternal serum bisphenol A (BPA) levels (case vs.
control mean maternal serum BPA (± SD), 2.59 ± 5.23 ng/ml vs. 0.77 ± 0.38 ng/ml, p=.02)
(Sugiura-Ogasawara et al., 2005). About half or more of 1st trimester spontaneously aborted fetuses
have chromosomal abnormalities, especially triploidy (Eiben et al., 1990), and BPA is a potent
cause of aneuploidy in mouse oocytes in vivo (Hunt et al., 2003), possibly by interfering with spindle
microtubule organization and chromosome segregation during meiosis (Can et al., 2005). In mice,
prenatal exposure to high-dose BPA produced increased fetal resorption (Morrissey et al., 1987).
Summary There was limited epidemiologic evidence for the role of environmental toxicants in
spontaneous abortion, including prenatal exposure to lead, arsenic, DDT/DDE, ethylene oxide,
active smoking, ETS, DBPs, chlorinated solvents, glycol ethers, other and unspecified solvents, and
ethylene oxide, and paternal occupational exposure to lead, chlorophenoxy herbicides other than
2,4,5-T, and other and unspecified solvents.
Stillbirths (Gestation Week ³ 20) Stillbirths are defined here at those occurring at GW 20 or
later, but some studies used other definitions (e.g., ≥28 wk). There are about 27,000 stillbirths
annually in the United States (Centers for Disease Control and Prevention, 2004). Major causes of
stillbirths include birth defects, infections, intrauterine growth restriction, gestational diabetes, and
preeclampsia (Health Canada, 2003).
Lead Maternal exposure, inadequate evidence: The Port Pirie birth cohort study found no
association between stillbirth and 2nd trimester maternal blood lead levels (only 11 cases)
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(McMichael et al., 1986). In a subsequent nested case-control study within this cohort, geometric
mean placental membrane lead levels were higher among pregnancies ending in stillbirth (2.73 µg/g,
95% CI 0.69–10.8, n=6) than controls (0.78 µg/g, 95% CI 0.61–1.00, n=22) but the difference was
not statistically significant (Baghurst et al., 1991). A large population-based case-control study found
a borderline association between stillbirth and self-reported prenatal occupational lead exposure
(Savitz et al., 1989a). A very small cohort of female survivors of childhood lead poisoning had a nonsignificantly increased risk of fetal death (any gestation length), but biomarkers of lead body burden
were not measured (Hu 1991). A small case-control study in Boston reported a nonsignificant association between stillbirth and drinking water lead levels in communities where mothers resided during
the 1st trimester (drinking water lead 2.5–80 vs. <2.5 µg/L, OR=2.1, 95% CI 0.6–7.2) (Aschengrau
et al., 1993). In a Norwegian retrospective cohort study, there was a statistically nonsignificant elevated risk of late spontaneous abortion, stillbirths and neonatal deaths among women with likely
high-level lead exposure based on job title (OR=3.7, 95% CI 0.6–13) (Irgens et al., 1998).
Paternal occupational exposure, inadequate evidence: A large U.S. population-based case-control study revealed no association between stillbirth and self-reported paternal occupational lead
exposure (Savitz et al., 1989a). Among a Norwegian cohort of printers, stillbirths were associated
with potential lead exposure based on job title (OR=2.0, 95% CI=0.9–4.7) (Kristensen et al.,
1993). A cohort of male lead smelter workers in British Columbia reported statistically nonsignificant associations between stillbirth and birth defects (12 stillbirths and 30 birth defect cases were
combined for analysis) and blood lead levels of at least 25 µg/dl (Alexander et al., 1996). The
dilution of stillbirths by birth defect cases reduces the usefulness of this study. In a large Norwegian
retrospective cohort study, late spontaneous abortion, stillbirths and neonatal deaths combined
were not associated with likely paternal high-level lead exposure based on job title (OR=1.2, 95%
CI 0.7–1.9) (Irgens et al., 1998).
Mercury Maternal exposure, inadequate evidence: Stillbirths were not associated with
detectable inorganic mercury levels in drinking water in the community of prenatal residence in
Massachusetts (OR=0.7, p > .05) (Aschengrau et al., 1993) or with prenatal occupational exposure
to airborne elemental mercury in North Carolina (OR=1.42, 95% CI 0.12–37.2, calculated from
data in paper) (Elghany et al., 1997).
Inorganic arsenic Maternal exposure, drinking water, inadequate evidence: A small case-control
study in Massachusetts reported no association between stillbirth and detectable drinking water
arsenic levels in the community of prenatal residence (drinking-water arsenic >0.8 µg/L vs. undetectable, OR=0.7; confidence interval not stated, p > .05; highest arsenic level was 2.6 µg/L)
(Aschengrau et al., 1993). In areas with drinking water arsenic levels above 100 µg/L, elevated
stillbirth risks were reported in ecologic studies in Hungary (rate per 1000 live births, exposed vs.
comparison region, 7.7 and 2.8, p=.03) (Borzsonyi et al., 1992) and Chile (exposed vs. comparison
region, RR=1.72, 95% CI 1.54–1.93) (Hopenhayn-Rich et al., 2000) and in a retrospective cohort study
in Bangladesh (RR=2.24, 95% CI 0.86–6.04, calculated from data in report) (Ahmad et al., 2001).
Maternal exposure, airborne, inadequate evidence: A review of epidemiologic and toxicologic
literature noted limited evidence of associations between fetal deaths and arsenic exposure in
humans and indicated that prenatal inorganic arsenic exposure can produce fetal death in experimental animals (Golub et al., 1998). Two original studies were not included in the latter review.
There was no association between stillbirth and prenatal airborne arsenic exposure (based on
self-reported work history) in a nationwide U.S. case-control study (OR=1.0, 95% CI 0.7–1.3)
(Savitz et al., 1989a). A case-control study in a Texan community with an arsenical pesticide plant
reported an association between stillbirth and estimated ambient air arsenic levels above 100 µg/m3
near the prenatal residence (OR=4.0, 95% CI 1.2–13.7) (Ihrig et al., 1998).
Paternal occupational exposure, airborne, inadequate evidence: A U.S. case-control study
found no association between stillbirth and self-reported paternal occupations with likely arsenic
exposure to airborne arsenic (OR=1.0, 95% CI 0.8–1.2) (Savitz et al., 1989a). A small case-control
study revealed an inverse association between stillbirth and parental employment in a Swedish
copper smelter (no data reported on maternal/paternal status but high-exposure jobs were held
mainly by men) (Wulff et al., 1995).
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Cadmium Maternal exposure, inadequate evidence: In a Massachusetts case-control, stillbirths were not associated with low-level drinking water cadmium levels in communities of prenatal residence (≥0.4 vs. <0.4 µg/L, OR = 1.2, p > .05) (Aschengrau et al., 1993). A Swedish
ecologic study demonstrated no overall increased risk of stillbirths among women living in municipalities with soil cadmium levels exceeding 1.6 µg/g (OR = 0.97, 95% CI 0.72–1.29); in the
municipality with the highest soil cadmium level (12 µg/g), stillbirth risk was elevated (OR = 2.17,
95% CI 1.06–4.49) (Landgren, 1996).
PCBs Maternal exposure, high-level exposure, inadequate evidence: A retrospective cohort
study of women exposed to high levels of PCBs, PCDFs and related toxicants during the Yucheng
incident revealed no increased risk of stillbirths (maternal serum PCB >46 vs. ≤46 ng/L, OR=1.35,
95% CI 0.35–5.26) (Yu et al., 2000).
Maternal exposure, low-level exposure, inadequate evidence: Reviewers found insufficient
evidence for an association between stillbirth and background PCB exposure (Longnecker et al.,
1997). In a Swedish retrospective cohort study of fishing families, stillbirth was not associated with
residence in a region with fish contaminated by relatively high PCB concentrations (OR=1.58, 95%
CI 0.50–5.04) (Axmon et al., 2000).
TCDD Paternal occupational exposure, inadequate evidence: Compared to Vietnam veterans
not exposed to Agent Orange, there were statistically nonsignificant elevated risks of stillbirth (odds
ratios calculated from data in paper) among exposure groups categorized as background (current and
baseline TCDD ≤10 ng/L, OR=1.89, 95% CI 0.68–5.12) or low (current TCDD ≤ 10 ng/L and baseline ≤110 ng/L, OR=1.90, 95% CI 0.64–5.43); the odds ratio for the high TCDD category (current
>10 ng/L and baseline >110 ng/L) was not calculated as there was only 1 exposed case father (Wolfe
et al., 1995). Among a cohort of British Columbia sawmill workers, stillbirths were not associated with
paternal occupational exposure to chlorophenate wood preservatives contaminated by TCDD and
related toxicants (per 100-h increment of exposure up to 3 mo before conception, OR=1.0, 95% CI
0.97–1.063) (Dimich-Ward et al., 1996).
Pesticides Maternal exposure, nonchlorophenoxy herbicides, inadequate evidence: In a California
case-cohort study, there was an association of borderline statistical significance between stillbirths/
neonatal deaths from birth defects and agricultural use of paraquat during GW 3–8 (OR=1.8, 95%
CI 0.9–3.9) in the same square mile section as the prenatal residence (Bell et al., 2001b).
Maternal exposure, DDT/DDE, limited evidence: A review of available studies found limited
evidence of an association between fetal deaths and biomarkers of maternal DDT/DDE exposure
(Weselak et al., 2006). Among recent studies included in their review, a retrospective cohort study
based on the U.S. Collaborative Perinatal Project reported a nonmonotonic dose-response relationship
between fetal deaths of any gestation length and prenatal serum DDE levels (per 60 µg/L maternal
serum DDE increment, OR=1.4, 95% CI 1.1–1.6) (Longnecker et al., 2005). The latter authors
stated that the association with maternal serum DDE was similar for spontaneous abortions and stillbirths but did not include supporting data. A subsequently reported Australian retrospective cohort
study observed no association between fetal death (any gestation length) and breast milk DDT or
DDE levels (e.g., 3rd vs. 1st tertile DDE, OR=0.76, 95% CI 0.41–1.39) (Khanjani & Sim, 2006).
Maternal exposure, organophosphate insecticides, inadequate evidence: A nested case-control
study reported an elevated stillbirth risk among women living in regions sprayed with malathion 1–4 wk
before the outcome (OR=1.95, 95% CI 0.88–4.35) (Thomas et al., 1992). A California-wide study
revealed that agricultural organophosphate application (mainly insecticides) during GW 3–8 in the
same square mile section as the prenatal residence was associated with stillbirths/neonatal deaths
from birth defects (OR=2.9, 95% CI 1.3–6.4) (Bell et al., 2001b) but not with other stillbirths (1st
trimester exposure, OR=1.0, 95% CI 0.7–1.3) (Bell et al., 2001c). Although suggestive, these findings require confirmation and exploration of dose-response relationships.
Maternal exposure, pyrethroid insecticides, inadequate evidence: Stillbirths/neonatal deaths
from causes other than birth defects in California were not associated with agricultural pyrethroid
insecticide application in the same square mile section as the maternal residence (1st trimester use,
OR=1.0, 95% CI 0.6–1.6) (Bell et al., 2001c). However, there was an association between stillbirths/neonatal deaths from birth defects and agricultural pyrethroid insecticide application during
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GW 3–8 in the same square mile section as the maternal residence (OR=4.9, 95% CI 1.9–12.9);
this analysis excluded subjects for whom other pesticide classes were also used during GW 3–8 (Bell
et al., 2001a, 2001b). The California study used a highly detailed statewide pesticide use database
and did not have to rely on individual pesticide exposure recall; however, it did not assess potential
relationships between stillbirths/neonatal deaths and preconceptual paternal exposure.
Maternal exposure, organochlorine insecticides, inadequate evidence: An Australian retrospective cohort study found no association between fetal death (any gestation length) and breast milk
dieldrin (3rd vs. 1st tertile, OR=0.73, 95% CI 0.42–1.27), heptachlor epoxide (OR=0.82, 95% CI
0.48–1.39) or oxychlordane levels (OR=0.54, 95% CI 0.29–1.00) (Khanjani & Sim, 2006).
Maternal exposure, unspecified insecticides, inadequate evidence: A case-control study in
California agricultural counties revealed an association of borderline statistical significance between
all stillbirths/neonatal deaths and 1st trimester indoor use of insecticides at the maternal residence
(OR=1.4, 95% CI 0.9–2.3) but not insecticide use on pets (OR=1.0, 95% CI 0.5–1.9) (Pastore
et al., 1997). There was a statistically nonsignificant elevated risk of stillbirths among Sudanese
women who reported prenatal spraying of pesticides (mainly insecticides) using hand pumps
indoors at home (OR=1.6, 95% CI 0.8–3.3) (Taha & Gray, 1993).
Maternal exposure, fungicides, inadequate evidence: An Australian retrospective cohort study
found no association between fetal death (any gestation length) and breast HCB levels (3rd vs. 1st
tertile, OR=0.49, 95% CI 0.26–0.90) (Khanjani & Sim, 2006).
Maternal exposure, unspecified pesticides, inadequate evidence: Reviewers found limited epidemiologic evidence for associations between fetal deaths and maternal exposure to unspecified
pesticides (Arbuckle & Sever, 1998). Among studies of stillbirths published before these reviews,
prenatal exposures linked to elevated stillbirth risks included maternal employment in agriculture
or horticulture (compared to all women in cohort, RR=5.55, 95% CI 1.51–14.2, based on 4
exposed case mothers) (McDonald et al., 1988), pesticide use at work (OR=1.6, 95% CI 1.3–2.1)
or home (OR=1.5, 95% CI 1.3–1.7) (Savitz et al., 1989b), occupational pesticide use in China
(lower 95% confidence limit = 4.48, p < .05; OR and upper 95% confidence limit not available
as there were 7 exposed case and 0 exposed control mothers) (Zhang et al., 1992), prenatal spraying
of pesticides on Sudanese farms (OR=3.6, 95% CI 1.6–8.0) (Taha & Gray, 1993), and maternal
occupational pesticide use during the 1st trimester (OR=2.7, 95% CI 1.5–4.8) or 2nd trimester
(OR=2.2, 95% CI 1.0–4.9) (Pastore et al., 1997). In the latter study, there was a relatively strong
association between the subgroup of stillbirths and neonatal deaths from complications of placenta
or cord and 1st trimester maternal occupational pesticide use (OR=4.8, 95% CI 2.0–11.4).
A recent Danish pregnancy cohort study revealed a statistically nonsignificant elevated risk of fetal
loss at any gestation length related to prenatal occupation as gardeners (OR=1.7, 95% CI 0.7–4.0, 5
exposed case mothers) (Zhu et al., 2006). Studies reporting no association between stillbirths and maternal pesticide exposure indices included case-control studies in the United States (occupation in agriculture, forestry or fishing, OR=0.8, 95% CI 0.5–1.3) (Savitz et al., 1989a), Colombia (employment in
semi-enclosed horticultural greenhouses, OR=0.99, 95% CI 0.66–1.48) (Restrepo et al., 1990b), and
California (1st trimester maternal residential proximity to agricultural crops, OR=1.0, 95% CI 0.9–
1.2; 1st trimester garden pesticide use, OR=0.7, 95% CI 0.6–3.1) (Pastore et al., 1997). Although
suggestive, the heterogeneity of exposure indices precludes firm conclusions.
Paternal occupational exposure, 2,4,5-T and chlorophenate wood preservatives, inadequate
evidence: See discussion of paternal occupational TCDD exposure earlier.
Paternal occupational exposure, unspecified pesticides, inadequate evidence: A retrospective
cohort study in Colombia reported no association between stillbirths and paternal employment in
semi-enclosed floriculture greenhouses in which a total of 127 different types of pesticides were used
(OR=0.87, 95% CI 0.42–1.83) (Restrepo et al., 1990b). Partners of male pesticide applicators had a
substantially increased risk of stillbirths in an Indian retrospective cohort study (crude OR=4.97, 95%
CI 3.85–6.42, calculated from data in paper); exposed men used backpacks to apply organochlorine
and other pesticides on cotton (Rupa et al., 1991). There may have been recall bias in this study as
only 2 birth defects were reported among 3016 pregnancies in the unexposed group compared to
128 among 4240 pregnancies in the exposed group. In a retrospective cohort of Norwegian farm
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families, stillbirth risk (based on official records) was not elevated (compared to nonfarm families,
RR=0.88, 95% CI 0.79–0.98) (Kristensen et al., 1997a). In a Spanish retrospective cohort study of
pregnancies conceived during the pesticide-use season (April–September), stillbirth (based on official
records) was associated with paternal occupation in agriculture (stillbirths caused by birth defects,
OR=1.62, 95% CI 1.01–2.60; other stillbirths, OR=1.35, 95% CI 1.11–1.65) (Regidor et al., 2004).
Although suggestive, the heterogeneity of exposure indices precludes firm conclusions.
Tobacco smoke Maternal active smoking, limited evidence: The U.S. Surgeon General
concluded that there is suggestive evidence of a causal relationship between stillbirth and maternal
active smoking (U.S. Department of Health and Human Services, 2001).
Maternal exposure, ETS, inadequate evidence: Stillbirths were not associated with ETS exposure at home (OR=1.1, 95% CI 0.6–2.4) or at work (OR=1.2, 95% CI 0.6–2.4) among Swedish
women (Ahlborg & Bodin, 1991). A pregnancy cohort study in California reported a statistically
nonsignificant association between stillbirth (defined as GW ≥20) and maternal serum cotinine
levels during early pregnancy (5th vs. 1st quintile, OR=3.36, 95% CI 0.81–14.0) (Kharrazi et al.,
2004). An expert panel review found sparse data on stillbirths and inadequate evidence of an association with prenatal ETS exposure (California Environmental Protection Agency, 2005).
Outdoor air pollution Maternal exposure, inadequate evidence: Daily stillbirths were weakly
associated with NO2 levels in ecologic studies in Brazil (5-pollutant model, daily stillbirths vs. NO2
(mg/m3), regression coefficient=0.0012 ± 0.0004 (Pereira et al., 1998) and the Czech Republic
(per NO2 increment of 50 µg/m3, OR=1.21, 95% CI 0.89–1.64) (Bobak & Leon, 1999). Stillbirths
in the United Kingdom were not associated with prenatal residence proximity to coke works (major
point sources of particulate air pollution) (≤2 vs. >2 km, OR=0.94, 95% CI 0.78–1.12) (Dolk et al.,
2000). A recent review concluded that the few epidemiologic studies of stillbirth provided inadequate evidence for an association with ambient air pollution (Glinianaia et al., 2004a).
Drinking water disinfection by-products Maternal exposure, limited evidence: Reviewers
found limited and fairly consistent evidence for an association between stillbirth and prenatal DBP
exposure (Bove et al., 2002; Graves et al., 2001; Nieuwenhuijsen et al., 2000). Studies in Nova
Scotia reported nonmonotonic dose-response relationships between stillbirths and drinking water
THM levels (≥100 vs. <50 µg/L, OR=1.66, 95% CI 1.09–2.52) (Dodds et al., 1999), especially
among the subset of stillbirths caused by asphyxia arising mainly from abruptio placenta (≥100 vs.
<50 µg/L, OR=4.57, 95% CI 1.93–10.8) (King et al., 2000). The association between stillbirth and
THMs was mainly related to BDCM (≥20 vs. <5 µg/L, OR=1.98, 95% CI 1.23–3.49) (King et al., 2000).
An expanded study in Nova Scotia and eastern Ontario revealed that stillbirths were associated with tap
water total THM levels (≥80 vs. <1 µg/L, OR=2.2, 95% CI 1.1–4.4) and with total THM exposure from
all sources including showering/bathing (5th vs. 1st quintile, OR=2.4, 95% CI 1.2–4.6) (Dodds et al.,
2004). Although this study observed nonmonotonic dose-response relationships, the authors noted that
the highest exposure subgroups consistently had the highest stillbirth risks. A retrospective cohort study in
England reported slightly increased stillbirth risks among women living in regions with the highest tap
water THM levels (≥60 vs. <30 µg/L, OR=1.11, 95% CI 1.00–1.23) (Toledano et al., 2005).
Drinking water nitrate Maternal exposure, inadequate evidence: A nested case-control study in
Massachusetts revealed no association between stillbirths and drinking water nitrate levels in the community of maternal residence at birth (0.3–4.5 vs. <0.2 mg/L, OR=0.8, p > .05) (Aschengrau et al., 1993).
Hazardous waste disposal sites Maternal exposure, inadequate evidence: A case-control study
within the 1988 U.S. National Maternal and Infant Health Survey reported a weak association
between stillbirth and prenatal residential proximity to U.S. EPA National Priority List2 (NPL)
hazardous waste disposal sites (maternal residence ≤1.6 km from a NPL site, OR=1.14, 95% CI
0.95–1.36) (Sosniak et al., 1994). In record-based retrospective cohort studies in the United
Kingdom, stillbirths were not associated with prenatal residential proximity to hazardous industries
(per unit change in an inverse distance squared function, OR=0.95, 95% CI 0.87–1.00) (Dummer
et al., 2003b), incinerators (per unit change in an inverse distance squared function, OR=1.04,
95% CI 0.90–1.19) (Dummer et al., 2003a), or hazardous waste landfill sites (<2 km from sites in
2
The EPA’s list of the most serious uncontrolled or abandoned hazardous waste disposal sites in the United States.
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the United Kingdom, OR=0.99, 95% CI 0.95–1.03; <2 km from landfill sites in Scotland,
OR=0.99, 95% CI 0.95–1.03 [Morris et al., 2003]).
Solvents Maternal occupational exposure, various and unspecified solvents, inadequate
evidence: A large population-based U.S. case-control study reported weak associations between
stillbirth and likely prenatal occupational exposure (inferred from job titles) to benzene (OR=1.3,
95% CI 1.0–1.8) and petroleum (OR=1.4, 95% CI 1.0–1.9) but not alcohols/glycols or chlorinated
hydrocarbons (Savitz et al., 1989a). A cohort study in Montreal revealed an association between
stillbirth and maternal 1st trimester occupational solvent exposure (OR=2.76, p < .01) (McDonald
et al., 1988). A case-control study in Brazil observed no association between stillbirth and prenatal
residential proximity to a petrochemical plant (<10 vs. ≥30 km, OR=0.71, 95% CI 0.20–2.56)
(keOliveira et al., 2002).
Paternal occupational exposure, various and unspecified solvents, inadequate evidence: A large
population-based U.S. case-control study observed no association between stillbirth and likely
paternal occupational exposure (inferred from job titles) to benzene, petroleum, alcohols/glycols or
chlorinated hydrocarbons (Savitz et al., 1989a). Fetal deaths (mainly stillbirths) were not associated
with paternal occupational solvent exposure inferred from job titles recorded on birth records in
Washington State (painting, OR=0.9, 95% CI 0.8–1.1; autobody work, OR=1.0, 95% CI 0.8–1.2;
printers, OR=1.1, 95% CI 0.8–1.3) (Daniell & Vaughan, 1988). A Norwegian cohort study of male
printers showed no association between stillbirth and occupational solvent exposure (OR=0.9,
95% CI 0.4–2.2) (Kristensen et al., 1993).
Summary There was limited epidemiologic evidence for the role of environmental toxicants in
stillbirths including prenatal exposure to DDT/DDE or DBPs.
Preterm Birth Preterm birth, defined as gestation length <37 wk, accounts for 75–85% of all
perinatal mortality in Canada (Public Health Agency of Canada, 2005). Affected infants have
increased risks of neurodevelopmental handicaps, infections, chronic respiratory disease, and
ophthalmologic problems (Health Canada, 2003). Preterm birth differs from intrauterine growth
restriction (see next section) with regard to etiology and outcome.
Lead Maternal exposure, limited evidence: Reviewers found limited and somewhat inconsistent epidemiologic evidence for an association between preterm birth and prenatal lead exposure
(Andrews et al., 1994). Preterm birth was generally associated with maternal but not paternal blood
lead levels (Figure 1). In the Port Pirie birth cohort study, there was a monotonic dose-response
relationship between preterm birth and prenatal blood lead levels with a 4.4-fold increased risk in
the highest exposure category (maternal blood lead ≥14 vs. <8 µg/dl, OR=4.4, 95% CI 1.2–17)
(McMichael et al., 1986). A nested case-control study within this cohort found somewhat higher
placental membrane lead levels among pregnancies ending in preterm birth (geometric mean lead
concentration, cases vs. controls, 1.24 µg/g [95% CI 0.91–1.67] vs. 0.78 µg/g [95% CI 0.61–1.00];
Baghurst et al., 1991). In two other birth cohort studies, preterm birth was not associated with cord
blood lead in Boston (per 1 µg/dl increment, OR=0.98, 95% CI 0.93–1.02) (Bellinger et al., 1991)
or with placental, prenatal blood, or cord blood lead levels in the former Yugoslavia (per 2nd
trimester maternal blood increment (µmol/L), OR=0.99, 95% CI 0.97–1.01) (Factor-Litvak et al.,
1991). All of these studies adjusted for prenatal smoking and other potential confounders.
In a Norwegian retrospective cohort study, preterm birth was related to likely high-level
occupational lead exposure based on job title (OR=1.9, 95% CI 1.1–3.3) (Irgens et al., 1998).
A case-cohort study in Mexico City reported a nonmonotonic dose-response relationship between
preterm birth and cord blood lead levels but only among primiparous women (cord blood lead ≥15
vs.<5.1 µg/dl, OR=2.6, 95% CI 1.0–6.7) (Torres-Sanchez et al., 1999). Gestation length was
inversely associated with placental lead concentration in a small Spanish case control study with
limited statistical analysis (Pearson’s r = –.32, p=.002) (Falcon et al., 2003). A retrospective cohort
study of occupationally exposed persons in Taiwan revealed a dose-response relationship between
preterm birth and maternal prenatal blood lead levels (maternal blood lead 10–19 vs. <10 µg/dl,
RR=1.97, 95% CI 0.92–3.86; blood lead ≥20 vs. <10 µg/dl, RR=1.86, 95% CI 0.68–4.28;
p-trend = .06) (Chen et al., 2006). In a Californian retrospective cohort study, women with maximum pregnancy blood lead levels of at least 10 µg/dl had a substantially increased risk of preterm
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Yr
Study
Grp
Compare
Huel
m
hair Pb > 14
1981
McMichael
m
BPb 08–10
1986
m
BPb 11–13
m
BPb 14+
Savitz
m
occup
1989
Bellinger
m
BPb 15+ vs. <5
1991
Factor
m
per 20 ug/dL
1991
Irgens
m
occup L
1998
m
occup M
Torres
m
BPb 15+ vs. <5
1999
Chen
m
10–19 ug/dL
2006
m
BPb 20+ vs. <10
Jelliffe
m
BPb 10+ vs. <10
2006
Berkowitz
m
airborne Pb
2006
Savitz
p
occup
1989
Irgens
p
occup H
1998
p
occup L
Lin
p
25+ ug/dL
1998
Chen
p
BPb 20+ vs. <10
2006
Odds ratio and 95% CI
0.1
0.2
0.5
1
2
5
10
FIGURE 1. Preterm birth vs. parental lead exposure (m=prenatal, p=paternal, Pb=lead, BPb=blood lead, occup=occupational).
birth (OR=3.2, 95% CI 1.2–7.4, compared to women with lower maximum levels); among women
with blood lead levels of at least 10 µg/dl, gestation length decreased by an average of 1 day per
increment of 1 µg/dl in 2nd trimester maximum maternal blood lead level (Jelliffe-Pawlowski et al.,
2006). Maternal exposures to airborne lead emissions in Shoshone County, Idaho (during a 15-mo
period when air emissions were high because of a damaged bag house), was not associated with
increased risk of preterm birth (OR=0.68, 90% CI 0.34–1.35) (Berkowitz et al., 2006).
Paternal occupational exposure, inadequate evidence: Three studies that inferred paternal lead
exposure from job histories yielded little or no evidence of an association with preterm birth (Irgens
et al., 1998; Kristensen et al., 1993; Savitz et al., 1989a). In a U.S. retrospective cohort study of
occupationally exposed men, there was a moderately strong association between preterm birth and
blood lead levels of at least 25 µg/dl for at least 5 yr (OR=3.0, 95% CI 1.4–6.8) (Lin et al., 1998).
A retrospective cohort study of persons occupationally exposed to lead in Taiwan reported no association between preterm birth and preconceptual blood lead levels of at least 20 µg/dl (OR=0.6,
95% CI 0.2–1.3) (Chen et al., 2006).
Inorganic arsenic Maternal occupational exposure, airborne arsenic, inadequate evidence:
A nation-wide case-control study in the United States found no association between preterm birth and
prenatal occupations likely exposed to airborne arsenic (OR=0.7, 95% CI 0.4–1.4) (Savitz et al., 1989a).
Maternal exposure, drinking water, inadequate evidence: Preterm births were associated with
prenatal residence in a region with well-water arsenic levels above 100 µg/L in a retrospective
cohort study in Bangladesh (RR=2.65, 95% CI 1.10–6.58, calculated from data in report) (Ahmad et al.,
2001). In an ecologic study, preterm birth was not associated with prenatal residence in regions of
Taiwan with high well water arsenic levels (OR=1.10, 95% CI 0.91–1.33) (Yang et al., 2003b).
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D. T. WIGLE ET AL.
Paternal occupational exposure, inadequate evidence: A U.S. nation-wide case-control study
observed no association between preterm birth and paternal occupations likely exposed to airborne
arsenic (OR=1.1, 95% CI 0.7–1.7) (Savitz et al., 1989a).
Cadmium Maternal exposure, inadequate evidence: A small retrospective cohort study in France
revealed a statistically nonsignificant elevated risk of preterm birth related to maternal hair cadmium
(≥0.42 vs. <0.42 µg/g, OR=2.69, 95% CI 0.53–15.0, calculated from data in paper) (Huel et al., 1981).
A birth cohort study in a lead smelter town in the former Yugoslavia reported no association between
gestation length and placental cadmium among nonsmoking women (change in gestation length per placenta cadmium increment of 1 nmol/g, 4.30 d, 95% CI –4.9 to 13.5) (Loiacono et al., 1992). A Swedish
ecologic study demonstrated no association between preterm birth and maternal residence in municipalities with soil cadmium levels exceeding 1.6 µg/g (OR=0.93, 95% CI 0.84–1.03); the risk in the
municipality with the highest soil cadmium level (12 µg/g) was not elevated (OR=0.86, 95% CI 0.61–
1.21) (Landgren, 1996). A small birth cohort study in a cadmium-polluted region in Japan showed an
association between preterm birth and prenatal urinary cadmium levels (≥2 vs. <2 µg/g creatinine,
crude OR=7.32, 95% CI 1.27–45.5, calculated from data in paper) (Nishijo et al., 2002). A small birth
cohort study in a cadmium-polluted region of China reported no association between preterm birth
among nonsmoking women and maternal or cord blood cadmium levels (cord blood cadmium >0.40
vs. ≤0.40 µg/L, OR=1.46, 95% CI 0.23–9.56, calculated from data in paper) (Zhang et al., 2004).
PCBs Maternal occupational exposure, limited evidence: Among women prenatally exposed
to airborne PCBs during capacitor production, there was an inverse dose-response relationship of borderline statistical significance between gestation length and estimated serum PCB levels (β=–1.1 d,
90% CI –2.0 to –0.1) (Taylor et al., 1989).
Maternal high-level environmental exposure, inadequate evidence: Follow-up to 1993–1994 of
women exposed to high-levels of PCBs during the 1979 Yucheng incident revealed an elevated
prevalence of stillbirths (exposed vs. unexposed, 4.2 vs. 1.7%, p=.07) (Yu et al., 2000). This finding
is consistent with an elevated risk possibly diluted by declining body burden of PCBs, PCDFs and
related toxicants over 15 yr.
Maternal low-level environmental exposure, inadequate evidence: After adjustment for the
relative concentration of docosahexaenoic acid (an n–3 polyunsaturated fatty acid in seafood) in
cord serum phospholipids, gestation length was not associated with prenatal serum PCB levels in a
small Faroe Islands birth cohort study (Grandjean et al., 2001). A small Spanish birth cohort study
revealed no association between preterm birth and mean maternal serum PCB levels (Ribas-Fito et al.,
2002). Preterm birth was weakly associated with maternal residence in zip code areas of New York State
with PCB-contaminated hazardous waste disposal sites (crude OR=1.10, 95% CI 1.08–1.11); this study
did not use exposure biomarkers and did not adjust for potential confounders (Baibergenova et al.,
2003). In a California pregnancy cohort study conducted during the 1960s (when population serum PCB
levels were substantially higher than currently) gestation length was inversely associated with prenatal
serum PCB (per natural log serum PCB increment, β=–3.9 ± 2.0 d) (Hertz-Picciotto et al., 2005). However, in a similar study with mothers recruited in 12 U.S. cities during 1959–1965, preterm birth was not
associated with prenatal serum PCB levels (≥4 vs. <2 µg/L, OR=1.11, 95% CI 0.55–2.24) (Longnecker
et al., 2005). There was also no association in cohort studies of Great Lakes fish eaters (change in gestation length per 2.7-fold maternal serum PCB increment, β=–0.08 wk [95% CI –0.75 to 0.59];
Weisskopf et al., 2005) or a representative sample of births in Victoria, Australia (preterm birth, breast
milk PCB 10–49 vs. <10 µg/kg lipid, OR=1.41, 95% CI 0.25–7.96; ≥50 vs. <10 µg/kg lipid, OR=2.30,
95% CI 0.40–13.3; p-trend=.43) (Khanjani & Sim, 2007).
TCDD Maternal exposure, inadequate evidence: Preterm birth during an 8-year follow-up of
women exposed at Seveso was not associated with maternal serum TCDD levels (per log increment, OR=1.3, 95% CI 0.7–2.3); there was also no association between gestation length and
maternal serum TCDD (per log increment, β=–1.2 d, 95% CI –2.9 to 0.5) (Eskenazi et al., 2003).
Paternal occupational exposure, inadequate evidence: Preterm birth was not associated with
exposure to potentially TCDD-contaminated chlorophenate wood preservatives among male
sawmill workers (per 100-h increment in cumulative exposure up to 3 mo before conception,
OR=1.00, 95% CI 0.99–1.001) (Dimich-Ward et al., 1996) or with paternal serum TCDD levels in
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the study of U.S. veterans exposed to Agent Orange (serum TCDD at conception ≥79 vs. ≤10 pg/g
lipid, OR=1.36, 95% CI 0.75–2.39) (Michalek et al., 1998). Similarly, preterm birth was
not related to serum TCDD levels among men exposed during production of trichlorophenol and
derivatives such as 2,4,5-T (per log serum TCDD increment, OR=0.8, 95% CI 0.6–1.1) (Lawson
et al., 2004).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: A large
ecologic study in four U.S. Midwest states revealed no relationship between preterm birth and
maternal residence in high-wheat rural counties (a proxy for chlorophenoxy herbicide exposure)
(standardized incidence ratio [SIR]=1.05, 95% CI 0.95–1.16) (Schreinemachers, 2003).
Maternal exposure, DDT/DDE, limited evidence: A large U.S. retrospective cohort study
reported a monotonic dose-response relationship between preterm birth and prenatal serum DDE
levels (≥60 vs. <15 µg/L, OR=3.1, 95% CI 1.8–5.4; p-trend=.0001); this study exploited
preserved prenatal blood samples collected during 1959–1965 when population DDT exposure
was much higher than now (Longnecker et al., 2001). A small birth cohort study in Spain revealed
an association between preterm birth and cord serum DDE levels (mean cord serum DDE, cases vs.
controls, 2.40 vs. 0.80 µg/L, p < 0.05) (Ribas-Fito et al., 2002). A case-cohort study in Mexico City
reported a nonmonotonic dose-response relationship between preterm birth and 1st trimester
maternal serum DDE levels (3rd vs. 1st tertile, OR=1.7, 95% CI 0.8–3.3, p-trend=0.17) (TorresArreola et al., 2003). A retrospective cohort study of women who consumed Great Lakes fish
showed no relationship between gestation length and maternal serum DDE levels (change in gestation length per natural log serum DDE increment, β=0.03 wk, 95% CI –0.50 to 0.57) (Weisskopf
et al., 2005). In an Australian retrospective cohort study, preterm birth was not associated with
breast milk DDT or DDE levels (2nd vs. 1st tertile DDE, OR=0.85, 95% CI 0.39–1.84; 3rd vs. 1st
tertile, OR=1.03, 95% CI 0.46–2.29) (Khanjani & Sim, 2006). A birth cohort study in the Salinas
Valley of California revealed no association between gestation length and maternal serum DDE (per
log10 serum DDE increment, β=–0.10 wk, 95% CI –0.40 to 0.20, p=.51) (Fenster et al., 2006).
Although 3 recent cohort studies found no relationship, the steep decline of population serum or
breast milk DDT/DDE levels over the past 40 yr may have contributed to negative findings.
Maternal exposure, other organochlorine insecticides, inadequate evidence: A case-cohort
study in Mexico City found a borderline dose-response relationship between preterm birth and 1st
trimester maternal serum β-hexachlorocyclohexane (β-HCH) levels (3rd vs. 1st tertile, OR=1.9, 95% CI
0.9–3.7, p-trend=.08) (Torres-Arreola et al., 2003). A California study reported an inverse dose-response
relationship between gestation length and maternal serum lipid-adjusted HCB levels (change in gestation length per natural log serum HCB increment, β=–0.47 wk, 95% CI –0.95 to –0.002, p=.05)
(Fenster et al., 2006). In an Australian retrospective cohort study, preterm birth was not associated
with breast milk dieldrin (3rd vs. 1st tertile, OR=1.22, 95% CI 0.61–2.45), heptachlor epoxide
(OR=1.02, 95% CI 0.49–2.11), oxychlordane (OR=0.93, 95% CI 0.39–2.22) or HCB (OR=1.27,
95% CI 0.54–3.00) (Khanjani & Sim, 2006).
Maternal exposure, organophosphate insecticides, inadequate evidence: A pregnancy cohort
study in New York found no association between gestation length and mean 3rd trimester maternal
urinary metabolite levels of chlorpyrifos (Berkowitz et al., 2004).
Maternal exposure, other specified insecticides or repellents, inadequate evidence: Preterm
birth was not associated with prenatal DEET exposure in a small Thai randomized trial (OR=1.00,
95% CI 0.54–1.85) (McGready et al., 2001). A pregnancy cohort study in New York found no association between gestation length and mean 3rd trimester maternal urinary pyrethroid metabolite
levels (Berkowitz et al., 2004).
Maternal exposure, fungicides, inadequate evidence: A case-cohort study in Mexico City found
no association between preterm birth and 1st trimester maternal serum HCB levels (OR=0.9, 95%
CI 0.5–1.8, p-trend=.80) (Torres-Arreola et al., 2003). A birth cohort study in the Salinas Valley of
California revealed an inverse association between gestation length and maternal serum HCB (per
log10 serum HCB increment, β=–0.47 wk, 95% CI –0.95 to –0.002, p=.05) (Fenster et al., 2006).
A pregnancy cohort study in New York found no association between gestation length and mean
3rd trimester maternal urinary pentachlorophenol levels (Berkowitz et al., 2004).
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Maternal exposure, unspecified pesticides, inadequate evidence: Preterm birth was not associated
with self-reported prenatal pesticide exposure at home (OR=1.0, 95% CI 0.7–1.4) or work (OR=1.1,
95% CI 0.6–2.1) in a large population-based case-control study (Savitz et al., 1989b). A nested case-control study reported an association between preterm birth and prenatal occupational pesticide exposure
in Colombian greenhouses (OR=1.86, 95% CI 1.59–2.17) (Restrepo et al., 1990b). A small cohort study
of mainly Hispanic pregnant women in an intense agricultural region of California revealed a lower than
expected preterm birth rate (5.6 vs. 8.9%) (Willis et al., 1993). A Danish birth cohort study showed no
association between preterm birth and prenatal occupation as gardeners (OR=1.4, 95% CI 0.8–2.4) or
farmers (OR=1.0, 95% CI 0.5–1.8) or with direct contact with pesticides at work (OR=0.7, 95% CI
0.1–5.7) (Zhu et al., 2006). The heterogeneity of exposure indices precludes firm conclusions.
Paternal occupational exposure, 2,4,5-T or chlorophenate wood preservatives, inadequate
evidence: See discussion of paternal occupational TCDD exposure earlier.
Paternal occupational exposure, other chlorophenoxy herbicides, inadequate evidence: In a
retrospective cohort study of Ontario farm families, preterm birth was associated with preconceptual paternal yard use (OR=2.5, 95% CI 0.9–7.3) but not crop use (OR=1.4, 95% CI 0.5–3.6) of
chlorophenoxy herbicides (Savitz et al., 1997a).
Paternal occupational exposure, nonchlorophenoxy herbicides, inadequate evidence: In a retrospective cohort study of Ontario farm families, preterm birth was associated with preconceptual
paternal atrazine use in yards (OR=4.9, 95% CI 1.6–15.0); there were also statistically nonsignificant elevated risks of preterm birth related to use on crops of atrazine (OR=2.4, 95% CI 0.8–7.0)
or glyphosate (OR=2.4, 95% CI 0.8–7.9) (Savitz et al., 1997a). As in most other studies, pesticidespecific risk estimates were not adjusted for exposure to other specified pesticides.
Paternal occupational exposure, unspecified pesticides, inadequate evidence: Preterm birth
was not associated with self-reported paternal pesticide exposure in the home (OR=1.0, 95%
CI 0.7–1.4), outdoors at home (OR=0.9, 95% CI 0.4–1.9) or at work (OR=1.1, 95% CI 0.7–1.8) in
a large U.S. population-based case-control study (Savitz et al., 1989b) or with agricultural work in a
large case-control study in North Carolina (OR= 0.5, 95% CI 0.2–1.4) (Savitz et al., 1997b). In a
nested case-control study, preterm birth was associated with paternal occupational pesticide use in
Colombian greenhouses (OR=2.75, 95% CI 2.01–3.76) (Restrepo et al., 1990b). A large Norwegian
retrospective cohort study found no association between preterm birth and parental employment in
farming (compared to nonfarm families, OR=0.95, 95% CI 0.91–0.99) (Kristensen et al., 1997a).
Tobacco smoke Maternal active smoking, sufficient evidence: The U.S. Surgeon General
concluded that there is suggestive evidence of a causal relationship between maternal active
smoking and preterm delivery and shortened gestation (U.S. Department of Health and Human
Services, 2004).
Maternal ETS exposure, sufficient evidence: Reviewers noted limited evidence of an association
between preterm birth and prenatal ETS exposure (Lindbohm et al., 2002) but an expert panel
recently concluded that there is sufficient evidence (California Environmental Protection Agency,
2005). Among reviewed studies, a Finnish report indicated a dose-response relationship between
preterm birth and maternal hair nicotine levels in segments corresponding to 3rd trimester exposure
(≥4.0 vs. <0.75 µg/g, OR=6.12, 95% CI 1.31–28.7; per µg/g (hair nicotine analyzed as continuous
variable), OR=1.22, 95% CI 1.07–1.39) (Jaakkola et al., 2001a), and a large cohort study of
nonsmoking Californian women noted a nonmonotonic dose-response relationship with 2nd trimester maternal serum cotinine levels (5th vs. 1st quintile, OR=1.78, 95% CI 1.01–3.13) (Kharrazi
et al., 2004). In a South African cohort study, preterm birth among nonsmoking women was not
associated with number of smokers in home (mean gestation lengths among unexposed women
and those exposed to 1 or 2+ smokers at home, respectively, were 38.4, 38.2, and 38.2 wk) (Steyn
et al., 2006). The U.S. Surgeon General reviewed eight available studies and concluded that there
was suggestive evidence of a causal relationship between prenatal ETS exposure and preterm birth
(U.S. Department of Health and Human Services, 2006).
Outdoor air pollution Maternal exposure, major ambient pollutants, limited evidence:
Reviewers noted limited evidence from studies published up to 2001 for a weak association
between preterm birth and prenatal exposure to ambient air pollutants including particulate matter
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(PM) and SO2 but no clear relationship with a critical gestational exposure period (Binkova et al.,
2005; Glinianaia et al., 2004a; Maisonet et al., 2004). Several subsequently published studies
revealed associations between preterm birth and 1st trimester maternal ambient air pollutant exposure with weaker or no association for exposure during later pregnancy. These included studies in
Lithuania (per 10 µg/m3 increment of NO2, OR=1.69, 95% CI 1.28–2.23 [in a multipollutant
model]) (Maroziene & Grazuleviciene 2002), Australia (4th vs. 1st quartile ozone level during 1st
trimester, OR=1.26, 95% CI 1.10–1.45; similar results for NO2 and SO2) (Hansen et al., 2006) and
Korea (4th vs. 1st quartile carbon monoxide (CO) level during 1st trimester, OR=1.26, 95% CI
1.11–1.44, p-trend < .001) (Leem et al., 2006).
A Vancouver study observed weak associations between preterm birth and ambient air
pollutant levels during the last month of pregnancy (e.g., per 1 ppm increment of CO,
OR = 1.08, 95% CI 1.01–1.15) but not the first month (e.g., per 1 ppm increment of CO,
OR = 0.95, 95% CI 0.89–1.01) (Liu et al., 2003). The risk of preterm birth in Taiwan was
elevated among women living close to a major freeway (<0.5 vs. 0.5–1.5 km, OR = 1.30, 95% CI
1.03–1.65) (Yang et al., 2003a). In Los Angeles, there was a dose-response relationship between
preterm birth and inverse-distance-weighted traffic density among women in their 3rd trimester
during fall-winter (OR = 1.15, 95% CI 1.05–1.26) but not spring-summer (Wilhelm & Ritz, 2003).
In further analysis of the Los Angeles study, preterm birth was associated with 1st trimester CO
concentrations near the maternal residence (women <1.6 km from monitoring station, per
1 ppm CO increment during 1st trimester, OR = 1.10, 95% CI 1.01–1.20; similar results for 3rd
trimester CO); preterm birth was not related to PM with a mass median aerodynamic diameter
< 2.5 µm (PM2.5) concentrations during early or late pregnancy (Wilhelm & Ritz, 2005). In a
California-wide case-control study, preterm birth was associated with ambient air PM2.5 levels
near the maternal residence during early or late pregnancy (4th vs. 1st quartile PM2.5 levels
during 1st gestational month, OR = 1.21, 95% CI 1.12–1.30, adjusted for CO levels; similar
results for PM2.5 during late gestation) (Huynh et al., 2006). The latter study revealed similar associations for CO levels during early or late pregnancy, independent of PM2.5. A time-series analysis
of daily preterm birth counts in 4 Pennsylvania counties during 1997–2001 revealed associations
with PM with a mass median aerodynamic diameter <10 µm (PM10) (per 50 µg/m3 increment,
OR=1.07, 95% CI 0.98–1.18) and SO2 levels (per 15 ppb increment, OR=1.15, 95% CI 1.00–1.32)
during the 6 wk before birth (Sagiv et al., 2005). In the latter study, the associations were somewhat stronger for PM10 and SO2 concentrations 2–5 d before birth, suggesting possible acute
effects of such exposure.
Maternal exposure, industrial emissions, inadequate evidence: Retrospective cohort studies in
Taiwan revealed weak associations between preterm birth and maternal residence within 3 km of a
major oil refinery (OR=1.41, 95% CI 1.08–1.82) (Lin et al., 2001a), within 2 km of a Portland
cement plant (OR=1.30, 95% CI 1.09–1.54) (Yang et al., 2003c), within 2 km of an industrial
complex including petrochemical, petroleum, steel, and shipbuilding industries (OR=1.11, 95% CI
1.02–1.21) (Tsai et al., 2003) or within 3 km of coal-based electricity-generating stations
(OR=1.14, 95% CI 1.01–1.30) (Tsai et al., 2004).
Drinking-water disinfection by-products Maternal exposure, inadequate evidence: Reviewers
found inadequate evidence for an association between preterm birth and THMs (Bove et al., 2002;
Graves et al., 2001). Among reviewed studies, there were weak associations between preterm birth
and THMs in retrospective cohort studies in Sweden (hypochlorite-treated vs. unchlorinated water,
OR=1.09, 95% CI 1.01–1.17) (Kallen & Robert, 2000) and Taiwan (chlorinated vs. unchlorinated
water supply, OR=1.37, 95% CI 1.20–1.56) (Yang, 2004) but not in Denver (Gallagher et al.,
1998), Nova Scotia (Dodds et al., 1999) and Norway (Jaakkola et al., 2001b). In a recent retrospective cohort study in Massachusetts, preterm birth was inversely associated with THM levels during
the 3rd trimester in the municipality of maternal residence (≥90th vs. <50th percentile, OR=0.88,
95% CI 0.81–0.94) but not with 3-chloro–4-(dichloromethyl)–5-hydroxy–2(5H)-furanone (MX),
mutagenic activity or haloacetic acid levels (Wright et al., 2004).
Drinking water nitrate Maternal exposure, inadequate evidence: A moderately large case-control
study in Prince Edward Island, Canada, revealed an association between preterm birth and median
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well water nitrate level in the region of maternal residence at birth (≥3.1 vs. ≤1.3 mg/L, OR=1.91,
95% CI 1.47–2.46, adjusted for prenatal smoking and several other potential confounders)
(Bukowski et al., 2001). Although suggestive, this finding requires confirmation. Groundwater nitrate
levels in rural areas may serve as proxies for pesticides and other contaminants (Ritter, 1990).
Hazardous waste sites Maternal exposure, inadequate evidence: Preterm birth was not associated with maternal residential proximity to any of 1221 NPL sites in the United States (≤1.6 vs.
>1.6 km, OR=0.99, 95% CI 0.86–1.16) (Boyle et al., 2004; Sosniak et al., 1994). In an Alaskan
study, maternal residence in villages with hazardous waste dumpsites was associated with a statistically nonsignificant increased risk of preterm birth (OR=1.24, 95% CI 0.89–1.74) (Gilbreath &
Kass, 2006b). Among studies of large single landfill sites, preterm birth risk was elevated in Los
Angeles (high-odor vs. unexposed region, mean difference in gestation length –1.8 d, p=.02)
(Kharrazi et al., 1997) and New Jersey (≤1 km and downwind vs. >1 km, high exposure period,
OR=2.10, 95% CI 1.01–4.36) (Berry & Bove, 1997) but not in Montreal (<4 km vs. unexposed
region, OR=0.97, 95% CI 0.88–1.07) (Goldberg et al., 1995). An ecologic study in Nova Scotia
revealed slightly increased risks of preterm birth in Sydney (RR=1.10, 95% CI 0.98–1.26) and the
rest of Cape Breton County (RR=1.13, 95% CI 1.04–1.22) (compared to the rest of Nova Scotia)
although the former is the site of a major hazardous waste site (Dodds & Seviour, 2001).
Solvents Maternal exposure, chlorinated solvents, inadequate evidence: A review of five epidemiologic studies published during 1990–2000 found inadequate evidence for an association
between preterm birth and prenatal residence in regions served by drinking water contaminated by
chlorinated solvents such as trichloroethylene (Bove et al., 2002). In a retrospective study at Camp
Lejeune, preterm birth was not associated with exposure to tetrachloroethylene-contaminated
drinking water (OR=1.1, 95% CI 0.9–1.3) (Sonnenfeld et al., 2001).
Maternal exposure, various and unspecified solvents, inadequate evidence: In a large U.S.
population-based case-control study, preterm birth was not associated with prenatal occupations
with likely exposure to benzene, petroleum, or alcohols/glycols (inferred from job titles) (Savitz
et al., 1989a). A small retrospective cohort study in Wisconsin observed no association between
preterm birth and prenatal occupations likely exposed to solvents (OR=1.2, 95% CI 0.4–3.1)
(Hewitt & Tellier, 1998). A very small cohort study in Canada reported an elevated risk of preterm
birth among women with self-reported 1st trimester occupational organic solvent exposure (crude
OR=3.26, 95% CI 0.78–15.7, calculated from data in paper) (Khattak et al., 1999).
Paternal occupational exposure, various and unspecified solvents, inadequate evidence: A retrospective cohort study based on Washington State birth certificates reported no association
between preterm birth and paternal employment in occupations with likely solvent exposure (e.g.,
painters, OR=1.0, 95% CI 0.8–1.3) (Daniell & Vaughan, 1988). A large U.S. population-based
case-control study reported no association between preterm birth and paternal occupations likely
exposed to benzene, petroleum or alcohols/glycols (inferred from job titles) (Savitz et al., 1989a).
Summary Epidemiologic evidence for the role of environmental toxicants in preterm birth
includes: (a) sufficient evidence—prenatal active smoking, ETS exposure; (b) limited evidence—
prenatal exposure to lead, PCBs (occupational exposure), DDT/DDE, outdoor air pollutants.
Fetal growth Deficit Intrauterine growth restriction (IUGR) is defined as a liveborn infant
below the 10th percentile of birth weight for gestational age. Other indicators of intrauterine growth
deficits include term low birth weight (birth weight <2500 g after at least 37 wk of gestation) and
low birth weight adjusted for gestation length. Fetal growth deficit (FGD) is defined here to encompass the above indicators and is associated with increased fetal and infant morbidity and mortality.
For context, prenatal smoking is the best proven preventable cause of FGD and is responsible for
30–40% of affected infants in Canada (Health Canada, 2003).
Lead Maternal exposure, limited evidence: FGD was generally associated with maternal but
not paternal blood lead levels (Figure 2). Two birth cohort studies conducted in lead smelter towns
revealed no association between FGD and prenatal blood, cord blood, or placental lead levels
(Loiacono et al., 1992; McMichael et al., 1986). A birth cohort study in Boston reported a doseresponse relationship between FGD and cord blood lead levels (relative risk increment per unit
increase in cord blood lead concentrations [µg/dl], 1.06, 95% CI 1.00–1.13) (Bellinger et al., 1991).
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
399
Study
Grp
Compare
Huel
m
hair Pb > 14
1981
Bellinger
m
BPb inc = 10
1991
Irgens
m
occup
1998
Seidler
m
occup
1999
Berkowitz
m
airborne Pb
2006
Jelliffe
m
BPb 10+
2006
Chen
m
BPb 10–19
2006
Yr
m
BPb 20+
Savitz
p
occup
1989
Kristensen
p
occup
1993
Irgens
p
occup
1998
Lin
p
BPb 25+
1998
Chen
p
BPb 10–19
2006
p
BPb 20+
Odds ratio and 95% CI
0.1 0.2
0.5
1
2
5
10
FIGURE 2. Fetal growth deficit vs. parental lead exposure (m=prenatal, p=paternal, Pb=lead, BPb=blood lead, occup=occupation).
Several recent cohort studies have consistently reported significant associations between FGD and
maternal lead exposure, including: (1) tibial bone lead (change in birth weight per unit increase in
bone lead level [µg/g] adjusted for gestation length, –7.29 ± 2.45 g, p=.003) (Gonzalez-Cossio
et al., 1997), (2) maternal occupations likely exposed to lead (OR=1.34, 95% CI 1.12–1.60) (Irgens
et al., 1998), OR=2.8, 95% CI 0.8–9.6 (Seidler et al., 1999), (3) cord blood lead (change in birth
weight per unit change in cord blood lead, –9.7 g, 95% CI –16.9 to –2.5) (Osman et al., 2000),
(4) placental lead concentration (change in birth weight per 0.1 µg/g placenta lead increment adjusted
for gestation length, –73.6 g, 95% CI –152.7 to 5.5) (note that the range of placenta lead concentrations
was 0.03–0.57 µg/g) (Odland et al., 2004) and (5) maternal blood lead level (10–19 µg/dl, OR=1.62,
95% CI 0.91–2.75; ≥20 µg/dl, OR=2.15, 95% CI 1.15–3.83, p-trend < .01 (Chen et al., 2006); maximum prenatal blood lead ≥10 vs. <10 µg/dl, OR=4.2, 95% CI 1.3–13.9 (Jelliffe-Pawlowski et al.,
2006). Maternal exposures to airborne lead emissions in Shoshone County, Idaho (during a 15-mo
period when air emissions were high because of a damaged bag house) was associated with FGD
(OR=1.92, 90% CI 1.33–2.76) (Berkowitz et al., 2006).
Paternal occupational exposure, inadequate evidence: A case-control and two retrospective
cohort studies revealed no association between FGD and paternal employment in jobs likely
exposed to lead (Irgens et al., 1998; Kristensen et al., 1993; Savitz et al., 1989a). Two retrospective
cohort studies in New York State and Taiwan showed no association between FGD and a history of
blood lead levels above 25 µg/dl for at least 5 yr before conception (OR=0.82, 95% CI 0.28–2.37)
(Lin et al., 1998) or preconceptual blood lead level of at least 20 µg/dl (OR=0.94, 95% CI 0.51–1.62)
(Chen et al., 2006).
Inorganic arsenic Maternal exposure, airborne, inadequate evidence: A U.S. case-control
study of FGD infants found no association with self-reported prenatal occupational exposure to
airborne arsenic (OR=0.8, 95% CI 0.4–1.5) (Savitz et al., 1989a).
Maternal exposure, drinking water, inadequate evidence : A birth cohort study in Chile
revealed reduced birth weight among infants of women living in a city with average drinking water
arsenic levels of 40 µg/L; birth weight adjusted for gestation length was 57 g less than that in a city
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with drinking-water arsenic levels below 1 µg/L (95% CI–123, 9) (Hopenhayn et al., 2003). However, the association between birth weight and individual tap water arsenic levels (β=–0.26 g, 95%
CI –0.85 to 0.31, per µg/L) was statistically nonsignificant.
Paternal occupational exposure, airborne, inadequate evidence: A U.S. case-control study of
FGD infants found no association with self-reported paternal occupational arsenic exposure
(OR=1.2, 95% CI 0.8–1.8) (Savitz et al., 1989a).
Cadmium Maternal exposure, inadequate evidence: In a small retrospective cohort study,
there was a statistically nonsignificant inverse association between FGD and maternal hair cadmium
(≥0.42 vs. <0.42 µg/g, OR=1.68, 95% CI 0.35–8.34); this study did not adjust for prenatal smoking (an important source of cadmium exposure and a known cause of FGD) (Huel et al., 1981).
Two birth cohort studies of nonsmoking pregnant women found no association between birth
weight adjusted for gestation length and maternal or cord blood or placental cadmium levels (Kuhnert
et al., 1987; Zhang et al., 2004). A very small Italian birth cohort study with limited statistical analysis reported an inverse correlation between birth weight of term infants and maternal blood cadmium (Pearson’s r=–0.55, p=.0003) (Salpietro et al., 2002). A small birth cohort study in a
cadmium-polluted region of Japan observed no association between birth weight and maternal
urinary cadmium levels (Nishijo et al., 2002), but there was an inverse relationship between height
at birth and 3rd trimester maternal blood cadmium (β=–0.59 ± 0.277 cm, p=.04) (Nishijo et al.,
2004). A recent Norway/Russia birth cohort study reported no association between birth weight
and maternal blood cadmium level (Pearson’s r=–0.23, p > .05); this study did not adjust for gestation length but only 2 of the 55 infants were preterm (Odland et al., 2004). Reviewers concluded
that high-dose prenatal cadmium exposure causes fetal growth deficits in experimental animals, but
there was little evidence for a relationship at the much lower exposure levels observed in humans
(Agency for Toxic Substances and Disease Registry, 1999a).
PCBs Prenatal occupational exposure, inadequate evidence: Among infants of women occupationally exposed to airborne PCBs, there was an inverse association of borderline statistical significance between birth weight adjusted for gestation length and prenatal serum PCB levels estimated from
those measured in a sub-sample of women (per 2.7-fold maternal serum PCB increment, β=–24 g,
90% CI –49 to 2) (Taylor et al., 1989). A German birth cohort study found no association between
FGD and prenatal PCB exposure inferred from a job-exposure matrix (exposed vs. unexposed,
OR=1.2, 95% CI 0.8–1.7) (Seidler et al., 1999).
Prenatal environmental exposure, inadequate evidence: Reviewers found inadequate evidence for an inverse association between birth weight and maternal exposure to background
environmental PCBs (Longnecker et al., 1997). Studies published since this review provide inconsistent evidence for an association. In a study of Swedish fishermen wives, low birth weight was
associated with prenatal serum PCB-153 levels (>400 vs. ≤400 ng/g lipid, OR = 2.3, 95% CI 0.9–
5.9), but there was no adjustment for gestation length (Rylander et al., 1998). Birth weight
adjusted for gestation length was inversely associated with cord plasma PCB levels in Holland (per
2.7-fold plasma PCB increment, β = –119.4 ± 53.7 g, p = .03) (Patandin et al., 1998). There was
an increased risk of FGD (borderline statistical significance) among Swedish fishing families in a
region where fish had relatively high PCB levels (contaminated vs. less contaminated region,
OR=1.4, 95% CI 0.9–2.1) (Rylander et al., 2000). A retrospective cohort study in New York State
found a weak but statistically significant association between low birth weight (adjusted for gestation
length and other potential confounders) and prenatal residence in regions with PCB-contaminated
hazardous waste disposal sites (OR= 1.04, 95% CI 1.02–1.07) (Baibergenova et al., 2003). In a
retrospective cohort study of Lake Michigan female anglers, birth weight adjusted for gestation
length was reduced among women in the highest serum PCB category (serum PCB 25–29 vs.
<5 µg/L, mean birth weight 2958 ± 224.0 vs. 3520 ± 103.3 g, p = .02); when analyzed by gender, the association was significant among boys but not girls (Karmaus & Zhu, 2004). Similarly, in
a California birth cohort, birth weight adjusted for gestation length was inversely related to maternal serum PCB levels among boys but not girls (per 2.7-fold serum PCB increment, respective
birth weight Z-scores for boys and girls were –0.53 ± 0.21 and 0.01 ± 0.16) (Hertz-Picciotto
et al., 2005).
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A Finnish study found no association between birth weight and breast milk PCB levels (Pearson’s
r=–.10, p=.22); analyses did not adjust for gestation length or other potential confounders
(Vartiainen et al., 1998). FGD was not associated with prenatal serum PCB levels in birth cohort
studies in the Faroe Islands (per 2.7-fold maternal serum PCB increment, β=–31.0 ± 99.9 g,
p=.76) (Grandjean et al., 2001) and Spain (per twofold cord serum PCB increment, β=–5.6 ± 36.1 g)
(Ribas-Fito et al., 2002). A U.S. Collaborative Perinatal Project cohort study revealed elevated risks
of FGD at higher maternal serum quartiles (4th vs. 1st quartile, OR=1.64, 95% CI 0.73–3.68), but
logistic regression based on PCB concentration as a continuous variable showed no association
(β=0.11 ± 0.10) (Longnecker et al., 2005). In a Japanese birth cohort, birth weight among mostly term
infants was not associated with breast milk PCB dioxin toxic equivalent (TEQ) levels (β=–5.09 ± 4.84)
(Tajimi et al., 2005). A retrospective cohort study of parents engaged in Great Lakes sport fishing
observed no association between birth weight adjusted for gestation length and maternal serum
PCB levels (change in birth weight per 2.7-fold serum PCB increment, β=29 g, 95% CI –110 to 168)
(Weisskopf et al., 2005). In a representative sample of births in Australia, FGD was not related to
breast milk PCB levels (≤50 µg/kg lipid vs. nondetectable, OR=0.87, 95% CI 0.34–2.22; >50 µg/kg
lipid, OR=0.61, 95% CI 0.22–1.66; p-trend=0.41) (Khanjani & Sim, 2007).
TCDD Maternal exposure, inadequate evidence: Low birth weight at term was not associated
with TCDD-contaminated soil (20–100 ng/g for 2+ yr or ≥100 ng/g for at least 6 mo) at or near the
prenatal residence in Missouri (OR=1.09, 95% CI 0.50–2.28) (Stockbauer et al., 1988). A birth
cohort study of women exposed to TCDD at Seveso found a statistically nonsignificant increased
risk of FGD during the first 8 yr of follow-up (per log10 maternal serum TCDD increment, OR=1.8,
95 % CI 0.7–4.3) (Eskenazi et al., 2003). In a Japanese birth cohort, birth weight among mostly term
infants was not associated with breast milk total TEQ from polychlorinated dibenzo-p-dioxins,
polychlorinated dibenzofurans, and coplanar polychlorinated biphenyls (β=–2.30 ± 2.62) (Tajimi
et al., 2005).
Paternal occupational exposure, inadequate evidence: Birth weight adjusted for gestation
length was not associated with paternal occupational exposure to chlorophenate wood preservatives known to be contaminated with TCDD and related toxicants (per 100-h increment in cumulative exposure up to 3 mo before conception, OR=1.00, 95% CI 0.99–1.001) (Dimich-Ward et al.,
1996). Among Vietnam veterans, FGD was not associated with paternal serum TCDD (≥79 vs. ≤10
pg/g lipid, OR=0.9, 95% CI 0.6–1.3) (Michalek et al., 1998). Birth weight adjusted for gestation
length was not related to serum TCDD levels among men exposed during production of trichlorophenol and derivatives such as 2,4,5-T (mean birth weight difference, TCDD ≥255 vs. <20 pg/g
lipid, 83 ± 52 g, p > .05) (Lawson et al., 2004).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: A large ecologic
study in four U.S. Midwest states revealed no relationship between FGD and maternal residence in
high-wheat rural counties (a proxy for chlorophenoxy herbicide exposure) (OR=1.05, 95% CI
0.94–1.17) (Schreinemachers, 2003).
Maternal exposure, other and unspecified herbicides, inadequate evidence: An ecologic study
in Iowa revealed associations between FGD and prenatal residence in municipalities with drinking
water supplies contaminated by the herbicides atrazine (maternal age-adjusted IUGR rate (cases
per 100 live births) vs. water atrazine level (µg/L), β=0.32, p=.001), metolachlor (β=0.26,
p=.006) or cyanazine (β=0.25, p=.009) (Munger et al., 1997). The ecologic design precludes
strong inferences from such findings. A German pregnancy cohort study revealed no association
between FGD and maternal occupational herbicide exposure (moderate vs. no exposure, OR=0.9,
95% CI 0.3–3.0, p-trend=.82) (Seidler et al., 1999).
Maternal exposure, DDT/DDE, limited evidence: There was a dose-response relationship between
FGD and prenatal serum DDE levels (≥60 vs. <15 µg/L, OR=2.6, 95% CI 1.3–5.2, p-trend=.04); this
study used preserved serum samples collected during the 1959–1966 U.S. Collaborative Perinatal
Project, a time when population-wide DDT exposure was much higher than now (Longnecker et al.,
2001). Term birth weight was inversely associated with maternal serum DDE levels in a retrospective
cohort study of women who consumed Great Lakes fish (change in birth weight per natural log serum
DDE increment, β=–146 g, 95% CI –35 to –257) (Weisskopf et al., 2005).
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Negative findings all came from relatively small studies. A small birth cohort study in Spain
revealed no association between birth weight adjusted for gestation length and cord serum DDE
levels (change in birth weight per unit change in cord serum DDE, β=–16.8 ± 37.8 g, p>0.05)
(Ribas-Fito et al., 2002). A Ukrainian pregnancy cohort study revealed increasing average birth
weight z-scores (adjusted for gestation length) among infants of women in higher breast milk DDE
tertiles (p-trend < .05) (Gladen et al., 2003). In an Australian retrospective cohort study, FGD was
not associated with breast milk DDT or DDE levels (2nd vs. 1st tertile DDE, OR=1.21, 95% CI
0.73–2.00; 3rd vs. 1st tertile, OR=0.79, 95% CI 0.45–1.39) (Khanjani & Sim, 2006). A birth cohort
study in the Salinas Valley of California revealed no association between birth weight (adjusted for
gestation length and other potential confounders) and maternal serum DDE (per log10 increment,
β=–46 g, 95% CI –129 to 37, p=.28) (Fenster et al., 2006).
Maternal exposure, other organochlorine insecticides, inadequate evidence: In an Australian
retrospective cohort study, FGD was not associated with breast milk dieldrin (3rd vs. 1st tertile,
OR=0.92, 95% CI 0.56–1.51), heptachlor epoxide (3rd vs. 1st tertile, OR=1.16, 95% CI 0.72–1.86),
or oxychlordane levels (3rd vs. 1st tertile, OR=0.93, 95% CI 0.58–1.67) (Khanjani & Sim, 2006).
A birth cohort study in the Salinas Valley of California revealed no association between birth weight
(adjusted for gestation length and other potential confounders) and maternal serum dieldrin (per
log10 increment, β=18 g, 95% CI –164 to 201, p=.84), heptachlor epoxide (per log10 increment,
β=44 g, 95% CI –105 to 194, p=.56) or oxychordane (per log10 increment, β=64 g, 95% CI –39
to 168, p=.22) (Fenster et al., 2006).
Maternal exposure, organophosphate insecticides, inadequate evidence: FGD was not associated with maternal residence in regions of San Francisco aerially sprayed with malathion during the
1st trimester (RR=0.90, 95% CI 0.54–1.49) (Thomas et al., 1992). In Mexican agricultural communities, FGD was associated with reduced cord blood acetylcholinesterase (AChE) levels (mean AChE
activity, cases vs. controls, 3.7 vs. 4.0 U/ml, p < .01) (Levario-Carrillo et al., 2004). In a New York
City birth cohort study, birth weight adjusted for gestation length was inversely associated with cord
plasma chlorpyrifos plus diazinon expressed as chlorpyrifos equivalents (mean birth weight difference, 3rd tertile vs. nondetectable, –186 g, 95% CI –327 to –45, p=.01) (Whyatt et al., 2004).
There was no association with maternal personal air chlorpyrifos plus diazinon levels over a 2-d
period during the 3rd trimester, suggesting that dietary sources may have been important.
Maternal exposure, other insecticides and repellents, inadequate evidence: In a randomized
clinical trial of prenatal DEET treatment in Thai refugee camps, the low birth weight (<2500 g) rate
was not increased among exposed (14.8%) compared to unexposed women (20.2%) (McGready
et al., 2001). The New York City study reported an inverse association of borderline statistical significance between birth weight adjusted for gestation length and cord plasma propoxur (a carbamate
insecticide) levels (mean birth weight difference, 3rd tertile vs. nondetectable, –66 g, 95% CI –147
to 15) (Whyatt et al., 2004).
Maternal exposure, fungicides, inadequate evidence: Prenatal employment in German daycare
centres with elevated indoor pentachlorophenol air concentrations was associated with reduced
birth weight adjusted for gestation length and other factors but not for maternal smoking (birth
weight difference, exposed vs. unexposed mothers, β=–217.1 ± 105.8 g, p=.04) (Karmaus &
Wolf, 1995). A small birth cohort study in Spain revealed no association between birth weight
adjusted for gestation length and other covariates and cord serum HCB (β=19.8 ± 50.9 g, p > .05)
or HCH levels (β=17.5 ± 17.6 g, p > .05) (Ribas-Fito et al., 2002). In an Australian retrospective
cohort study, FGD was not associated with breast milk HCB levels ( 2nd vs. 1st tertile, OR=0.89,
95% CI 0.54–1.48; 3rd vs. 1st tertile, OR=0.98, 95% CI 0.58–1.67) (Khanjani & Sim, 2006).
A birth cohort study in the Salinas Valley of California revealed no association between birth weight
(adjusted for gestation length and other potential confounders) and maternal serum HCB (per log10
increment, β=–23g, 95% CI –154 to 108, p=.73) or β-HCH (per log10 increment, β=25 g, 95%
CI –154 to 108, p=.73) (Fenster et al., 2006).
Maternal exposure, unspecified pesticides, inadequate evidence: A large U.S. case-control
study reported an association between FGD and self-reported prenatal pesticide exposure at home
(OR=1.5, 95% CI 1.1–2.1) but not occupational exposure (OR=1.2, 95% CI 0.6–2.3) (Savitz et al.,
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403
1989b). In a Shanghai case-control study, FGD was associated with self-reported periconceptual
maternal occupational pesticide exposure (OR=2.9, 95% CI 1.0–8.6) (Zhang et al., 1992). In
Mexican agricultural communities, FGD was associated with self-reported prenatal residential
proximity to crop areas with intense pesticide use (OR=2.3, 95% CI 1.0–5.3) (Levario-Carrillo
et al., 2004). A Danish birth cohort study reported no association between FGD and prenatal occupation as gardeners (OR=1.0, 95% CI 0.6–1.6) or farmers (OR=0.6, 95% CI 0.3–1.0) or with use
of pesticides at home or work (OR=0.5, 95% CI 0.1–2.9) (Zhu et al., 2006). The heterogeneity of
exposure indices precludes firm conclusions.
Paternal occupational exposure, 2,4,5-T and chlorophenate wood preservatives, inadequate
evidence: See discussion of paternal occupational TCDD exposure earlier.
Paternal occupational exposure, other chlorophenoxy herbicides, inadequate evidence:
A retrospective cohort study of Ontario farm families found no association between FGD and preconceptual paternal crop (OR=0.7, 95% CI 0.4–1.2) or yard (OR=0.8, 95% CI 0.4–1.6) chlorophenoxy herbicide use; however, there was an association of borderline statistical significance for
users of yard herbicides without protective equipment (OR=1.8, 95% CI 0.9–3.4) (Savitz et al.,
1997a).
Paternal occupational exposure, unspecified pesticides, inadequate evidence: A large U.S.
case-control study reported an association between FGD and self-reported paternal pesticide exposure at home (OR=1.4, 95% CI 0.9–2.3) or work (OR=1.5, 95% CI 1.1–2.0) (Savitz et al., 1989b).
A large Norwegian cohort study found a slightly reduced FGD risk among infants of farm operators
compared to infants of nonfarmers in agricultural municipalities (RR=0.90, 95% CI 0.88–0.93)
(Kristensen et al., 1997a).
Tobacco smoke Maternal active smoking, sufficient evidence: The U.S. Surgeon General
concluded that there is sufficient evidence of a causal relationship between prenatal active smoking
and FGD (U.S. Department of Health and Human Services, 2004).
Maternal ETS exposure, limited evidence: Reviewers (Lindbohm et al., 2002) and an expert
panel (California Environmental Protection Agency, 2005) noted limited evidence of an association
between FGD and prenatal ETS exposure. The U.S. Surgeon General reviewed 46 available studies
and concluded that there was sufficient evidence of a causal relationship between prenatal ETS
exposure and a small reduction in birth weight but did not clearly distinguish between reduced
birth weight from preterm birth and that from FGD (U.S. Department of Health and Human Services, 2006). Among nonsmoking pregnant women, FGD was associated with self-reported prenatal
ETS exposure in studies in the Czech Republic (OR=1.19, 95% CI 0.96–1.47) (Dejmek et al.,
2002), China (mean birth weight difference, adjusted for gestation length, exposed vs. unexposed
women, –37 g, 95% CI –82.6 to 8.4) (Ha et al., 2002), India (OR=2.10, 95% CI 1.27–3.48) (Goel
et al., 2004), Poland (change in birth weight per unit increase in log 2nd trimester maternal serum
cotinine (over the range 0–9 ng/ml), β=–100.5 ± 60.4 g, p=.09) (Hanke et al., 2004), and California
(per log 2nd trimester maternal serum cotinine increment, OR=1.41, 95% CI 0.91–2.17) (Kharrazi
et al., 2004). In a Korean cohort study of nonsmoking women, birth weight at term was inversely
associated with ETS exposure (confirmed by urinary cotinine measurements) among the subgroup
of women with the GSTT1 null type polymorphism (mean birth weight difference, exposed vs.
unexposed, –236 g, 95% CI –455 to –17) but not among those with the GSTT1 wild type or GSTM1
null or wild types (Hong et al., 2003).
Among women who smoked during part or all of pregnancy, FGD was associated with
self-reported ETS exposure in the Czech Republic (≤10 cigarettes/d, OR=2.14, 95% CI 1.67–2.73;
>10 cigarettes/d, OR=3.43, 95% CI 2.19–5.36) (Dejmek et al., 2002) and Sweden (1st trimester
exposure, OR=2.60, 95% CI 0.99–6.86) (Dejin-Karlsson & Ostergren, 2003). The latter study
observed no association with ETS exposure during the 2nd and 3rd trimesters, adjusted for 1st
trimester exposure (OR=1.27, 95% CI 0.64–2.49).
In other studies, FGD was not associated with self-reported prenatal ETS exposure at home or
work (Hruba and Kachlik 2000; Jedrychowski et al., 2004; Matsubara et al., 2000; Perera et al.,
2004; Windham et al., 2000), maternal hair nicotine levels in segments corresponding to 3rd
trimester exposure (Jaakkola et al., 2001a), or postpartum maternal plasma cotinine (β=–0.02,
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p=.42) (Perera et al., 2004). In a South African cohort study, FGD among nonsmoking women was
not associated with self-reported ETS exposure at home (Steyn et al., 2006). In a Spanish cohort
study, birth weight was reduced among women with elevated hair nicotine concentrations (≥18 vs.
<3 ng/mg, birth weight difference and SD = –247.1 ± 118.6 g) but this study did not adjust for
maternal active smoking (Pichini et al., 2003).
Outdoor air pollution Maternal exposure, major ambient air pollutants, limited evidence:
Reviewers noted limited evidence in studies published up to 2001 for an association between
FGD and ambient air pollutant levels (Binkova et al., 2005; Glinianaia et al., 2004a; Maisonet
et al., 2004). In recently published studies, associations were observed in Lithuania (1st trimester
formaldehyde, 3rd vs. 1st tertile, OR=2.39, 95% CI 1.07–5.32; no association with formaldehyde
levels during later pregnancy) (Maroziene & Grazuleviciene, 2002), Vancouver (per 5 ppb increment of 1st trimester SO2, OR=1.07, 95% CI 1.00–1.14; similar results for NO2 and CO) (Liu
et al., 2003), Los Angeles County (women in 3rd trimester during fall-winter months, 5th vs. 1st
quintile of distance-weighted traffic density, OR=1.39, 95% CI 1.16–1.67, p-trend=.004)
(Wilhelm & Ritz, 2003), South Korea (per interquartile SO2 increment, OR = 1.06, 95%
CI 1.02–1.11) (Lee et al., 2003), Taiwan (birth weight deficit per 1 µg/m3 SO2 increment during 1st
trimester, 0.52 g, 95% CI 0.09–2.63; similar results for PM10; no association with levels during later
pregnancy) (Yang et al., 2003), Brazil (birth weight deficit per 10 µg/m3 PM10 increment during 1st
trimester, 14 g, 95% CI 0.4–27; per 1 ppm CO increment during 1st trimester, 23 g, 95%
CI 4.9–41; no association with pollutant levels during later pregnancy) (Gouveia et al., 2004), Los
Angeles (per 10 µg/m3 PM10 increment during 1st trimester, multipollutant model OR=1.36, 95%
CI 1.12–1.65; per 1 ppm CO increment, OR=1.15, 95% CI 0.98–1.35) (Wilhelm and Ritz 2005),
and Nova Scotia, Canada (4th vs. 1st quartile SO2 during 1st trimester, OR = 1.36, 95%
CI 1.04–1.78; similar results for PM10 but no association with any pollutant during later pregnancy) (Dugandzic et al., 2006). FGD was not associated with ambient air pollution in a Nevada
study (3rd trimester PM10 >45 vs. <20 µg/m3, OR=1.11, 95% CI 0.71–1.71; similar findings for
CO and ozone) (Chen et al., 2002a).
Maternal exposure, industrial air pollutants, inadequate evidence: A small retrospective cohort
study in Taiwan reported an association between term low birth weight and prenatal residence
within about 3 km of petrochemical manufacturing facilities (exposed vs. unexposed women,
OR=1.77, 95% CI 1.00–3.12) (Lin et al., 2001b), but a larger study found no association with
prenatal residence in several cities with petrochemical industries (exposed vs. unexposed women,
OR=1.07, 95% CI 0.95–1.22) (Yang et al., 2002).
Drinking water disinfection by-products Maternal exposure, limited evidence: Reviewers
found limited evidence for an association between FGD and indices of maternal DBP exposure
(Bove et al., 2002; Graves et al., 2001). Among reviewed studies, there were dose-response relationships between FGD and prenatal community drinking water THM levels in Iowa (e.g., chloroform 1–9 µg/L vs. nondetectable, OR=1.3, 95% CI 0.9–1.8; ≥10 µg/L, OR=1.8, 95% CI 1.1–2.9)
(Kramer et al., 1992) and New Jersey (total THM >100 vs. ≤20 µg/L, OR=1.50, 95% CI 1.04–2.09,
p-trend<0.05) (Bove et al., 1995). In a retrospective cohort study in Denver, there was a relatively
strong association between term low birth weight and prenatal drinking water THM levels above
60 µg/L (relative to ≤20 µg/L, OR=5.9, 95% CI 2.0–17) (Gallagher et al., 1998). A retrospective
cohort study in Massachusetts found associations between FGD and 3rd trimester municipal drinking water DBP indices above the 90th compared to less than 50th percentile concentrations (total
THM, OR=1.13, 955 CI 1.07–1.20; MX, OR=1.14, 95% CI 0.95–1.37; mutagenic activity,
OR=1.25, 95% CI 1.04–1.51) but not HAAs (OR=0.97, 95% CI 0.77–1.23) (Wright et al., 2004).
In a retrospective cohort study in Arizona, there were associations between FGD and 3rd trimester
municipal drinking-water total THMs (3rd vs. 1st tertile total THM, OR=1.09, 95% CI 1.00–1.18)
and HAAs (3rd vs. 1st tertile total HAAs, OR=1.25, 95% CI 0.96–1.64), especially dibromoacetic
acid (OR=1.49, 95% CI 1.09–2.04) (Hinckley et al., 2005).
Other studies reported weak or no association between FGD and measured THM levels in
Nova Scotia (3rd trimester THM ≥100 vs. <50 µg/L, OR=1.08, 95% CI 0.99–1.18) (Dodds et al.,
1999), prenatal residence in communities using chlorinated versus unchlorinated water in Sweden
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
405
(OR=1.07, 95% CI 0.96–1.19) (Kallen & Robert, 2000), Taiwan (OR=0.90, 95% CI 0.75–1.09)
(Yang et al., 2000), and Norway (OR=1.00, 95% CI 0.91–1.10) (Jaakkola et al., 2001b) or swimming
pool use in the United Kingdom (≥2 vs. 0 h/wk, birth weight difference 17 g, 95% CI –11 to 45)
(Nieuwenhuijsen et al., 2002).
Drinking water nitrate Maternal exposure, limited evidence: A moderately large case-control
study in Prince Edward Island, Canada, revealed an association between FGD and median well
water nitrate level in the region of maternal residence at birth (≥3.1 vs. ≤1.3 mg/L, OR=2.40, 95%
CI 1.75–3.27, adjusted for prenatal smoking and several other potential confounders) (Bukowski
et al., 2001).
Solvents Maternal exposure, trichloroethylene or tetrachloroethylene in drinking water, inadequate evidence: Reviewers noted that two of three studies found increased risks of FGD related to
prenatal residence in regions with trichloroethylene- and tetrachloroethylene-contaminated drinking
water (Bove et al., 2002). The positive studies were conducted at Camp Lejeune (main contaminant
was tetrachloroethylene, OR=1.2, 95% CI 1.0–1.3) (Sonnenfeld et al., 2001) and Woburn (main
contaminants were trichloroethylene and tetrachloroethylene). A peer review panel concluded that
studies in Woburn did not support an association between adverse pregnancy outcomes and drinking
water contamination (Massachusetts Department of Public Health, 1998).
Prenatal exposure, other and unspecified solvents, inadequate evidence: In a large U.S. populationbased case-control study, FGD was not associated with prenatal occupations likely exposed
to benzene, petroleum or alcohols/glycols (inferred from job titles) (Savitz et al., 1989a). A small
hospital-based case-control study in California reported no association with self-reported maternal
1st trimester occupational solvent exposure (OR=1.4, 95% CI 0.73–2.6) (Windham et al., 1991).
A retrospective cohort study in New Jersey reported an association between FGD and drinkingwater carbon tetrachloride levels in water systems serving the prenatal residence (>1 vs. ≤1 µg/L,
OR=1.75, 95% CI 1.13–2.70) (Bove et al., 1995). There was a statistically nonsignificant increased
risk of FGD among women with prenatal occupational solvent exposure in a retrospective cohort
study in Germany (OR=2.2, 95% CI 0.8–6.1, p-trend=0.13) (Seidler et al., 1999). In a retrospective cohort study of Swedish female biomedical research laboratory employees, mean birth weight
was reduced among those exposed to diethylether (compared to unexposed women, mean difference –155 g, 95% CI –356 to 46) but not acetone, chloroform or toluene (Wennborg et al., 2000).
A retrospective cohort study in China reported lower birth weight among nonsmoking women
occupationally exposed to aromatic solvents including benzene, toluene, styrene, and xylene
(adjusted for gestation length, compared to unexposed women, birth weight difference –79.0 g,
95% CI –156.0 to –1.9) (Ha et al., 2002). In Singapore, term low birth weight was not associated
with maternal employment as cleaners or related jobs with likely solvent exposure (Chia et al.,
2004).
Paternal occupational exposure, unspecified solvents, inadequate evidence: A retrospective
cohort study based on Washington State birth certificates reported an association between FGD
and paternal employment in one of four occupations with likely solvent exposure (painters,
OR=1.9, 95% CI 1.0–3.4) (Daniell & Vaughan, 1988). In a large U.S. population-based case-control
study, FGD was associated with paternal occupations likely exposed to benzene (OR=1.5, 95% CI
1.1–2.3) but not petroleum or alcohols/glycols (inferred from job titles) (Savitz et al., 1989a).
A Norwegian retrospective cohort study of male printers observed no association between FGD and
paternal occupational exposure to solvents only (OR=1.0, 95% CI 0.75–1.3) or to solvents and
lead (OR=0.9, 95% CI 0.64–1.2) (Kristensen et al., 1993). A retrospective cohort study in China
reported no association between birth weight (adjusted for gestation length) among offspring of
men occupationally exposed to aromatic solvents (Ha et al., 2002). In Singapore, term low birth
weight was associated with paternal employment as cleaners or related jobs with likely solvent
exposure (OR=1.32, 95% CI 1.12–1.55); the crude exposure index precludes strong inferences
(Chia et al., 2004).
Summary There is sufficient epidemiologic evidence that prenatal active smoking causes FGD;
limited epidemiologic evidence supports associations between FGD and prenatal exposure to lead,
DDT/DDE, ETS, outdoor air pollution and drinking water DBPs and nitrate.
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D. T. WIGLE ET AL.
Birth Defects Birth defects are defined as physical or biochemical defects (e.g., cleft palate,
phenylketonuria) that are present at birth and may be inherited or environmentally induced. About
2–3% of liveborn infants have one or more birth defects. In addition to true differences in risk,
reported birth defect prevalence rates may vary because of differences in rates of 1st trimester
screening and pregnancy termination and variable diagnosis and reporting. They are the leading
cause of infant deaths and can cause lifelong disability but their causes remain largely unknown
(Arias & Smith, 2003). During the late 20th century, birth prevalence rates in the Atlanta birth
defect monitoring system declined for several types of birth defects (especially neural tube defects
[NTDs]) but increased for coarctation of the aorta, hypospadias (a urethral closure defect on the
ventral surface of the penis), and cystic kidney (International Clearinghouse for Birth Defect Monitoring
Systems, 2000). There are over 100 distinct types of birth defect recognized, but most registries only
track the most common types. For instance, the Metropolitan Atlanta Congenital Defects Program
tracks 44 distinct birth defect categories (Centers for Disease Control and Prevention, 2005a).
The following discussion of potential environmental causes of birth defects focuses on specific
major birth defect types but epidemiologic studies with limited statistical power have often assessed
the overall risk of birth defects in relation to environmental exposure indices. Epidemiologic studies
of total birth defects are problematic because of the likely substantial etiologic heterogeneity of
diverse types of birth defects; such studies may obscure true relationships with specific birth
defects. The level of epidemiologic evidence for associations between major birth defects and environmental factors is summarized in Table 2.
Neural Tube Birth Defects
Lead Maternal exposure, inadequate evidence: NTDs were not associated with average
municipal drinking-water lead levels in communities of maternal residence in a Canadian ecologic
study (mean levels in case and control communities were 10.3 and 11.5 µg/L) (Elwood & Coldman,
1981) but were associated with the percent of homes with tap water lead levels exceeding 10 µg/L
in an English ecologic study (Bound et al., 1997). A Massachusetts case-control study reported no
association between NTDs and drinking water lead levels in the community of maternal residence
at delivery (≥1 vs. <1 µg/L, OR=0.8, p > .05); the use of such a low cutoff point greatly reduces
the chance of observing an association if it exists (Aschengrau et al., 1993). In a population-based
case-control study in California, NTDs were associated with periconceptual maternal residential
proximity to NPL sites containing lead (OR=2.0, 95% CI 0.9–4.1) (Croen et al., 1997). A Norwegian
retrospective cohort study observed an association between NTDs and prenatal occupations likely
exposed to airborne inorganic lead (inferred from job titles) (OR=2.9, 95% CI 1.1–6.4) but this
study had no information on blood lead levels (Irgens et al., 1998). A small study in Texas reported
higher mean amniotic fluid lead levels among NTD cases compared to controls (12.0 ± 0.6 vs. 5.7 ± 0.1
µmol/L, p < .0001) (Dawson et al., 1999). There was no association between NTDs and maternal
blood lead levels above 6 µg/dl (OR=1.5, 95% CI 0.6–4.3) in a large Texan case-control study
(Brender et al., 2006).
Paternal occupational exposure, inadequate evidence: A Norwegian retrospective cohort study
and a Texan case-control study, respectively, found no association between NTDs and paternal
occupational lead exposure (inferred from job titles) (OR=1.0, 95% CI 0.7–1.4) (Irgens et al., 1998)
or self-reported paternal occupational lead exposure (OR=1.3, 95% CI 0.8–2.3) (Brender et al., 2002).
Arsenic, cadmium, mercury Maternal exposure, inadequate evidence: Case-control studies in
Canada and Massachusetts reported no association between community drinking water cadmium
levels and risk of anencephaly (mean concentration, case vs. control communities, 8.97 vs. 9.00 µg/L)
(Elwood & Coldman, 1981) or the broader category of central nervous system (CNS) birth defects
(>1 vs. ≤1 µg/L, OR=1.6, p > .05) (Aschengrau et al., 1993). The Canadian study also revealed no
association with drinking water mercury levels (mean concentration, case vs. control communities,
7.09 vs. 7.95 µg/L) (Elwood & Coldman, 1981). In a large Texan case-control study, there was no
association between NTDs and biomarker levels exceeding the 95th percentile for arsenic in the
third U.S. National Health and Nutrition Examination Survey (NHANES III) (1999–2000) (maternal
urinary arsenic ≥38.8 µg/L, OR=0.0, 95% CI 0.0–2.0, 0/56 exposed cases), cadmium (urinary
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
407
TABLE 2. Role of Environmental Toxicants in Birth Defects
Toxicant
Exposure
Neural tube
Cardiac
Orofacial
Lead
Prenatal
Paternal
Prenatal
Paternal
Prenatal
Prenatal
Prenatal
Paternal
Prenatal
I
I
I
I
I
I
I
I
I
I
I
I
I
Arsenic
Mercury
Cadmium
Mixed metals
PCBs
TCDD
2,4,5-T, chlorophenate
wood preservatives
Chlorophenoxy herbicides
(other than 2,4,5-T)
Other or unspecified
herbicides
DDT/DDE
Organophosphate
insecticides
Other or unspecified
insecticides
Fungicides
Unspecified pesticides
Active smoking
ETS
Outdoor air pollutionh
Drinking water DBPs
Drinking water
nitrate, nitrite
Hazardous waste disposal
sites
Incinerators
Chlorinated solvents
Glycol ethers
Unspecified solvents
I
I
Prenatal
Paternal
Paternal
I
L
I
I
I
I
Prenatal
I
I
I
I (amides,
glyphosate)
I
Paternal
Prenatal
Prenatal
Prenatal
Urinary tract
Male genital
I
Yucheng—Sg
Environ—I
I
I
I
I
Musculo-skeletal
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
Prenatal
I
I
I
I
Prenatal
Prenatal
Paternal
Prenatal
Prenatal
Prenatal
Paternal
Prenatal
Prenatal
I
I
I
I
I
I
I
I
I
L
L
L
I
I
I
L
I
I
I
I
Prenatal
L
L
I
I
Prenatal
Prenatal
Prenatal
Prenatal
Paternal
I
I
L
I
L
I
I
L
L
I
I
I
I
I
I
I
I
I
L
I
L
I
I
I
I
Phthalates
I
I
I
I
I
I
L
I
I
I
Ii
Note. TCDD, 2,3,7,8-tetrachlorodibenzo-p-dioxin.
g
This category refers to developmental tooth abnormalities.
h
Major pollutants from fossil fuel combustion.
i
Reduced anogenital index in male infants, an indicator of incomplete masculinization.
cadmium ≥1.24 µg/L, OR=1.0, 95% CI 0.1–17.0, 1/24 exposed cases), or mercury (urinary
mercury ≥5.62 µg/L, OR=1.8, 95% CI 0.8–3.7) (Brender et al., 2006). Reviewers noted sparse and
inadequate evidence for an association between NTDs and drinking-water arsenic exposure in
humans, a conclusion that remains applicable almost a decade later (DeSesso et al., 1998).
PCBs Maternal exposure, inadequate evidence: In a population-based case-control study in
California, NTDs were weakly associated with periconceptual maternal residential proximity to NPL
sites containing PCBs (OR=3.5, 95% CI 0.9–10.6) (Croen et al., 1997). A case-control study of
Mexican-American women in Texas found no association between NTDs and individual or summed
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D. T. WIGLE ET AL.
PCB congener concentrations in maternal serum (sum of 7 PCB congeners >32 ng/g lipid vs.
<LOD, OR=0.7, 95% CI 0.3–1.6) (Suarez et al., 2005).
TCDD Maternal exposure, inadequate evidence: A retrospective cohort study of births in
French communities with solid waste incinerators and unexposed comparison communities
reported no association between NTDs and expert-rated hazard of incinerator emissions (high vs.
low exposure, OR=0.83, 95% CI 0.35–1.96) (Cordier et al., 2004).
Paternal occupational exposure, limited evidence: A case-control study in Atlanta reported no
association between anencephaly or spina bifida and self-reported paternal exposure to Agent
Orange in Vietnam (respective ORs=0.80 and 1.19, CIs not stated) (Erickson et al., 1984). In a retrospective cohort study of Vietnam veterans and partners, there was an elevated risk of CNS defects
of borderline statistical significance among offspring of men with low or high TCDD categories
(defined earlier) compared to offspring of unexposed men (OR=4.18, 95% CI 0.96–21.3, calculated from data in paper, 5 exposed case fathers) (Wolfe et al., 1995). Among male sawmill workers
in British Columbia, NTDs were associated with maximum preconceptual chlorophenate exposure
intensity (h/yr) (75th vs. 25th percentile, OR=2.35, 95% CI 1.1–5.3) and less strongly with cumulative exposure (h) during the 3 mo before conception (75th vs. 25th percentile, OR=1.27, 95% CI
0.8–2.0) (Dimich-Ward et al., 1996). The latter findings are consistent with a role for cumulative
exposure to PCDD and PCDF contaminants that bioaccumulate in body lipids. Reviewers noted
inadequate epidemiologic evidence for an association between NTDs and paternal TCDD exposure
(Longnecker et al., 1997). A subsequent review concluded that there was limited epidemiologic
evidence for an association between spina bifida and paternal exposure to phenoxy herbicides
potentially contaminated by TCDD (National Academy of Sciences, 2003).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: A retrospective cohort study in four U.S. states reported no association between CNS birth defects and prenatal
residence in high-wheat counties (a proxy for agricultural use of chlorophenoxy herbicides including
2,4-D and MCPA) (high- vs. low-wheat counties, OR=0.81, 95% CI 0.46–1.42) (Schreinemachers,
2003). A pooled analysis of 2 case-control studies in California revealed a statistically nonsignificant
elevated NTD risk related to prenatal residence less than 1 km from areas with documented agricultural
use of 2,4-D or derivatives (OR=1.5, 95% CI 0.8–2.7, in a multipesticide model) (Rull et al., 2006).
Maternal exposure, other herbicides, inadequate evidence: In the California pooled analysis,
NTDs were associated with prenatal residential proximity to agricultural applications of amide
herbicides (OR=2.2, 95% CI 1.0–5.3) and a statistically nonsignificant association with glyphosate
use (OR=1.5, 95% CI 0.8–2.9) (Rull et al., 2006).
Maternal exposure, insecticides, inadequate evidence: In the California pooled analysis, NTDs
were not associated with prenatal residential proximity to agricultural applications of halogenated
organic insecticides (OR=0.9, 95% CI 0.6–1.3) but elevated NTD risks were related to organophosphate (OR=1.3, 95% CI 0.9–1.8), methyl carbamate insecticides (OR=1.5, 95% CI 1.0–2.3)
and specific carbamates including carbaryl (OR=1.7, 95% CI 0.8–3.9) and methomyl (OR=1.4,
95% CI 0.8–2.5) (Rull et al., 2006). This study also reported statistically nonsignificant associations
between NTDs and specific organophosphate insecticides including naled (OR=2.7, 95% CI 0.9–8.2)
and oxydemeton-methyl (OR=3.4, 95% CI 0.8–14.3) but not with malathion (OR=1.0, 95%
CI 0.4–2.7) or chlorpyrifos (OR=1.3, 95% CI 0.7–2.3).
Maternal exposure, fungicides, inadequate evidence: In the California pooled analysis, NTDs
were associated with prenatal residential proximity to agricultural applications of fungicides in the
benzimadazole chemical class (OR=2.2, 95% CI 1.1–4.7) but not with dicarboximides (OR=1.1,
95% CI 0.6–2.1) or dithiocarbamates (OR=0.7, 95% CI 0.3–1.5) (Rull et al., 2006). This study
observed an association of borderline statistical significance with the benzimidazole fungicide
benomyl (OR=2.3, 95% CI 0.9–5.6).
Maternal exposure, unspecified pesticides, inadequate evidence: Reviewers noted limited
epidemiologic evidence for an association between NTDs (anencephaly, spina bifida) and prenatal
occupational pesticide exposure (Sever et al., 1997). Among Norwegian farm families, CNS birth
defect risk was not elevated (compared to nonfarm families, OR=0.94, 95% CI 0.73–1.20) but was
associated with farm use of tractor pesticide spraying equipment for orchards or greenhouses
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409
(OR=2.30, 95% CI 1.31–4.04) (Kristensen et al., 1997b). Combined analysis of two case-control
studies in California revealed an association between NTDs and maternal residential proximity to
NPL sites containing pesticides (<1 vs. ≥1 mile, OR=2.2, 95% CI 0.9–5.2) (Croen et al., 1997).
A California study revealed associations between NTDs and prenatal professional pesticide use in
homes (OR=1.6, 95% CI 1.1–2.5) and maternal use of garden pesticides (OR=2.9, 95% CI 1.3–6.7)
but not with pet flea treatments (Shaw et al., 1999). Two reports of a Mexican case-control study of
anencephaly revealed strong associations with self-reported periconceptual maternal occupation in
agriculture; in the more recent report, the odds ratio was 4.57 (95% CI 1.05–20.0) (Blanco Munoz
et al., 2005).
A Finnish case-control study found no association between NTDs and maternal 1st trimester
work in agriculture (OR=1.2, 95% CI 0.6–2.4); among the 38 exposed case mothers, half were
considered lightly exposed (Nurminen et al., 1995). NTDs were associated with prenatal occupation in agriculture in a case-control study in the Netherlands (OR=3.4, 95% CI 1.3–9.0); however,
among the few women employed in agriculture, similar proportions of case (4/9) and control (5/10)
mothers reported pesticide exposure (Blatter et al., 1996a). NTDs were not associated with
self-reported periconceptual maternal occupational pesticide exposure in case-control studies in
California (OR=0.9, 95% CI 0.2–3.8) (Shaw et al., 1999) or Texas (OR=1.2, 95% CI 0.3–4.8)
(Brender et al., 2002). In a California case-control study, CNS defects were not associated with prenatal residence in a census tract with a NPL site containing pesticides (OR=1.02, 95% CI 0.68–1.55)
(Orr et al., 2002). The heterogeneous and nonspecific exposure indices preclude strong inferences.
Paternal occupational exposure, 2,4,5-T and chlorophenate wood preservatives, inadequate
evidence: See paternal occupational TCDD exposure described earlier.
Paternal occupational exposure, unspecified pesticides, inadequate evidence: Reviewers found
limited epidemiologic evidence for an association between NTDs and paternal occupational pesticide exposure (Sever et al., 1997). Among studies published before the latter review, CNS birth
defects were associated with farm use of tractor pesticide spraying equipment for orchards or
greenhouses (OR=2.30, 95% CI 1.31–4.04) (Kristensen et al., 1997b). Statistically nonsignificant
elevated risks were observed in Texas (anencephaly, paternal occupations likely exposed to pesticides, OR=1.28, 95% CI 0.77–2.13; occupation as farm or ranch workers, OR=1.73, 95% CI
0.84–3.55) (Brender & Suarez, 1990), Minnesota (CNS defects, paternal occupation as licensed
pesticide applicator, RR=1.10, 95% CI 0.50–2.40, 6 exposed case fathers) (Garry et al., 1996), and
the Netherlands (spina bifida, moderate to high vs. no periconceptual paternal occupational
pesticide exposure, OR=1.7, 95% CI 0.7–4.0) (Blatter et al., 1997). The Minnesota study also
noted increased risks of NTDs among nonfarm families (neither parent was a licensed pesticide
applicator) living in agricultural regions (e.g., residence in corn/soybean vs. noncrop regions,
OR=1.42, 95% CI 1.09–1.86; wheat/sugar beet/potato regions, OR=1.49, 95% CI 0.92–2.40);
the crude exposure indicator and inability to adjust for potential confounders preclude strong
inferences (Garry et al., 1996).
Among recently reported studies, a country-wide Norwegian retrospective cohort study
reported no association between NTDs and paternal occupation in agriculture (based on census
records) (OR=0.91, 95% CI 0.67–1.21) (Irgens et al., 2000). NTDs were not associated with
self-reported paternal periconceptual occupational pesticide exposure in case-control studies in
California (OR=0.9, 95% CI 0.5–1.7) (Shaw et al., 1999) or Texas (OR=1.2, 95% CI 0.5–2.8)
(Brender et al., 2002). Anencephaly risk was elevated (but not statistically significant) among
offspring of men with partner-reported periconceptual occupation in agriculture in Mexico
(OR=2.50, 95% CI 0.73–8.64) (Lacasana et al., 2006). The inconsistent findings and the heterogeneity of exposure indices precludes firm conclusions.
Outdoor air pollution Maternal exposure, inadequate evidence: A retrospective cohort study
in France observed no association between NTDs and traffic density near the prenatal residence
(>50,000 vs. <10,000 vehicles/d, OR=1.05, 95% CI 0.44–2.51) (Cordier et al., 2004).
Paternal occupational exposure, inadequate evidence: A large Norwegian retrospective cohort
study reported no association between NTDs and paternal occupation as drivers (OR=1.07, 95%
CI 0.80–1.43) (Irgens et al., 2000).
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Drinking-water disinfection by-products Maternal exposure, limited evidence: Reviewers
found limited evidence for an association between NTDs and prenatal DBP exposure (Bove et al.,
2002; Graves et al., 2001). However, a meta-analysis of four epidemiologic studies published by
1999 (Magnus et al., 1999) indicated a modest association between NTDs and THM exposure indices (OR=1.49, 95% CI 1.08–2.05) (Hwang & Jaakkola, 2003). Among subsequently published
studies, NTDs were associated with drinking water BDCM but not chloroform levels in Nova Scotia
(BDCM ≥20 vs. <5 µg/L, OR=2.5, 95% CI 1.2–5.1; chloroform ≥100 vs. <50 µg/L, OR=1.2, 95%
CI 0.7–2.3) (Dodds & King, 2001). There was no association among recent studies in Sweden
(nationwide, maternal residence in hypochlorite-treated vs. nonchlorinated municipalities,
OR=1.4, 95% CI 0.7–1.4) (Kallen & Robert, 2000) and Norway (chlorinated high-color vs. unchlorinated low colour water, OR=0.7, 95% CI 0.2–2.0) (Hwang et al., 2002). A report of two
case-control studies in California noted inconsistent findings with no evidence of a dose-response
relationship in either study (Shaw et al., 2003b). Halogenated acetic acids cause NTDs in
experimental animals; the potential teratogenicity of many other DBPs has not been evaluated
(Nieuwenhuijsen et al., 2000).
Drinking-water nitrate Prenatal drinking water nitrate level, limited evidence: Reviewers noted
limited evidence for an association between birth defects (especially neural tube) and drinking water
nitrate in humans but no evidence that nitrate is teratogenic in experimental animals (Fan & Steinberg,
1996). A case-control study in Australia reported an association between CNS defects and maternal
drinking water source at birth (groundwater with nitrate >15 mg/L vs. surface water with nitrate
<1 mg/L, OR=3.5, 95% CI 1.1–14.6, adjusted for maternal age and parity but not other potential
confounders) (Dorsch et al., 1984). A Canadian population-based case-control study observed a
statistically nonsignificant elevated risk of NTDs among women with elevated residential well water
nitrate levels (≥26 vs. ≤0.1 mg/L, OR=2.30, 95% CI 0.73–7.29) (Arbuckle et al., 1988). A large
case-control study in California revealed an association between nitrate levels in drinking water
from ground sources serving the periconceptual maternal residence and anencephaly (≥36 vs. <5 mg/L,
OR=6.9, 95% CI 1.9–24.9) but not spina bifida (OR=1.1, 95% CI 0.25–4.5) (Croen et al., 2001).
However, this study found no association between NTDs and maternal periconceptual dietary
intake of nitrate, nitrite, or N-nitroso compounds, suggesting that nitrate might be a marker for
other teratogenic agents in groundwater. A small case-control study in Texas reported an elevated
risk of NTDs among women with elevated drinking-water nitrate levels at the maternal periconceptual residence (≥3.5 vs. <3.5 mg/L, OR=1.9, 95% CI 0.8–4.6) (Brender et al., 2004).
Hazardous waste disposal sites Maternal exposure, limited evidence: Reviewers found limited
evidence for a weak association between birth defects, especially NTDs and cardiac defects, and
maternal prenatal residential proximity to hazardous waste disposal sites (Dolk & Vrijheid, 2003).
They noted several important limitations of epidemiologic studies of birth defects and environmental
contaminants, including the heterogeneity and relative rarity of birth defects, crude exposure
indices, and inadequate assessment of potential confounders. For instance, in a population-based
case-control study in New York State (excluding New York City), CNS defects were associated with
prenatal residential proximity to hazardous waste disposal sites (<1.6 vs. ≥1.6 km, OR=1.29, 95%
CI 1.05–1.59) but not with an exposure index based on chemical toxicity and likelihood of exposure (high vs. no exposure, OR=1.48, 95% CI 0.69–3.16) (Geschwind et al., 1992). Subsequently,
a similar study in New York State reported no association between prenatal residential proximity to
NPL sites (<1.6 vs. ≥1.6 km, OR=0.92, 95% CI 0.79–1.08, adjusted for proximity to Toxic Release
Inventory [TRI] sites) (Marshall et al., 1997) but did find a borderline association between CNS
defects and prenatal residential proximity to TRI sites emitting solvents (≤0.5 vs. >0.5 km,
OR=1.39, 95% CI 0.97–2.01). In a population-based case-control study in California, NTDs were
weakly associated with periconceptual maternal residential proximity to NPL sites (<1.6 vs. ≥1.6 km,
OR=1.4, 95% CI 0.8–2.4); there were stronger associations with the subsets of NPL sites
containing pesticides (OR=2.2, 95% CI 0.9–5.2), lead (OR=2.0, 95% CI 0.9–4.1), pyrene
(OR=3.1, 95% CI 1.0–8.6), PCBs (OR=3.5, 95% CI 0.9–10.6), or benzene (OR=1.9, 95% CI 0.9–3.6)
(Croen et al., 1997). A European case-control study revealed a nonmonotonic dose-response
relationship between NTDs and hazard categories based on expert-rated potential for toxicant
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exposure via air or water (high vs. low hazard, OR=1.89, 95% CI 0.84–4.29, p-trend=0.64)
(Vrijheid et al., 2002b). In a large case-control study of minority-group women in California, there
was no association between CNS defects and prenatal residence in a census tract with at least one
NPL site containing solvents, pesticides or a category comprising PCBs, dioxins and polycyclic
aromatic hydrocarbon (PAHs) (Orr et al., 2002). A United Kingdom-wide retrospective cohort study
revealed no association between NTDs and prenatal residential proximity to any of 774 hazardous
waste sites (≤2 vs. >2 km, OR=1.07, 95% CI 0.95–1.20) (Elliott et al., 2001). Three studies were
not included in the review by Dolk and Vrijheid (2003). A retrospective cohort study in Scotland
found no association between NTDs and prenatal residential proximity to hazardous waste sites (≤2 vs.
>2 km from any of 61 sites, OR=0.71, 95% CI 0.36–1.42) (Morris et al., 2003). In an ecologic
study, NTD risk was elevated among women living in Sydney, Nova Scotia, site of a major hazardous
waste disposal site (SIR=1.83, 95% CI 1.08–3.09) but not in the remainder of Cape Breton County
(Dodds & Seviour, 2001). In a small Alaskan retrospective cohort study, there was a statistically
nonsignificant elevated risk of CNS defects among women in Native villages with hazardous waste
sites (higher vs. lower hazard dumpsite, OR=2.36, 95% CI 0.37–14.7) (Gilbreath & Kass, 2006a).
Incinerators Maternal exposure, inadequate evidence: A UK retrospective cohort study
reported a weak association between NTDs and prenatal residential proximity to incinerators (per
unit change in an inverse distance function, OR=1.13, 95% CI 1.04–1.23) but not crematoria
(Dummer et al., 2003a, 2003b). In a large French retrospective cohort study, NTDs were not
associated with prenatal residence in communities with solid waste incinerators or expert-rated
potential for exposure to dioxin, metal or dust emissions (high vs. low hazard, OR=0.83, 95% CI
0.35–1.96) (Cordier et al., 2004).
Solvents Recent meta-analyses reported associations between the broad category of major birth
defects and periconceptual parental occupational solvent exposure (maternal, summary OR=1.64,
95% CI 1.16–2.30; paternal, summary OR=1.47, 95% CI = 1.18–1.83) (Logman et al., 2005).
Maternal exposure, chlorinated solvents, inadequate evidence: A retrospective cohort study in
New Jersey revealed increased NTD risks for women with 1st trimester residence in communities
served by drinking water with trichloroethylene levels above 10 µg/L (compared to ≤1 µg/L,
OR=2.53, 90% CI 0.91–6.37) or carbon tetrachloride levels above 1 µg/L (compared to ≤1 µg/L,
OR=5.39, 90% CI 1.12–19.0) (Bove et al., 1995).
Maternal occupational exposure, glycol ethers, limited evidence: NTDs were associated with
maternal 1st trimester occupational exposure to glycol ethers in Europe (OR=1.94, 95% CI 1.16–3.24)
(Cordier et al., 1997) and Texas (OR=∞, 95% CI 1.8–∞, 7 exposed cases and 0 exposed control
mothers) (Brender et al., 2002). Reviewers concluded that NTDs were associated with maternal 1st
trimester occupational exposure to glycol ethers (Shi & Chia, 2001).
Maternal exposure, unspecified solvents, inadequate evidence: In a case-control study in the
Netherlands, spina bifida was not associated with 1st trimester maternal occupational solvent
exposure (alcohol, yes vs. no, OR=0.9, 95% CI 0.5–1.7; other organic solvents, OR=1.5, 95% CI
0.6–4.0) (Blatter et al., 1996b). Combined analysis of 2 Californian case-control studies revealed no
association between NTDs and periconceptual maternal residence within 1 mile of NPL sites containing
volatile organic carbons (VOCs) (OR=1.4, 95% CI 0.8–2.3) (Croen et al., 1997). A hospital-based
case-control study in France found no association between CNS defects and prenatal occupational solvent exposure (OR=1.4, 90% CI 0.6–3.2) (Cordier et al., 1992). In a population-based case-control
study, CNS defects were associated with prenatal residential proximity to hazardous waste disposal
sites containing solvents (<1.6 vs. ≥1.6 km, OR=1.24, 95% CI 1.01–1.54) (Geschwind et al.,
1992). In a similar New York State study, there was a borderline association between CNS defects
and prenatal residential proximity to TRI sites emitting solvents (≤0.5 vs. >0.5 km, OR=1.39, 95%
CI 0.97–2.01) (Marshall et al., 1997). A case-control study in Texas reported an association
between NTDs and periconceptual maternal use of solvents at home or work (OR=2.5, 95%
CI 1.3–4.7) (Brender et al., 2002).
Paternal occupational exposure, unspecified solvents, limited evidence: A recent meta-analysis
of 5 epidemiologic studies published during 1976–2000 reported an association between NTDs
and paternal occupational solvent exposure (summary RR=1.86, 95% CI=1.40–2.46) (Logman
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et al., 2005). Among studies not included in the meta-analysis, CNS defects were not associated
with paternal occupation as printers in Norway (compared to other occupations, SIR=1.1, 95%
CI 0.53–2.0) (Kristensen et al., 1993) or self-reported periconceptual paternal occupational solvent
use in Texas (OR=0.8, 95% CI 0.5–1.4) (Brender et al., 2002).
Summary
Epidemiologic evidence for the role of environmental toxicants in neural tube birth defects
includes: limited evidence—prenatal exposure to DBPs, nitrate (drinking water), hazardous waste
disposal sites (residential proximity), glycol ethers (occupational); paternal occupational exposure to
TCDD, unspecified solvents.
Cardiac Birth Defects
Lead Maternal exposure, inadequate evidence: A case-control study in Massachusetts found
no association between cardiac birth defects and drinking-water lead levels in communities of
prenatal residence (>1 vs. ≤1 µg/L, OR=1.13, 95% CI 0.60–2.14); 90% of lead concentrations
were below 1 µg/L (Zierler et al., 1988). Further investigation with about fourfold more cases
revealed a borderline association between cardiac birth defects and lead levels in the water supply
serving the maternal residence during the 1st trimester (>1 vs. ≤1 µg/L, OR=2.2, 95% CI 0.9–5.7)
(Aschengrau et al., 1993). In a lead-polluted region of Italy, an ecologic study reported an elevated
risk of cardiac birth defects during 1982–1986 when emission levels were higher (SIR=2.59, 95%
CI 1.68–3.82) but not during 1991–1995 (SIR=0.97, 95% CI 0.57–1.54); annual air lead emissions
in the region decreased by 93% between 1979 and 1997 (Vinceti et al., 2001). In the Baltimore–
Washington Infant Study, there was a statistically nonsignificant elevated risk of pulmonary vein
defects among infants of women who were likely exposed to lead at home or work during the 3-mo
periods before and after conception (OR=1.57, 95% CI 0.64–3.47) (Jackson et al., 2004).
Paternal exposure, inadequate evidence: Case-control studies within the Baltimore–Washington
Infant Study found associations between pulmonary atresia and self-reported paternal frequent
exposure to lead soldering (OR=4.1, 95% CI 1.2–11) (Correa-Villasenor et al., 1993) and between
total anomalous pulmonary vein return and likely paternal occupational lead exposure (based on
industrial hygiene assessment, a job-exposure matrix or self-reports) during the 6 mo before
conception (OR=1.83, 95% CI 1.00–3.42) (Jackson et al., 2004).
Arsenic, cadmium, mercury, mixed metals Maternal exposure, inadequate evidence: In a
Massachusetts case-control study, infants of women living in communities with detectable drinkingwater inorganic mercury or arsenic, respectively, had increased risks of coarctation of the aorta
(arsenic >0.8 vs. ≤0.8 µg/L, OR=3.4, 95% CI 1.3–8.9) and patent ductus arteriosus (mercury >0.2 vs.
≤0.2 µg/L, OR=1.6, 95% CI 0.95–2.6) (Zierler et al., 1988). A hospital-based case-control study
reported no association between cardiovascular birth defects and maternal residence in communities with detectable drinking-water arsenic (>0.8 vs≤0.8 µg/L, OR=1.3, p > .05) or cadmium
(>0.4 vs≤0.4 µg/L, OR=1.6, p > .05) (Aschengrau et al., 1993). In a Swedish retrospective cohort
study, there was an elevated risk of confirmed cardiac birth defects (1.69, 95% CI 0.55–4.44)
among infants of women living in parishes near a copper smelter known to produce air emissions of
lead, arsenic, mercury, and cadmium (Wulff et al., 1996).
Paternal occupational exposure, inadequate evidence: A Swedish retrospective cohort study
reported no association between cardiac birth defects and prenatal maternal residential proximity
to a copper smelter (RR=0.51, 95% CI 0.26–1.00) or paternal employment as a blue collar worker
at the smelter (OR=0.55, 95% CI 0.30–1.01) (Wulff et al., 1996). A small Texan case-control study
found no relationship between NTDs and paternal occupations likely exposed to arsenic (OR=1.5,
95% CI 0.7–3.0) (Brender et al., 2006).
TCDD Maternal exposure, inadequate evidence: A retrospective cohort study of births in
French communities with solid waste incinerators and unexposed comparison communities
reported no association between cardiac birth defects and expert-rated hazard of incinerator emissions (high vs. low exposure, conotruncal defects, OR=0.97, 95% CI 0.58–1.60; other cardiac
defects, OR=1.05, 95% CI 0.72–1.53) (Cordier et al., 2004).
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Paternal occupational exposure, inadequate evidence: In a retrospective cohort study of
Vietnam veterans and partners, there was an elevated risk of cardiovascular birth defects among
offspring of men with low (OR=2.39, 95% CI 1.02–5.24) but not high serum TCDD levels
(OR=0.95, 95% CI 0.28–2.65) (TCDD categories defined earlier) (Wolfe et al., 1995). Among
offspring of male sawmill workers, conotruncal and septal defects were not associated with preconceptual chlorophenate exposure intensity (per 100 h exposure during peak exposure year up to 3 mo
before conception, OR=0.95, p > .05) or with cumulative exposure during the 3 mo before conception (per 100 hr exposure, OR=0.94, p > .05) (Dimich-Ward et al., 1996).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: A retrospective cohort study found no association between cardiac birth defects and prenatal residence in
high-wheat counties, a proxy for agricultural use of chlorophenoxy herbicides (high- vs. low-wheat
counties, OR=1.23, 95% CI 0.70–2.17) (Schreinemachers, 2003).
Maternal exposure, unspecified herbicides, inadequate evidence: The Baltimore–Washington
case-control study noted an association between transposition of great arteries and self-reported
periconceptual maternal herbicide exposure at home or work (OR=3.6, 95% CI 1.6–8.2) (Loffredo
et al., 2001). Further studies are needed to assess this finding and to explore specific herbicides and
dose-response relationships.
Maternal exposure, insecticides, inadequate evidence: In a California case-control study,
conotruncal defects were not associated with prenatal pet flea treatments (OR=1.2, 95% CI 0.8–1.8)
or professional indoor pesticide application (OR=1.2, 95% CI 0.7–2.0) (Shaw et al., 1999). There
were statistically nonsignificant elevated risks in case-control studies in Baltimore–Washington
(transposition of great arteries, self-reported maternal periconceptual insecticide exposure at home
or work, OR = 1.5, 95% CI 0.9–2.6) (Loffredo et al., 2001) and California (conotruncal defects,
self-reported prenatal occupational insecticide exposure, OR = 2.1, 95% CI 0.8–5.1) (Shaw
et al., 2003a).
Maternal exposure, unspecified pesticides, inadequate evidence: Conotruncal defects were
associated with prenatal residential proximity to crops (OR=1.4, 95% CI 0.9–2.0) or garden
pesticide application (OR=3.1, 95% CI 1.3–7.3) in California (Shaw et al., 1999). The Baltimore–
Washington case-control study reported an association between transposition of great arteries and
self-reported maternal periconceptual exposure at home or work to rodenticides (OR=5.1, 95%
CI 1.7–14.9) (Loffredo et al., 2001). The Baltimore–Washington case-control study reported a
statistically nonsignificant elevated risk of total anomalous pulmonary vein return defects related to
self-reported periconceptual maternal pesticide exposure at home or work (OR=2.06, 95% CI
0.82–5.15) (Correa-Villasenor et al., 1991). A retrospective cohort study revealed a statistically nonsignificant elevated congenital heart defect risk (as evidenced by heart murmurs) related to prenatal
pesticide exposure in Colombian greenhouses (crude OR=2.16, 95% CI 0.69–6.54, calculated
from data in paper) (Restrepo et al., 1990a). A subsequent report of the California study revealed an
association between conotruncal defects and prenatal occupational exposure to insecticides (as
noted earlier) but not to pesticides other than insecticides (OR=1.0, 95% CI 0.5–2.2) (Shaw et al.,
2003a). In another California case-control study, limited to minority groups, cardiovascular birth
defects were not associated with prenatal residence in a census tract with a NPL site containing
pesticides (OR=0.83, 95% CI 0.56–1.25) (Orr et al., 2002). The heterogeneity of these studies with
regard to type of cardiac defect evaluated and nonspecific exposure indices precludes strong
inferences.
Paternal occupational exposure, 2,4,5-T or chlorophenate wood preservatives, inadequate
evidence: A case-control study of cardiovascular birth defects in Atlanta revealed no overall association with exposure to Agent Orange (based on self-reports and military records) (OR=0.97, CI not
reported) but noted elevated risks of transposition of the great vessels (OR=1.49) and coarctation
of the aorta (OR=1.89) (Erickson et al., 1984). See also paternal occupational TCDD exposure
discussed earlier.
Paternal occupational exposure, unspecified pesticides, inadequate evidence: There was an
increased risk of cardiorespiratory birth defects among offspring of male licensed pesticide applicators in Minnesota (RR=1.69, 95% CI 1.04–2.76); this study did not present data separately for
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cardiac defects (Garry et al., 1996). Among other residents of agricultural counties (after excluding
licensed applicators), there were increased cardiorespiratory birth defect risks in areas with corn/
soybeans (RR=1.43, 95% CI 1.17–1.76) or wheat, sugar, and/or potatoes (RR=1.90, 95% CI 1.37–2.63).
In the Norwegian farm cohort study, cardiac birth defect risk was not elevated among offspring of
farmers (compared to nonfarm families, RR=0.83, 95% CI 0.68–1.02); 57% of fathers and 34% of
mothers worked at least 500 h/yr on their farms (Kristensen et al., 1997b). A California case-control
study reported no association between cardiac birth defects and periconceptual paternal occupational pesticide exposure (OR=1.0, 95% CI 0.6–1.9) (Shaw et al., 1999).
Outdoor air pollution Maternal exposure, limited evidence: A large case-control study in
southern California reported dose-response relationships between certain cardiac defects and
ambient air pollutant levels near the prenatal residence during the 2nd gestational month including
ozone (per 1 pphm ozone, aortic defects, OR=1.56, 95% CI 1.16–2.09; pulmonary valve defects,
OR=1.34, 95% CI 0.96–1.87) and CO (per 1 ppm CO, VSD, OR=1.33, 95% CI 1.00–1.78) (Ritz
et al., 2002). These associations were independent of socioeconomic status (SES) and other air
pollutants but this record-based study had no information on potential confounders such as prenatal
smoking. A retrospective cohort study in France observed an association between conotruncal
defects and traffic density near the prenatal residence (>50,000 vs. <10,000 vehicles/d, OR=1.88,
95% CI 1.07–3.30, p-trend=.02) (Cordier et al., 2004). A population-based case-control study in
Texas reported associations between ambient air pollutant levels near the maternal residence during
GW 3–8 and cardiac defects including ASD (4th vs. 1st quartile PM10, OR=2.27, 95% CI 1.43–3.60,
p-trend=.0001), VSD (4th vs. 1st quartile SO2, OR=2.16, 95% CI 1.51–3.09, p-trend < .0001)
and Tetralogy of Fallot (4th vs. 1st quartile CO, OR=2.04, 95% CI 1.26–3.29, p-trend=.002)
(Gilboa et al., 2005).
Drinking water disinfection by-products Maternal exposure, drinking water, limited evidence:
HAAs and haloacetonitriles cause cardiac defects in experimental animals (Graves et al., 2001;
Nieuwenhuijsen et al., 2000). However, the five epidemiologic studies of cardiac birth defects and
DBPs available to reviewers yielded inadequate evidence of an association (Bove et al., 2002).
A meta-analysis of three studies published by 1999 (Bove et al., 1995; Dodds et al., 1999; Magnus
et al., 1999) showed no association between major cardiac defects and DBP exposure indices (summary OR=0.95, 0.77–1.17) (Hwang & Jaakkola, 2003). In subsequently reported studies, cardiac
defects were associated with DBP indices in Sweden (total THM >10 vs. ≤10 µg/L, OR=1.30, 95%
CI 1.08–1.56) (Cedergren et al., 2002) and Norway (ventricular septal defect, chlorinated
high-color vs. unchlorinated low-color water, OR=1.81, 95% CI 1.05–3.09) but not in California
(community drinking-water THM levels 50–74 vs. <1.0 µg/L, OR=1.5, 95% CI 0.7–3.5) (Shaw
et al., 2003b).
Drinking water nitrate Maternal exposure, drinking water, inadequate evidence: In a casecontrol study of cardiac defects in Massachusetts, there was no association with drinking water
nitrate in the community of maternal residence at birth (>0.1 vs. ≤0.1 mg/L, OR=1.08, 95% CI
0.72–1.62) (Zierler et al., 1988). A Swedish cohort study reported an association between major
cardiac birth defects and water nitrate levels in the municipality of maternal residence at birth (≥4.0 vs.
<1.0 mg/L, crude OR=1.61, 95% CI 0.95–2.72, CI calculated from data in paper) (Cedergren
et al., 2002).
Hazardous waste disposal sites Maternal exposure, limited evidence: In a population-based
case-control study in California, there was a statistically nonsignificant elevated risk of conotruncal
defects among women with periconceptual residential proximity to NPL sites (<1.6 vs. ≥1.6 km,
OR=1.8, 95% CI 0.8–4.2); there were stronger associations with the subsets of NPL sites containing
lead (OR=2.3, 95% CI 0.8–6.4), arsenic (OR=2.3, 95% CI 0.8–6.4) or 1,1-dichloroethylene
(OR=2.0, 95% CI 0.8–5.2) (Croen et al., 1997). A European case-control study reported associations between maternal residential proximity to hazardous waste landfill sites and cardiac septal
(<3 vs. 3–7 km, OR=1.49, 95% CI 1.09–2.04) and artery/vein defects (OR=1.81, 95% CI 1.02–3.20)
(Dolk et al., 1998). Further analysis of the European study revealed a nonmonotonic dose-response
relationship between hazard categories based on expert-rated potential for toxicant exposure
via air or water and both cardiac defects (high vs. low hazard, OR=1.65, 95% CI 0.86–3.16,
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p-trend=0.31) and artery/vein defects (high vs. low hazard, OR=2.19, 95% CI 0.93–5.17,
p-trend=0.31) (Vrijheid et al., 2002b).
A United Kingdom-wide retrospective cohort study revealed a weak association between
cardiovascular birth defects and prenatal residential proximity to any of 774 hazardous waste sites
(≤2 vs. >2 km, OR=1.11, 95% CI 1.03–1.21) (Elliott et al., 2001). In a similar study limited to
Scotland, there was no association (≤2 vs. >2 km, OR=1.03, 95% CI 0.85–1.26) (Morris et al.,
2003). In a large case-control study in California, there was no association between cardiovascular
defects and prenatal residence in a census tract with at least one NPL site containing solvents,
pesticides or a category comprising PCBs, dioxins, and PAHs (Orr et al., 2002). A large populationbased case-control study in Dallas County revealed weak associations between maternal residential
proximity to any of 276 hazardous waste sites and cardiac birth defects (<1.6 vs. ≥1.6 km,
OR=1.2, 95% CI 1.1–1.4); there was a somewhat stronger association for the subgroup of
endocardial cushion defects (OR=1.8, 95% CI 1.0–3.1) (Malik et al., 2004). There was a statistically nonsignificant elevated risk of cardiovascular birth defects related to maternal residential
proximity to hazardous waste disposal sites in an ecologic study in Sydney, Nova Scotia (SIR=1.27,
95% CI 0.93–1.75) (Dodds & Seviour, 2001).
Incinerators Maternal exposure, inadequate evidence: A UK retrospective cohort study
reported a weak association between cardiac defects and prenatal residential proximity to incinerators (per unit change in inverse distance function, OR=1.12, 95% CI 1.03–1.22) but not crematoria
or hazardous industries (Dummer et al., 2003a, 2003b). In a large French retrospective cohort
study, conotruncal defects were not associated with prenatal residence in a communities with solid
waste incinerators or expert-rated potential for exposure to dioxin, metal or dust emissions (high vs.
low hazard, OR=0.97, 95% CI 0.58–1.60) (Cordier et al., 2004).
Solvents Prenatal occupational exposure, glycol ethers, inadequate evidence: In a multicentre
European case-control study, cardiac defects were associated with maternal 1st trimester occupational exposure to glycol ethers (OR=1.45, 95% CI 0.99–2.13) (Cordier et al., 1997).
Prenatal exposure, unspecified solvents, limited evidence: The Baltimore–Washington Infant
Study found an imprecise and statistically nonsignificant increased risk of total anomalous pulmonary venous return defects among women with self-reported periconceptual exposure to paint,
paint strippers, solvents, or degreasing agents (OR=3.4, 95% CI 0.7–16, 3 exposed case mothers)
(Correa-Villasenor et al., 1991). In a country-wide Finnish case-control study, 1st trimester maternal
occupational organic solvent exposure was associated with VSD (OR=1.5, 95% CI 1.0–3.7) but not
with total cardiovascular defects (OR=1.3, 95% CI 0.8–2.2) (Tikkanen & Heinonen, 1991) or conal
defects (OR=0.6, 95% CI 0.2–1.4) (Tikkanen & Heinonen, 1992b). Further analyses revealed associations between 1st trimester maternal occupational exposure to dyes, lacquers or paint and VSD
(crude OR=5.4, 95% CI 1.7–17, calculated from data in paper) and conus arteriosus syndrome
(crude OR=6.7, 95% CI 1.8–24) (Tikkanen & Heinonen, 1992a). A hospital-based case-control
study in France found no association between cardiac birth defects and prenatal occupational
solvent exposure (OR=1.3, 90% CI 0.3–6.2) (Cordier et al., 1992). Combined analysis of 2 Californian
case-control studies revealed no association between conotruncal defects and periconceptual
maternal residence within 1 mile of a NPL site containing VOCs (OR=1.3, 95% CI 0.6–3.2) (Croen
et al., 1997). The Baltimore–Washington Infant Study reported an elevated risk of isolated coarctation of the aorta among women with the highest frequency of 1st trimester occupational solvent
exposure (daily vs. never, OR=3.2, 95% CI 1.3–7.9) (Wollins et al., 2001). A large case-control
study in California reported a borderline association between conotruncal defects and prenatal
occupations likely exposed to aliphatic hydrocarbons (OR=1.6, 95% CI 0.8–3.3); the association
was stronger among infants with the GST M1 polymorphism (GSTM1) (OR=4.6, 95% CI 1.0–19)
(Shaw et al., 2003a).
Paternal occupational exposure, unspecified solvents, inadequate evidence: A Norwegian
cohort study of offspring of male printers revealed no increased risk of cardiac defects (compared to
other occupations, SIR=0.7, 95% CI 0.33–1.2) (Kristensen et al., 1993). In the Baltimore–Washington
Infant Study, there were dose-response relationships between intensity of paternal occupational
exposure to paint strippers (potentially containing trichloroethylene) and coarctation of the aorta
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(≥2–3 times/wk vs. unexposed, OR=4.3, 95% CI 1.0–14.4; p-trend=.012) and VSD (≥2–3 times/wk
vs. unexposed, OR=7.0, 95% CI 1.6–24.6; p-trend=.007) (Correa-Villasenor et al., 1993).
Summary There is limited epidemiologic evidence for associations between cardiac birth
defects and prenatal exposures: outdoor air pollution, DBPs in drinking water, residential proximity
to hazardous waste disposal sites, unspecified solvents (occupational).
Orofacial Clefts
Lead Maternal exposure, inadequate evidence: A case-control study in Massachusetts found a
statistically nonsignificant elevated risk of facial/neck birth defects among infants of women living in
communities with drinking-water lead levels above 1 µg/L (OR=1.7, p > .05) (Aschengrau et al.,
1993). A European case-control study reported an association between isolated cleft palate and
self-reported maternal 1st trimester occupational lead exposure (OR=3.0, 95% CI 1.1–8.6)
(Lorente et al., 2000). An ecologic study of a region of Italy polluted by air emissions from ceramic
tile plants showed that the risk of oral cleft defects was elevated during 1982–1986 (SIR=2.28,
95% CI 1.16–4.07) but decreased in recent years (1991–1995) after lead emissions declined
substantially (SIR=1.31, 95% CI 0.42–3.16) (Vinceti et al., 2001).
Paternal occupational exposure, inadequate evidence: Retrospective Norwegian cohort studies
found borderline associations between paternal occupational lead exposure inferred from job titles
and cleft lip (OR=1.6, 95% CI 1.0–2.5) (Kristensen et al., 1993) and cleft palate (OR=1.3, 95% CI
0.9–2.0) (Irgens et al., 1998). These studies did not measure blood lead levels and the men were
likely exposed to other toxicants (e.g., solvent exposure among printers).
TCDD Maternal exposure, inadequate evidence: A retrospective cohort study of births in
French communities with solid waste incinerators and unexposed comparison communities
reported no association between facial clefts and expert-rated hazard of incinerator emissions (high
vs. low exposure, OR=1.01, 95% CI 0.64–1.59) (Cordier et al., 2004).
Paternal exposure, inadequate evidence: In a retrospective cohort study of Vietnam veterans
and partners, there was no association between ear, face or neck defects and paternal serum TCDD
categories (low or high serum TCDD vs. unexposed or serum TCDD <10 pg/g lipid, OR=1.09,
95% CI 0.42–2.62 (calculated from data in paper)) (Wolfe et al., 1995).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: A retrospective cohort study found no association between cleft lip and/or palate and prenatal residence in
high-wheat counties, a proxy for chlorophenoxy herbicide exposure (high- vs. low-wheat counties,
OR=1.12, 95% CI 0.62–2.01) (Schreinemachers, 2003).
Maternal exposure, organophosphate insecticides, inadequate evidence: A small nested casecontrol study in San Francisco reported a statistically nonsignificant association between orofacial
defects and prenatal residence ≤1 km from areas sprayed with malathion (OR = 3.35, 95% CI
0.61–18.5) (Thomas et al., 1992).
Maternal exposure, unspecified insecticides, inadequate evidence: In a California case-control
study, there was no association between isolated cleft lip with or without cleft palate and 1st trimester
professional indoor insecticide application (OR=0.9, 95% CI 0.6–1.3) or pet flea treatments
(OR=0.8, 95% CI 0.6–1.2) (Shaw et al., 1999). Further analysis of this study revealed no association with maternal periconceptual occupational insecticide exposure (OR=0.9, 95% CI 0.4–2.2)
(Shaw et al., 2003a).
Maternal exposure, unspecified pesticides, inadequate evidence: In a large Finnish case-control
study, orofacial clefts were associated with maternal 1st trimester employment in agriculture (compared to other work, OR=1.9 95% CI 1.1–3.5) (Nurminen et al., 1995). In the California study,
there was no association between isolated cleft lip with or without cleft palate and 1st trimester
maternal occupational pesticide exposure (OR=1.1, 95% CI 0.6–2.0), maternal application of
garden pesticides (OR=1.2, 95% CI 0.5–2.8) or residential proximity to crops treated with pesticides (OR=0.9, 95% CI 0.6–1.4) (Shaw et al., 1999). Further analysis of this study revealed no association with maternal periconceptual occupational exposure to pesticides other than insecticides
(OR=1.0, 95% CI 0.6–1.7) (Shaw et al., 2003a). In another California case-control study, limited to
minority groups, oral clefts were not associated with prenatal residence in a census tract with an
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NPL site containing pesticides (OR=0.89, 95% CI 0.45–1.74) (Orr et al., 2002). The inconsistent
findings and the heterogeneous and nonspecific exposure indices preclude strong inferences.
Paternal exposure, 2,4,5-T, inadequate evidence: A case-control study in Atlanta revealed no
association between exposure to Agent Orange (based on self-reports and military records) and cleft
lip without cleft palate (OR=1.07, CI not reported) or with cleft palate (OR=0.76, CI not reported)
(Erickson et al., 1984). See also paternal occupational TCDD exposure above.
Paternal occupational exposure, unspecified pesticides, inadequate evidence: A cohort study of
Norwegian farm families found no increased risks of cleft lip with or without cleft palate (compared
to nonfarm families, OR=0.94, 95% CI 0.74–1.20) (Kristensen et al., 1997b). Cleft palate was not
associated with periconceptual paternal occupational pesticide exposure in a large California casecontrol study (OR=1.1, 95% CI 0.8–2.1) (Shaw et al., 1999).
Tobacco smoke Maternal active smoking, limited evidence: The U.S. Surgeon General concluded that there is suggestive evidence of a causal relationship between prenatal active smoking
and oral clefts (U.S. Department of Health and Human Services, 2004).
Maternal ETS exposure, inadequate evidence: An expert panel concluded that there is inadequate evidence for an association between oral cleft of other specific birth defects and prenatal ETS
exposure (California Environmental Protection Agency, 2005).
Outdoor air pollution Prenatal environmental exposure, inadequate evidence: A large casecontrol study in southern California reported a weak association between cleft lip/palate and ambient
air ozone levels near the prenatal residence during the 2nd gestational month (per 1 pphm ozone,
OR=1.13, 95% CI 0.90–1.40) (Ritz et al., 2002). A retrospective cohort study in France observed
no association between facial cleft defects and traffic density near the prenatal residence (>50,000
vs. <10,000 vehicles/d, OR=1.07, 95% CI 0.67–1.72) (Cordier et al., 2004). In a population-based
case-control study in Texas, cleft lip and/or palate were marginally associated with ambient air PM10
levels near the maternal residence during GW 3–8 (4th vs. 1st quartile, OR=1.37, 95% CI 0.94–2.00,
p-trend=.09) but not with SO2, CO, ozone, or NO2 levels (Gilboa et al., 2005).
Paternal occupational exposure, inadequate evidence: A large Norwegian retrospective cohort
study reported no association between cleft lip and/or cleft palate and paternal occupation as drivers
(OR=1.10, 95% CI 0.91–1.36) (Irgens et al., 2000).
Drinking-water disinfection by-products Prenatal drinking water DBP levels, inadequate
evidence: Reviewers noted limited evidence for an association between orofacial clefts and DBP
exposure indices (Bove et al., 2002), but a meta-analysis of three epidemiologic studies published
by 1999 (Bove et al., 1995; Dodds et al., 1999; Magnus et al., 1999) revealed no association
(summary OR=1.09, 95% CI 0.79–1.51) (Hwang & Jaakkola, 2003). There was also no association
in subsequently published studies in Sweden (hypochlorite-treated vs. nonchlorinated drinking
water, OR=1.4, 95% CI 0.7–1.4) (Kallen & Robert, 2000), Nova Scotia (chloroform ≥100 vs. <50
µg/L, OR=1.5, 95% CI 0.8–2.8), (Dodds and King 2001) or Norway (chlorinated high color vs.
nonchlorinated low color, OR=0.90, 95% CI 0.52–1.58) (Hwang et al., 2002). A case-control
study in California reported a statistically nonsignificant association between isolated cleft lip/palate
and periconceptual drinking water THM levels with a nonmonotonic dose-response relationship
(THM 50–74 vs. <1 µg/L, OR=1.9, 95% CI 0.8–4.5) (Shaw et al., 2003b). Reviewers concluded
that halogenated acetic acids and chloroform cause craniofacial defects in experimental animals
(Graves et al., 2001; Nieuwenhuijsen et al., 2000).
Hazardous waste disposal sites Prenatal residential proximity, inadequate evidence: In a
population-based case-control study, oral cleft defects were not associated with prenatal residential
proximity to hazardous waste disposal sites (<1.6 vs. ≥1.6 km, OR=1.15, 95% CI 0.87–1.51);
there was a statistically nonsignificant elevated risk related to proximity to sites containing pesticides
(OR=1.27, 95% CI 0.84–1.92) (Geschwind et al., 1992).
In a population-based case-control study in California, there was no association between oral
cleft defects and maternal periconceptual residential proximity to NPL sites (<1.6 vs. ≥1.6 km,
OR=1.0, 95% CI 0.5–2.3) or the subsets of sites containing lead heavy metals or solvents (Croen
et al., 1997). A European case-control study reported a statistically nonsignificant elevated risk of
cleft palate among women with prenatal residential proximity to hazardous waste landfill sites
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(<3 vs. 3–7 km, OR=1.63, 95% CI 0.77–3.41) (Dolk et al., 1998). In a large case-control study in
California, there was no association between oral cleft defects and prenatal residence in a census
tract with at least one NPL site containing solvents, pesticides or a category comprising PCBs,
dioxins, and PAHs (Orr et al., 2002).
Incinerators Maternal exposure, inadequate evidence: In a large French retrospective cohort
study, facial cleft defects were not associated with prenatal residence in a communities with solid
waste incinerators or expert-rated potential for exposure to dioxin, metal or dust emissions (high vs.
low hazard, OR=1.01, 95% CI 0.64–1.59) (Cordier et al., 2004).
Solvents Maternal exposure, glycol ethers, limited evidence: In a multicenter European
case-control study, cleft lip/palate was associated with maternal 1st trimester occupational exposure
to glycol ethers (OR=1.97, 95% CI 1.20–3.25) (Cordier et al., 1997). Further analysis of the
European study, limited to subjects who worked during pregnancy, revealed associations between
cleft lip/palate and 1st trimester maternal occupational exposure to glycol ethers (OR=2.10, 95%
CI 1.14–3.88) and other solvents (see later discussion) (Lorente et al., 2000). A recent large casecontrol study in California found a borderline association between cleft palate plus other defects
and prenatal occupational exposure to glycol ethers and derivatives (OR=2.2, 95% CI 0.8–7.0)
(Shaw et al., 2003a).
Maternal exposure, chlorinated solvents, limited evidence: A retrospective cohort study in
New Jersey revealed associations between oral cleft defects and 1st trimester maternal residence in
communities served by drinking water with trichloroethylene levels above 10 µg/L (compared to
≤1 µg/L, OR=2.24, 90% CI 1.16–4.20) or carbon tetrachloride levels above 1 µg/L (compared to
≤1 µg/L, OR=3.60, 90% CI 0.75–12.5) (Bove et al., 1995). In a French case-control study, facial
defects were associated with self-reported 1st trimester occupational exposure to solvents (any vs.
none, OR=1.62, 95% CI 1.04–2.52) including halogenated solvents (aliphatic, OR=4.40, 95% CI
1.41–16.2; aromatic, OR=1.78, 95% CI 0.89–3.54) (Laumon et al., 1996). Among subjects who
worked during pregnancy, a European case-control study revealed a statistically nonsignificant
elevated risk of cleft lip/palate related to 1st trimester maternal occupational trichloroethylene
exposure (OR=3.21, 95% CI 0.49–20.9, based on 4 exposed case and 2 exposed control mothers)
(Lorente et al., 2000).
Maternal exposure, other and unspecified solvents, inadequate evidence: A hospital-based
case-control study in France observed increased risk of oral cleft defects related to prenatal occupational solvent exposure (OR=3.3, 90% CI 0.8–18.1) (Cordier et al., 1992). Combined analysis of 2
Californian case-control studies revealed no association between oral cleft defects and periconceptual maternal residence within 1 mile of a NPL site containing VOCs (OR=0.8, 95% CI 0.4–1.8)
(Croen et al., 1997). Among subjects who worked during pregnancy, a European case-control study
revealed associations between cleft lip/palate and 1st trimester maternal occupational exposure to
aliphatic alcohols (OR=1.67, 95% CI 0.86–3.26) (Lorente et al., 2000). A recent large case-control
study in California found a borderline association between isolated cleft palate and prenatal occupational exposure to aliphatic hydrocarbons (OR=2.2, 95% CI 0.9–5.7) (Shaw et al., 2003a).
Paternal exposure, unspecified solvents, inadequate evidence: A Norwegian cohort study of offspring of male printers revealed an increased risk of cleft lip/palate (compared to other occupations,
SIR=1.6, 95% CI 0.97–2.5) (Kristensen et al., 1993).
Summary There is limited epidemiologic evidence for associations between orofacial birth defects
and prenatal active smoking and occupational exposure to glycol ethers or chlorinated solvents.
Developmental Tooth Abnormalities
Natal teeth Maternal high-level PCB/PCDF exposure, sufficient evidence: Prenatal consumption of cooking oil contaminated with high levels of PCBs, PCDFs, and related toxicants in Taiwan
was associated with natal teeth (prevalence, exposed vs. unexposed, 11/127 vs. 0/113, OR=∞,
95% CI 3.0 to ∞) (Rogan et al., 1988). Further investigation revealed a dose-response relationship
between a history of natal teeth and maternal serum PCB levels (0.0% among unexposed children;
5.3, 11.5, and 13.0% among those with increasing maternal serum PCB tertiles, p-trend=.003)
(Wang et al., 2003a).
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Maternal background PCB/PCDD/PCDF exposure, inadequate evidence: In a small Finnish
cohort, natal teeth were not associated with breast milk TCDD-TEQ exposure from PCDDs/PCDFs
or PCBs (cases vs. noncases, mean PCDD/PCDF-TEQ 11.9 vs. 8.6 pg/g milk lipid, p=.11; mean
PCB-TEQ 7.2 vs. 5.3 pg/g milk lipid, p=.31) (Alaluusua et al., 2002).
Lactational PCB/PCDF exposure, inadequate evidence: Among Yucheng children, there was an
irregular relationship between a history of natal teeth and breastfeeding duration (0.0% among unexposed children; 14.3% among forumula-fed Yucheng children and 0 and 10%, respectively, among
Yucheng children breast-fed for shorter or longer periods, p-trend=.08) (Wang et al., 2003a).
Hypomineralized enamel and other developmental tooth defects Maternal high-level PCB/
PCDF exposure, limited evidence: Prenatal consumption of cooking oil contaminated with high
levels of PCBs, PCDFs, and related toxicants in Taiwan was associated with missing permanent
tooth germ (exposed vs. unexposed, 5/18 vs. 1/44; OR=16.5, 95% CI 1.6–411, calculated from
data in paper) (Lan et al., 1989). Further investigation revealed a dose-response relationship
between the prevalence at age 7–11 yr of other developmental tooth defects (fusion, microdontia,
pigmentation, enamel hypoplasia, impaction) and maternal serum PCB levels (2.7% among
unexposed children; 9.1, 11.5, and 24.0% among increasing maternal serum PCB tertiles,
p-trend=.001) (Wang et al., 2003a).
Maternal background PCB exposure, inadequate evidence: A cross-sectional study of Slovenian
children found a higher prevalence of enamel defects among residents of a PCB-contaminated
region compared with those from a relatively uncontaminated region (22 vs. 13%, p < .001) (Jan &
Vrbic, 2000). Unfortunately, this study did not assess the relationship between enamel defects and
biomarkers of PCB exposure.
Maternal PCDD/PCDF/PCB-TEQ or TCDD exposure, inadequate evidence: A small Finnish
cohort study (102 children) showed that hypomineralized enamel defects at age 6–7 yr were associated with total lactational TCDD-TEQ exposure from PCDDs/PCDFs and PCBs; the statistical significance was stronger for PCDD/PCDF-TEQ (p=.004) than for PCB-TEQ (p=.07); this short
communication did not include other statistical data on the strength of these associations or potential confounders (Alaluusua et al., 1999). A small retrospective cohort study of persons exposed
before age 10 yr to TCDD at Seveso revealed an association between developmental tooth enamel
defects (opacities or hypoplasia) and serum TCDD at baseline soon after the incident (≥238 vs.
<238 pg/g lipid, OR=2.4, 95% CI 1.3–4.5) (Alaluusua et al., 2004).
Summary There is sufficient epidemiologic evidence that high-level prenatal exposure to
PCBs, PCDFs, and related toxicants can produce developmental tooth abnormalities, including
hypomineralized enamel defects of permanent teeth.
Musculoskeletal Birth Defects
Cadmium Maternal exposure, inadequate evidence: A Massachusetts case-control study
reported no association between drinking water cadmium levels in the community of maternal
residence at delivery and risk of musculoskeletal birth defects (OR=0.9, p > .05) (Aschengrau
et al., 1993).
TCDD Paternal occupational exposure, inadequate evidence: In a retrospective cohort study
of Vietnam veterans and partners, there was no association between musculoskeletal birth defects
and paternal serum TCDD levels categorized as low (current level >10 and initial level ≤110 pg/g
lipid, OR=1.08, 95% CI 0.72–1.60) or high (current level >10 and initial level >110 pg/g lipid,
OR=0.89, 95% CI 0.58–1.32, calculated from data in paper) (Wolfe et al., 1995).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: A retrospective cohort study found an association between musculoskeletal birth defects and maternal
residence in high-wheat counties, a proxy for agricultural chlorophenoxy herbicide use (high- vs.
low-wheat counties, OR=1.50, 95% CI 1.06–2.12) (Schreinemachers, 2003). There was also an
association between such exposure and polydactyly/syndactyly (OR=2.43, 95% CI 1.26–4.71) but
not clubfoot (OR=0.84, 95% CI 0.39–1.80).
Maternal exposure, unspecified herbicides, inadequate evidence: A case-control study in
New York State reported no association between limb reduction defects and maternal residence in
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counties with high agricultural herbicide use (high vs. low acres treated (1987 census), OR=0.82,
95% CI 0.41–1.64 (Lin et al., 1994).
Maternal exposure, organophosphate insecticides, inadequate evidence: The San Francisco
study observed a statistically nonsignificant association between limb defects and 1st trimester
maternal residence in areas sprayed with malathion (OR=1.73, 95% CI 0.87–3.46) (Thomas et al.,
1992).
Maternal exposure, unspecified insecticides, inadequate evidence: A case-control study in
New York State reported no association between limb reduction defects and maternal residence in
counties with high agricultural insecticide use (high vs. low acres treated (1987 census), OR=0.68,
95% CI 0.34–1.36 (Lin et al., 1994). In a California case-control study, limb reduction defects were
associated with self-reported 1st trimester professional indoor insecticide application at the maternal residence (OR=1.6, 95% CI 1.0–2.7) but not with pet flea insecticide use (OR=0.9, 95% CI
0.6–1.4) (Shaw et al., 1999). Further analysis of this study revealed no association with periconceptual maternal occupational insecticide exposure (OR=0.7, 95% CI 0.2–3.4, only 2 exposed case
mothers) (Shaw et al., 2003a).
Maternal exposure, unspecified pesticides, inadequate evidence: Among women prenatally
exposed to pesticides in Colombian semi-enclosed greenhouses, there was an elevated risk of
congenital dislocated hip (crude OR = 2.90, 95% CI 1.19–6.99, calculated from data in paper)
(Restrepo et al., 1990a). A retrospective cohort study in Minnesota observed increased risks of
musculoskeletal defects in families (neither parent was a licensed pesticide applicator) living in
regions with corn/soybean (compare to noncrop regions, OR = 1.36, 95% CI 1.18–1.58) or
wheat/sugar beet/potato crops (OR = 1.75, 95% CI 1.37–2.22) (Garry et al., 1996). In a
California case-control study, limb reduction defects were associated with self-reported 1st
trimester professional (OR = 3.5, 95% CI 1.2–9.9) but not maternal application of garden pesticides (OR = 1.5, 95% CI 0.5–4.6, 5 exposed case mothers) (Shaw et al., 1999). Further analysis
of this study revealed no association with prenatal residential proximity to crops treated with
pesticides (OR = 1.0, 95% CI 0.6–1.8) or occupational exposure to insecticides or noninsecticide pesticides (OR = 0.7, 95% CI 0.3–1.9, only 5 exposed case mothers) (Shaw et al., 2003a).
In a cohort study of women employed in agriculture in Washington State, limb reduction defect
risk was elevated (compared to neither parent in agriculture, OR = 2.8, 95% CI 1.2–6.3) (Engel
et al., 2000).
In a California case-control study, limb reduction defects were not associated with parental
employment in agriculture (either parent vs. neither, OR=0.9, 95% CI 0.4–1.7) but were related to
intensity of restricted agricultural pesticide use in counties of maternal residence (high vs. minimal
use, OR=1.9, 95% CI 1.2–3.1, p-trend=0.02) (Schwartz & LoGerfo, 1988). A case-control study in
New York State reported no association between limb reduction defects and maternal occupations
likely exposed to pesticides (OR=0.7, 95% CI 0.4–1.5) or with maternal residence in counties with
high per capita acreage in farms (high vs. low acreage, OR=1.49, 95% CI 0.53–4.23, calculated
from data in paper) (Lin et al., 1994). A Finnish case-control study revealed no association between
skeletal defects and 1st trimester maternal employment in agriculture (OR=0.8, 95% CI 0.4–1.7,
only 3 exposed case mothers) (Nurminen et al., 1995). In New York State, musculoskeletal defects
were not associated with maternal residence within 1.6 km of a hazardous waste disposal site
known to contain pesticides (OR=0.80, 95% CI 0.51–1.26) or a TRI industrial site known to emit
pesticides into air (OR=1.12, 95% CI 0.93–1.35) (Marshall et al., 1997). Another California
case-control study, limited to racial or ethnic minority populations, revealed no association
between musculoskeletal defects and prenatal residence in a census tract with a NPL site containing
pesticides (OR=1.19, 95% CI 0.92–1.54) (Orr et al., 2002). The inconsistent findings and the
heterogeneous and nonspecific exposure indices and the variable scope and great heterogeneity of
musculoskeletal defects preclude strong inferences.
Paternal occupational exposure, 2,4,5-T, inadequate evidence: A case-control study of musculoskeletal birth defects in Atlanta revealed no association with exposure to Agent Orange (based on
self-reports and military records) (OR=1.06, CI not reported) (Erickson et al., 1984). See also paternal
occupational TCDD exposure discussed earlier.
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Paternal occupational exposure, unspecified pesticides, inadequate evidence: Among infants
of women age 30 or older, musculoskeletal defects were associated with paternal occupation as
licensed pesticide applicators in Minnesota (OR = 2.52, 95% CI 1.58–4.01); there was no association among offspring of younger mothers (OR = 0.94, 95% CI 0.52–1.71) (Garry et al., 1996).
A case-control study in New York State found no association between limb reduction defects and
paternal occupations likely exposed to pesticides (OR = 0.9, 95% CI 0.5–1.6) or the subgroup
employed as farmers (OR= 1.1, 95% CI 0.5–2.7) (Lin et al., 1994). Limb reduction defect risk
was not elevated among all infants in Norwegian farm families (OR = 0.84, 95% CI 0.51–1.37)
but was elevated on all farms reporting pesticide expenditures during agricultural censuses
(OR = 1.79, 95% CI 0.98–3.26) and the subgroup of grain farms with such expenditures
(OR = 2.50, 95% CI 1.06–5.90) (Kristensen et al., 1997b). A retrospective cohort study of the
general Norwegian population reported no association between limb reduction defects and
paternal occupation in agriculture, based on birth certificates (OR = 0.89, 95% CI 0.56–1.33)
(Irgens et al., 2000). A case-control study in California found no association between limb reduction defects and self-reported periconceptual paternal occupational pesticide exposure
(OR = 1.2, 95% CI 0.6–2.2) (Shaw et al., 1999).
Outdoor air pollution Paternal occupational exposure, inadequate evidence: A large Norwegian retrospective cohort study reported no association between club foot and paternal occupation
as drivers (OR=0.94, 95% CI 0.79–1.10) (Irgens et al., 2000).
Drinking-water disinfection by-products Prenatal drinking-water DBP level, inadequate evidence:
A large population-based case-control study in New York State found no association between
musculoskeletal birth defects and drinking water THM levels (≥100 vs. <100 µg/L, OR=0.76, 95%
CI 0.61–0.95) (Marshall et al., 1997).
Drinking-water nitrate/nitrite Prenatal drinking-water DBP level, inadequate evidence: A casecontrol study in Australia reported an association between musculoskeletal defects and maternal
drinking water source at birth (groundwater with nitrate generally >15 mg/L vs. surface water with
nitrate <1 mg/L, OR=2.9, 95% CI 1.2–8.0, adjusted for maternal age and parity but not other
potential confounders) (Dorsch et al., 1984).
Hazardous waste disposal sites Maternal exposure, inadequate evidence: In a populationbased case-control study in New York State, musculoskeletal defects were weakly associated with
prenatal residential proximity to any hazardous waste disposal site (<1.6 vs. ≥1.6 km, OR=1.16,
95% CI 1.06–1.26); the association was somewhat stronger for the subgroup of sites with a high
exposure index, based on chemical toxicity and likelihood of exposure (high vs. no exposure,
OR=1.75, 95% CI 1.31–2.34) and for sites containing pesticides (OR=1.20, 95% CI 1.05–1.38)
(Geschwind et al., 1992). Subsequently, a similar study in New York State reported no association
between musculoskeletal defects and prenatal residential proximity to NPL sites (<1.6 vs. ≥1.6 km,
OR=1.00, 95% CI 0.94–1.07, adjusted for proximity to TRI sites) (Marshall et al., 1997).
A European case-control study reported an association between gastroschisis and maternal residential proximity to hazardous waste landfill sites (<3 vs. 3–7 km, OR=3.19, 95% CI 0.95–10.8) (Dolk
et al., 1998). A United Kingdom-wide retrospective cohort study revealed no association between
abdominal wall defects and prenatal residential proximity to any of 774 hazardous waste sites (≤2 vs.
>2 km, OR=1.03, 95% CI 0.86–1.25) (Elliott et al., 2001). In a large case-control study in California,
there were statistically nonsignificant elevated risks of musculoskeletal defects among women living
in a census tract with at least 1 NPL site containing solvents, pesticides, or a category comprising
PCBs, dioxins and PAHs (e.g., pesticides, OR=1.19, 95% CI 0.92–1.54) (Orr et al., 2002). In a
small Alaskan retrospective cohort study, musculoskeletal defects were associated with maternal
residence in Native villages with hazardous waste sites (higher vs. lower hazard dumpsite,
OR=4.27, 95% CI 1.76–10.4) (Gilbreath & Kass, 2006a). Given the heterogeneity of musculoskeletal
birth defects and exposure indices, it is difficult to draw firm conclusions from the few available
epidemiologic studies.
Incinerators Maternal exposure, inadequate evidence: In a large French retrospective cohort
study, limb reduction defects were not associated with prenatal residence in a communities with
solid waste incinerators (OR=0.78, 95% CI 0.51–1.20) (Cordier et al., 2004).
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Solvents Prenatal occupational exposure, glycol ethers, inadequate evidence: In a multicentre
European case-control study, musculoskeletal birth defects were weakly related to maternal 1st trimester occupational exposure to glycol ethers (OR=1.32, 95% CI 0.85–2.05) (Cordier et al., 1997).
Maternal exposure, unspecified solvents, inadequate evidence: In a similar New York State
study, there was no association between musculoskeletal defects and prenatal residential proximity
to hazardous waste disposal sites containing solvents (<1.6 vs. ≥1.6 km, OR=1.03, 95% CI 0.81–1.31)
or TRI sites emitting solvents (<1.6 vs. ≥1.6 km, OR=1.02, 95% CI 0.93–1.12) (Marshall et al., 1997).
Paternal occupational exposure, unspecified solvents, inadequate evidence: A hospital-based
case-control study in France found no association between musculoskeletal defects and prenatal
occupational solvent exposure (OR=1.7, 90% CI 0.7–4.2) (Cordier et al., 1992). A Norwegian
cohort study of offspring of male printers revealed no increased risk of clubfoot (compared to other
occupations, SIR=0.9, 95% CI 0.70–1.1) or other limb reduction defects (Kristensen et al., 1993).
Summary There inadequate epidemiologic evidence for associations between musculoskeletal
birth defects and any of the environmental contaminants examined.
Urinary-Tract Birth Defects
TCDD Maternal exposure, inadequate evidence: A retrospective cohort study of births in
French communities with solid waste incinerators and unexposed comparison communities
revealed that expert-rated hazard of incinerator emissions was associated with obstructive urinary
tract defects (high vs. low exposure, OR=1.93, 95% CI 0.94–3.93, p-trend=0.07) but not with
renal dysplasia (OR=1.30, 95% CI 0.57–2.97) (Cordier et al., 2004).
Paternal occupational exposure, inadequate evidence: In a retrospective cohort study of Vietnam veterans and partners, urinary tract birth defects were associated with paternal serum TCDD
levels categorized as low (OR=1.97, 95% CI 0.70–4.97) or high (OR=2.12, 95% CI 0.81–5.12)
(calculated from data in paper)) (Wolfe et al., 1995). These findings preclude firm inferences as the
odds ratios were not statistically significant (there were only 17 cases among the 3 exposure categories) and there was no testing or adjustment for potential confounders. A review concluded that
early gestational exposure to relatively low doses of TCDD and dioxin-like chemicals can cause
ureteral hyperplasia and hydronephrosis in experimental animals (Birnbaum, 1995).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: A retrospective cohort study of Minnesota and 3 adjacent states found no association between genitourinary
defects and maternal residence in high-wheat regions, a proxy for chlorophenoxy herbicide exposure (high- vs. low-wheat rural counties, OR=1.01, 95% CI 0.65–1.55) (Schreinemachers, 2003).
Paternal occupational exposure, 2,4,5-T, inadequate evidence: See also paternal occupational
TCDD exposure discussed earlier.
Paternal occupational exposure, unspecified pesticides, limited evidence: Among offspring of
male licensed pesticide applicators in Minnesota, there was an increased risk of urogenital birth
defects compared to the general population (OR=1.69, 95% CI 1.06–2.64) (Garry et al., 1996).
This study also found increased risks of these defects among offspring of women living in high-crop
regions of Minnesota (high corn/soybean vs. noncrop regions, OR=1.56, 95% CI 1.29–1.89; high
wheat/sugar beet/potato regions, OR=2.25, 95% CI 1.67–3.03). A cohort study of Norwegian farm
families reported an association between kidney and other urinary tract birth defects and pesticide
expenditures on farms with orchards and/or greenhouses (OR=2.94, 95% CI 1.19–7.29); there was
no increased risk of such defects among all farm families (OR=0.82, 95% CI 0.53–1.27) (Kristensen
et al., 1997b). The heterogeneity of exposure indices precludes firm conclusions.
Outdoor air pollution Maternal exposure, inadequate evidence: A retrospective cohort study
in France reported an association between obstructive urinary tract defects and traffic density near
the prenatal residence (>50,000 vs. <10,000 vehicles/d, OR=2.10, 95% CI 0.94–4.70, p-trend=.07)
and an elevated risk of renal dysplasia (OR=1.75, 95% CI 0.73–4.20) (Cordier et al., 2004).
Drinking-water disinfection by-products Prenatal drinking-water DBP level, limited evidence:
Reviewers noted limited epidemiologic evidence for an association between urinary tract birth
defects and THM levels (Bove et al., 2002). Combined analysis of studies in Massachusetts and
Norway (Aschengrau et al., 1993; Magnus et al., 1999) indicated an association (summary
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423
OR=2.27, 95% CI 1.34–3.85) (Hwang & Jaakkola, 2003). In a Swedish retrospective cohort study,
maternal residence in communities using chlorinated drinking water was not associated with hypospadias (hypochlorite-treated vs. unchlorinated water, OR=1.1, 95% CI 0.6–2.0) or major kidney
malformations (OR=1.4, 95% CI 0.7–3.0) (Kallen & Robert, 2000). Haloacetonitriles produce
urogenital defects in experimental animals (Graves et al., 2001).
Incinerators Maternal exposure, inadequate evidence: In a large French retrospective cohort
study, renal dysplasia was associated with prenatal residence in communities with solid waste incinerators (OR=1.55, 95% CI 1.10–2.20) but not with expert-rated potential for exposure to dioxin,
metal, or dust emissions (high vs. low hazard, OR=1.30, 95% CI 0.57–2.97) (Cordier et al., 2004).
This study also found borderline associations between obstructive urinary tract birth defects and
prenatal residence in communities with solid waste incinerators (OR=1.22, 95% CI 0.90–1.65)
and expert-rated potential for exposure to dioxin, metal, or dust emissions (high vs. low hazard,
OR=1.93, 95% CI 0.94–3.93) (Cordier et al., 2004).
Solvents Maternal exposure, glycol ethers, inadequate evidence: In a multicenter European
case-control study, urinary tract defects were not associated with maternal 1st trimester occupational exposure to glycol ethers (OR=1.25, 95% CI 0.59–2.63) (Cordier et al., 1997).
Maternal exposure, unspecified solvents, inadequate evidence: A hospital-based case-control
study in France found no association between urinary tract defects and prenatal occupational
solvent exposure (OR=1.1, 90% CI 0.2–5.2) (Cordier et al., 1992).
Summary There is limited epidemiologic evidence for an association between male urinarytract birth defects and prenatal exposure to drinking water DBPs.
Male Genital Birth Defects A recent systematic review of epidemiological studies for
evidence linking indicators of prenatal serum levels of maternal estrogen with hypospadias and
cryptorchidism found no strong evidence to indicate that prenatal exposure to estrogens (including
some environmental exposures) was linked to disturbed development of the male reproductive
organs (Storgaard et al., 2006).
PCBs Maternal exposure, inadequate evidence: A small German case-control study observed
no association between cryptorchidism and infant adipose tissue PCB levels (median concentrations, cases vs. controls, 558 vs. 561 µg/kg lipid, p > 0.05) (Hosie et al., 2000).
TCDD Maternal exposure, inadequate evidence: A retrospective cohort study of births in
French communities with solid waste incinerators and unexposed comparison communities
reported no association between hypospadias and expert-rated hazard of incinerator emissions
(high vs. low exposure, OR=1.12, 95% CI 0.53–2.35, p-trend > .05) (Cordier et al., 2004).
Paternal exposure, 2,4,5-T or chlorophenate wood preservatives, inadequate evidence: In a
retrospective cohort study of Vietnam veterans, genital tract birth defects were not associated with
paternal serum TCDD levels categorized as low or high (OR=1.66, 95% CI 0.41–6.10) calculated
from data in paper) (Wolfe et al., 1995). Among British Columbia sawmill workers exposed to
chlorophenate wood preservatives known to be contaminated with TCDD and related toxicants,
genital-tract defects were weakly associated with hours of exposure during the 3 preconceptual
months (75th vs. 25th percentile, OR=1.29, 95% CI 0.9–1.5) (Dimich-Ward et al., 1996).
Pesticides Maternal exposure, chlorophenoxy herbicides, inadequate evidence: In a populationbased case-control study in Arkansas, hypospadias (cases identified from statewide registry) was
associated with agricultural use of the herbicide diclofop-methyl within 0.5 km of the maternal residence during GW 6–16, the window of genital-tract development (≥0.3 vs. 0 lb active ingredient,
OR=2.33, 95% CI 1.02–5.31) (Meyer et al., 2006).
Maternal exposure, other herbicides, inadequate evidence: The Arkansas study found no association between hypospadias and use within 0.5 km of the maternal residence during GW 6–16 of
alachlor (>0 vs. 0 lb, OR=0.56, 95% CI 0.35–0.89), atrazine (≥3.6 vs. 0 lb, OR=1.02, 95%
CI 0.58–1.79), dicamba (≥0.04 vs. 0 lb, OR=0.91, 95% CI 0.38–2.14), trifluralin (>8.5 vs. 0 lb,
OR=0.60, 95% CI 0.23–1.56), or diuron (>0 vs. 0 lb, OR=0.78, 95% CI 0.37–1.62) (Meyer et al.,
2006).
Maternal exposure, DDT/DDE, inadequate evidence: In a case-control study nested within the
U.S. Collaborative Perinatal Project pregnancy cohort, 3rd trimester maternal serum DDE levels
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D. T. WIGLE ET AL.
were not associated with cryptorchidism (per natural log serum DDE (µg/g lipid) increment,
OR=1.07, 95% CI 0.97–1.18) or hypospadias (OR=1.01, 95% CI 0.90–1.16) (Longnecker et al.,
2002). There was no association in a Mexico City case-control study between cryptorchidism and
maternal plasma DDE (>840 vs. ≤840 µg/g lipid, OR=0.48, 95% CI 0.15–1.60) or DDT levels (≥50
vs. <50 µg/g lipid, OR=1.13, 95% CI 0.24–5.29 (Flores-Luevano et al., 2003). A case-control study
nested within a California pregnancy cohort enrolled during 1959–1967 (when population serum
DDT/DDE levels were much higher than recently) revealed no association between cryptorchidism
and prenatal serum DDE (4th vs. 1st quartile, OR=1.34, 95% CI 0.51–3.48, p-trend=.75) or DDT
(OR=1.01, 95% CI 0.44–2.28, p-trend=.38) (Bhatia et al., 2005). This study also found no association between hypospadias and maternal serum DDE (OR=1.18, 95% CI 0.46–3.02, p-trend=.82)
or DDT (OR=0.79, 95% CI 0.33–1.89, p-trend=.30). A Finnish/Danish case-control study of cryptorchidism reported higher median breast milk DDT plus DDE concentrations in case compared to
control mothers (140.4 vs. 116.3 ng/g lipid, p=.27) (Damgaard et al., 2006). A small German casecontrol study reported a statistically nonsignificant relationship between cryptorchidism and infant adipose tissue p,p′-DDE levels (median concentrations, cases vs. controls, 265 vs. 170 µg/kg lipid, p > .05)
(Hosie et al., 2000). DDE and other androgen receptor antagonists (vinclozolin, procymidone,
linuron) produce feminization of prenatally exposed male rats with reduced anogenital distance
and induced areolas at low doses and hypospadias, retained nipples, undescended testes, and
epididymal agenesis at higher doses (Gray et al., 2001).
Maternal exposure, other organochlorine insecticides, inadequate evidence: A small German
case-control study observed associations between cryptorchidism and infant adipose tissue heptachlor (median concentrations, cases vs. controls, 5.2 vs. 2.4 µg/kg lipid, p=.01) and HCB levels
(61 vs. 20 µg/kg lipid, p=.01) (Hosie et al., 2000). The Finnish/Danish case-control study revealed
higher median breast milk levels for 17 of 21 organochlorine pesticides in case compared to control
mothers with differences being statistically significant for trans-chlordane (but not oxychlordane)
and of borderline significance for HCB and α-endosulfan (Damgaard et al., 2006).
Maternal exposure, other insecticides, inadequate evidence: The Arkansas study found no association between hypospadias and use within 0.5 km of the maternal residence during GW 6–16 of
carbaryl (>0 vs. 0 lb, OR=0.80, 95% CI 0.20–3.18) or permethrin (>0 vs. 0 lb, OR=0.37, 95% CI
0.16–0.86) (Meyer et al., 2006).
Maternal exposure, unspecified pesticides, inadequate evidence: Infants of women occupationally exposed to pesticides in Colombian semi-enclosed greenhouses had an elevated risk of cryptorchidism (OR=5.08, 95% CI 1.67–15.7) (Restrepo et al., 1990a). In a hospital-based case-control
study in Spain, there was a dose-response relationship between cryptorchidism and expert-rated
agricultural pesticide use intensity near the maternal residence (low vs. no use, OR=0.93, 95% CI
0.43–2.01; medium use, OR=1.56, 95% CI 0.72–3.38; high use, OR=2.32, 95% CI 1.26–4.29)
(Garcia-Rodriguez et al., 1996). In a very large Danish case-control study, maternal occupation in
farming or gardening was associated with cryptorchidism (OR=1.36, 95% CI 1.05–1.77) but not
hypospadias (OR=0.90, 95% CI 0.42–1.92) (Weidner et al., 1998). In a small Sicilian case-control
study, self-reported prenatal occupational or domestic pesticide exposure was related to a statistically nonsignificant elevated risk of cryptorchidism (OR=2.74, 95% CI 0.72–10.4) but not hypospadias (OR=0.42, 95% CI 0.05–3.56) (Carbone et al., 2007). Although suggestive, the heterogeneity of
exposure indices and inconsistent findings precludes firm conclusions.
A proportional birth prevalence study in the United Kingdom revealed no association between
hypospadias and maternal occupations likely exposed to pesticides (OR=0.84, 95% CI 0.50–1.41)
(Vrijheid et al., 2003). A retrospective cohort study in four Midwest states found no association
between male genital tract defects and prenatal residence in high-wheat counties, a proxy for chlorophenoxy herbicide exposure (high- vs. low-wheat counties, OR=1.03, 95% CI 0.51–2.09)
(Schreinemachers, 2003). In the Arkansas case-control study noted above, hypospadias was not
associated with total agricultural pesticide use within 0.5 km of the maternal residence during GW
6–16 (per 0.5 lb pesticide active ingredient, OR=0.82, 95% CI 0.70–0.96) (Meyer et al., 2006).
Paternal exposure, 2,4,5-T or chlorophenate wood preservatives, inadequate evidence: See
also paternal occupational TCDD exposure discussed earlier.
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Paternal occupational exposure, unspecified pesticides, inadequate evidence: As noted earlier,
a study of male licensed pesticide applicators in Minnesota reported an increased risk of urogenital
birth defects compared to the general population but did not include data for urological and genital
defects separately (Garry et al., 1996). Among Norwegian farm families, cryptorchidism was associated with pesticide expenditure (yes/no, OR=1.70, 95% CI 1.16–2.50); the association was
somewhat stronger for pesticide purchases on the subset of farms with field vegetables (OR=2.32,
95% CI 1.34–4.01) (Kristensen et al., 1997b). The latter study also reported associations between
hypospadias and tractor spray equipment (OR=1.38, 95% CI 0.95–1.99) and the use of such
equipment on farms with grain crops (OR=1.51, 95% CI 1.00–2.26). In a case-control study in the
Netherlands, paternal employment in occupations with likely pesticide exposure was associated
with cryptorchidism (OR=3.8, 95% CI 1.1–13.4) (Pierik et al., 2004).
In a very large Danish case-control study, paternal occupation in farming or gardening was not
associated with cryptorchidism (OR=1.08, 95% CI 0.94–1.23) or hypospadias (OR=1.15, 95% CI
0.83–1.58) (Weidner et al., 1998). A record-based retrospective cohort study in Norway revealed
no association between hypospadias and paternal occupation in agriculture (OR=0.68, 95% CI
0.34–1.23) (Irgens et al., 2000). A Sicilian case-control study revealed a statistically nonsignificant
association between cryptorchidism and self-reported preconceptual paternal employment in
agriculture (OR=2.45, 95% CI 0.63–9.59) but not with probable preconceptual pesticide exposure
based on job title (OR=0.60, 95% CI 0.21–1.74) (Carbone et al., 2007). This study also revealed
no association between hypospadias and paternal employment in agriculture (OR=1.61, 95% CI
0.29–9.01) or likely pesticide exposure (OR=1.07, 95% CI 0.42–2.73).
Outdoor air pollution Maternal exposure, inadequate evidence: A retrospective cohort
study in France observed no association between hypospadias and traffic density near the prenatal residence (>50,000 vs. <10,000 vehicles/d, OR = 1.33, 95% CI 0.59–3.01) (Cordier et al.,
2004).
Hazardous waste disposal sites Maternal exposure, inadequate evidence: A European casecontrol study reported an association between hypospadias and maternal residential proximity to
hazardous waste landfill sites (<3 vs. 3–7 km, OR=1.96, 95% CI 0.98–3.92) (Dolk et al., 1998).
A United Kingdom-wide retrospective cohort study revealed a weak association between NTDs and
prenatal residential proximity to any of 774 hazardous waste sites (≤2 vs. >2 km, OR=1.11, 95%
CI 1.03–1.21) (Elliott et al., 2001). A similar study limited to Scotland also found no association (≤2 vs.
>2 km from any of 61 sites, OR=0.84, 95% CI 0.58–1.22) (Morris et al., 2003).
Incinerators Maternal exposure, inadequate evidence: In a French retrospective cohort study,
hypospadias was not associated with prenatal residence in communities with solid waste incinerators (OR=0.88, 95% CI 0.66–1.19) or expert-rated potential for exposure to dioxin, metal or dust
emissions (high vs. low hazard, OR=1.12, 95% CI 0.53–2.35) (Cordier et al., 2004).
Solvents Paternal occupational exposure, inadequate evidence: A Norwegian cohort study of
offspring of male printers revealed no increased risk of cryptorchidism (compared to other occupations,
SIR=0.6, 95% CI 0.36–1.0) or hypospadias (OR=1.0, 95% CI 0.48–1.7) (Kristensen et al., 1993).
Phthalates Maternal exposure, inadequate evidence: In a pregnancy cohort study, there was
a dose-response relationship between reduced anogenital index (anogenital distance divided by
weight at examination, an indicator of incomplete masculinization) among male infants age 2–36
mo and prenatal urinary phthalate metabolite levels (e.g., mono-n-butyl phthalate ≥75th vs. <25th
percentile, OR = 10.2, 95% CI 2.54–42.2) (Swan et al., 2005). Other associations included
monobenzyl phthalate (OR= 3.8, 95% CI 1.03–13.9), monoethyl phthalate (OR = 4.7, 95%
CI 1.2–17.4), and monoisobutyl phthalate (OR = 9.1, 95% CI 2.3–35.7). The median maternal
urinary concentrations of phthalate metabolite concentrations associated with reduced anogenital index were below the U.S. 25th percentile for women. Although this was a well-designed
study, the findings require confirmation before a higher level of evidence can be assigned.
Several phthalates disrupt male genital-tract development in offspring of experimental animals
exposed during pregnancy (Foster, 2006).
Summary There was inadequate epidemiologic evidence for associations between male genital
birth defects and exposure to the environmental contaminants examined here.
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Neuropsychological Deficits
Although population-based statistics for neuropsychological deficits are not routinely
available, surveys indicate that several hundred thousand U.S. children have disabling childhood mental health conditions including mental retardation, learning disabilities, autism, and
attention deficit hyperactivity disorder (ADHD) (Halfon & Newacheck, 1999). An expert panel
estimated that 10% of neurobehavioral disorders (dyslexia, ADHD, autism, cognitive deficits,
and mental retardation) are attributable to environmental contaminants (National Academy of
Sciences, 2000b). The complex changes underlying brain growth and development during
pregnancy and childhood confer unique susceptibility to neurotoxins such as lead and methylmercury (MeHg) (Clarkson 2002; Mendola et al., 2002; National Academy of Sciences, 2000c;
Newland, 2002). The level of epidemiologic evidence for associations between developmental
milestones, cognitive function and problem behaviors and environmental factors is summarized
in Table 3.
Developmental Milestones
Lead Prenatal or childhood exposure, inadequate evidence: A large cross-sectional study
(based on NHANES II) reported an association between delayed onset of sitting, speaking and walking
and ranked order of current blood lead levels (Schwartz & Otto, 1987). The authors stated that
childhood blood lead levels reflect both prenatal and postnatal exposure.
Methylmercury In the following discussion, high-level MeHg exposure refers to doses
sufficient to cause clinical symptoms of poisoning in mothers and/or infants (as during Minamata
and Iraq episodes). Low-level exposure refers to background sources not linked to overt signs of
poisoning and does not imply low risk or safety.
High-level maternal exposure, sufficient evidence: Follow-up of 16 infants with congenital
MeHg poisoning at Minamata revealed that most could not understand maternal speech or walk at
age 3 yr; only half could walk alone and all had speech impairment at age 9–14 yr (Harada 1977).
Among children prenatally exposed to high doses of MeHg in Iraq, maternal hair mercury levels
were associated with delayed motor development and speech (Amin-Zaki et al., 1981). A WHO
expert panel (Marsh et al., 1981, 1987) concluded that extrapolation of the Iraqi data suggested the
TABLE 3. Role of Environmental Toxicants in Developmental Milestones, Cognitive Function, and Problem Behaviors
Toxicant
Exposure
Developmental milestones
Cognitive function
age 0–2 yr
Cognitive function
age ≥3 yr
Lead
Prenatal
Childhood
I
I
L
L
Prenatal
High-level—S
Low-level—I
High-level—L
High-level—S
Low-level—I
High-level—L
High-level—S
Low-level—L
I
I
I
L
High-level—S
Low-level—S
High-level—S
Low-level—L
I
I
I
I
High-level—S
Low-level—L
L
I
I
I
Methylmercury
Arsenic
Manganese
PCBs
DDT/DDE
Organophosphate
insecticides
Other herbicides
Environmental
tobacco smoke
Unspecified solvents
Childhood
Childhood
Prenatal
Childhood
Prenatal
Lactational
Prenatal
Lactational
Childhood
Paternal
Prenatal
Childhood
Prenatal
Note. TCDD, 2,3,7,8-tetrachlorodibenzo-p-dioxin.
L
I
Problem behaviors
L
I
I
High-level—L
I
I
L
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potential for increased risk of delayed developmental milestones at prenatal hair mercury levels as
low as 10–20 µg/g (World Health Organization 1990), a view shared by subsequent reviewers
(Clarkson, 2002; Newland, 2002).
Low-level maternal exposure, inadequate evidence: Extrapolation of dose-response data from
the Iraqi MeHg poisoning incidents indicates that neurodevelopmental milestone delay during
infancy may occur at prenatal hair mercury levels as low as 10 µg/g (Cox et al., 1989). However,
direct observation of populations with maternal hair mercury levels in the range of about 1–40 µg/g
have yielded generally negative results. A birth cohort study in the Faroe Islands observed no association between age at milestone development and prenatal hair (geometric mean 4.5 µg/g; 85th
percentile 10 µg/g) or cord blood mercury levels (finding stated without supporting data), although
early milestone development was associated with infant hair mercury level at age 12 mo (e.g., age
at sitting, Spearman r=–.10, p=.01) (Grandjean et al., 1995). Infant hair mercury at age 1 yr was
associated with breast-feeding duration, a predictor of early milestone attainment in this study. The
Faroe Islands study did not adjust for potential confounders such as maternal smoking. A small birth
cohort study in Peru reported no association between age at milestone development and prenatal
hair mercury levels (geometric mean 7.1 µg/g, range 0.9–29) (Marsh et al., 1995). In the Seychelles
Islands birth cohort study, age at walking was associated with maternal hair mercury levels (range
0.5–27 µg/g) in boys but not girls; age at talking was not associated with maternal hair mercury
(Axtell et al., 1998; Myers et al., 1997). The Peruvian and Seychelles studies only reported data on
p-values, not strength of associations.
High-level childhood exposure, limited evidence: Follow-up of 12 Iraqi infants postnatally
exposed to MeHg mainly through breast-feeding revealed that 4 had delayed motor milestones
(ages at sitting, standing, walking) and 8 had delayed language development (i.e., they did not
respond to simple verbal communication by age 2 yr even though hearing was not impaired) (Amin-Zaki
et al., 1981). The range of prenatal hair mercury concentrations was 2–384 µg/g and 40% of infants
had blood mercury ≥400 µg/L.
PCBs High-level maternal exposure, limited evidence: Compared to unexposed children, prenatally exposed Yucheng children were delayed on many of 32 developmental milestones and
experienced psychomotor and cognitive deficits at older ages (Guo et al., 2004).
Summary Epidemiologic evidence for the role of environmental toxicants in delayed developmental milestones includes: (a) sufficient—high-level prenatal MeHg exposure; (b) limited—high-level
childhood MeHg exposure, high-level prenatal exposure to PCBs, PCDFs, and related toxicants.
Cognitive Function: Children Age 0–2
Lead Maternal exposure, limited evidence: Several birth cohort studies found inverse doseresponse relationships between Bayley’s Mental Development Index (Bayley MDI) scores during
infancy and prenatal or cord blood lead levels adjusted for potential confounders (Bellinger et al.,
1984, 1988; Gomaa et al., 2002; Shen et al., 1998; Wasserman et al., 1992). The Sydney and Port
Pirie birth cohort studies reported no association between Bayley MDI at age 2 and prenatal or cord
blood lead (Cooney et al., 1989a; Wigg et al., 1988). In the Cincinnati birth cohort study, prenatal
blood lead was inversely associated with Bayley MDI at age 6 mo per log maternal blood lead increment, (β=–5.91 ± 2.68) but not at age 2 (β=3.30 ± 1.71) (Dietrich et al., 1990). In the Oswego
birth cohort study (consumers of Lake Ontario sports-caught fish), Fagan intelligence test scores at
ages 6 and 12 mo were not associated with cord blood lead levels (Darvill et al., 2000; Stewart
et al., 2000). However, a recent small birth cohort study in Atlanta reported an apparent inverse
association between Fagan intelligence scores at age 7 mo and very low prenatal blood lead levels
(all were below 5 µg/dl) (difference in mean maternal blood lead, 15th vs. 85th percentile Fagan
Scores, 0.94 ± 0.26 (SD) vs. 0.44 ± 0.15 µg/dl, t = 5.77, p < .001) (Emory et al., 2003). After adjustment for other prenatal and postnatal lead concentrations, the strongest lead exposure predictor of
Bayley MDI at age 2 yr in a Mexico City birth cohort was 1st trimester maternal blood lead (change
in Bayley MDI per 1 SD of loge lead concentration, β = –3.5, p = 0.03, there was also a statistically
non-significant inverse relationship with 1st trimester maternal plasma lead concentration (β = −2.4,
p = 0.19).
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Postnatal exposure, limited evidence: Three of 4 major birth cohort studies reported associations between Bayley MDI scores before age 3 yr and postnatal blood lead levels. The Boston birth
cohort study reported an inverse association between Bayley MDI scores at age 2 yr and blood lead
levels at ages 6–24 mo (Bellinger et al., 1988). In the Port Pirie birth cohort study, Bayley MDI
scores at age 2 yr were inversely associated with blood lead levels (µg/dl) at age 6 mo (β=–0.16,
p=.07) but not with levels at birth or ages 15 or 24 mo (Wigg et al., 1988). Bayley MDI scores at
age 2 yr were inversely associated with current blood lead levels in the Yugoslavia birth cohort study
(MDI vs. log current blood lead, β=–5.31 ± 2.44) (Factor-Litvak et al., 1999; Wasserman et al.,
1992). In a Mexico City birth cohort study, Bayley MDI scores at age 2 yr were inversely associated
with natural log current blood lead (β=–1.04, p < .01) (Tellez-Rojo et al., 2006) (see also maternal
exposure discussed earlier). In the Cincinnati birth cohort study, Bayley MDI scores at age 24 mo
were favourably associated with both prenatal and postnatal blood lead (change in MDI per log current blood lead increment, β=3.20 ± 1.70) (Dietrich et al., 1990). A small cross-sectional study in
New York City found Bayley scale deficits among children age 12–36 mo with current blood lead
levels at least 10 µg/dL (compared to <10 µg/dL, β=–6.2, 95% CI –10.8, –1.7) (Mendelsohn et al.,
1998).
Methylmercury High-level prenatal exposure, sufficient evidence: Among 15 mother–infant
pairs with high prenatal exposures during the 1972 Iraq MeHg poisoning episode, 40% (6 mothers,
6 infants) had clinically obvious neurotoxicity when examined less than 1 mo after exposure ended
(Amin-Zaki et al., 1974a). Severe manifestations included less than 3rd percentile head circumference (3 infants), total blindness (4), and severe hearing deficit (4). Most infant blood mercury levels
exceeded 200 µg/L; the blood mercury range among the 6 infants with severe neurotoxicity was
564–4220 µg/L. Repeat measurements indicated that infant blood mercury levels were about twice
those of their mothers at delivery and during follow-up to age 6 mo. When reexamined at age
1–7 yr, 22 Japanese children with high prenatal MeHg exposure (neonatal Minamata disease) all
had abnormal chewing, swallowing, speech, gait, and coordination, and most had abnormal tendon reflexes and involuntary movements (Harada, 1977). Among 9 Minamata cases examined at
age 3 yr, 6 children were unable to walk alone and 3 had no head control or ability to sit without
support. Several literature reviews concluded that high-level prenatal MeHg exposure causes severe
neurotoxic effects in infants, including mental retardation (Clarkson, 2002; Harada, 1978, 1995;
Harada et al., 1999; Myers & Davidson, 2000; Myers et al., 1998, 2000; National Academy of Sciences, 2000c; United Nations Environment Programme, 2002).
Low-level maternal exposure, inadequate evidence: Three of 5 birth cohort studies reported
no association between neuropsychological test scores before age 3 yr and prenatal mercury exposure indices in populations potentially exposed through consumption of marine or fresh water fish.
Among infants of Seychellois women with hair mercury levels greater than 12 µg/g compared to
3 µg/g or lower, mean Fagan visual recognition memory scores at age 6 mo were virtually identical
(60.5 ± 7.8 vs. 60.8 ± 7.6) (Myers et al., 1995). Similarly, there was no association between maternal hair mercury and Bayley MDI scores at 19 or 29 mo (Davidson et al., 1995). Among infants of
women who consumed Great Lakes fish, Fagan test scores of infant IQ at ages 6 and 12 mo were
not associated with maternal hair mercury levels (e.g., at age 12 mo, r=0.05, p=.42) (Darvill
et al., 2000). In the Avon Longitudinal Study of Parents and Children birth cohort study,
MacArthur Communicative Development Inventory and Denver Developmental Screening Test
subscale scores at age 15 mo were not associated with umbilical cord tissue mercury levels (e.g.,
change in vocabulary comprehension score per quartile change in umbilical cord tissue mercury,
β=6.1 ± 24.1[SE], p=.80) (Daniels et al., 2004). In a pregnancy cohort study in Massachusetts,
there was an inverse dose-response relationship between visual recognition memory novelty
preference scores at age 6 mo and 3rd trimester maternal hair mercury levels, independent of
prenatal fish consumption and several other potential confounders (change in score per unit
change in maternal hair mercury [µg/g], β=–7.5, 95% CI –13.7 to –1.2) (Oken et al., 2005). A
pregnancy cohort study of nonsmoking women in Poland reported inverse associations between
Bayley MDI or psychomotor development index (PDI) scores below 85 at age 1 yr and cord blood
mercury (≥ median vs. < median, OR=3.58, 95% CI 1.40–9.14) and maternal blood mercury
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levels (OR=2.82, 95% CI 1.17–6.79); note—this study did not report data for Bayley MDI and
PDI separately (Jedrychowski et al., 2006).
PCBs High-level maternal exposure, sufficient evidence: Compared to unexposed children,
Yucheng children had a statistically nonsignificant Bayley MDI score deficit at ages 6–30 mo (difference in mean Bayley MDI 6.0 points, t=1.67, p=.10; t and p calculated from data in paper)
(Rogan et al., 1988). A WHO expert group concluded that high-level prenatal exposure to PCBs,
PCDFs, and related compounds caused childhood cognitive deficits and persistent global developmental delays, mild behavior disorders, and hearing deficits (Brouwer et al., 1998; World Health
Organization, 2000).
Low-level maternal exposure, limited evidence: In the Michigan cohort, the Fagan test of visual
recognition memory at 7 mo was inversely associated with cord serum PCB levels (change in MDI
score per unit change in cord serum PCB, β=–0.35 points, F(1,76)=10.2, p < .005) (Jacobson
et al., 1985). The authors stated that, unlike the Bayley MDI, the Fagan test during infancy is predictive of childhood cognitive function, likely because it measures visual discrimination and short-term
memory, which are essential for information processing (Jacobson & Jacobson, 1996a). In the North
Carolina birth cohort, Bayley MDI scores at ages 6 and 12 mo were not associated with breast milk
PCB levels (change in MDI score at age 12 mo per unit change in breast milk PCB at birth [µg/g
lipid], β=–0.54 ± 0.54(SD), p=.32) (Gladen et al., 1988). A birth cohort study in the Netherlands
revealed a favourable but statistically nonsignificant relationship between Bayley MDI scores at age
7 mo and maternal plasma PCB (sum of 4 noncoplanar congeners) (change in MDI per natural log
increment of maternal plasma PCB, β=2.3 ± 1.7, p=.18) (Koopman-Esseboom et al., 1996). In a
German birth cohort, Bayley MDI at age 7 mo was inversely associated with breast milk PCB
concentrations at 2 and 4 wk after delivery (change in MDI per unit change in breast milk PCB,
β=–0.69 ± 0.41(SE), p=.05) but not with cord plasma PCB (change in MDI per unit change in
cord plasma PCB, β=0.06 ± 0.38(SE), p=.43) (Winneke et al., 1998). In the Oswego birth cohort,
Fagan Test of Infant Intelligence scores at ages 6 and 12 mo were not associated with breast milk
total PCBs (change in Fagan test score at age 12 mo per unit change in breast milk PCB, β=–0.075
points, t=–0.51, p=.30); this study did not assess cumulative lactational PCB exposure (Darvill
et al., 2000). However, the latter study observed inverse relationships between Fagan scores and
cord blood total PCBs (change in Fagan test score at age 12 mo per unit change in cord blood PCB,
F(1,207)=2.04, p=.08) and cord blood highly chlorinated PCBs (change in Fagan test score at age
12 mo per unit change in cord blood highly chlorinated PCB, F(1,207)=4.08, p=.02); the study
did not report β values for the latter analyses). In a German birth cohort, breast milk PCB concentrations about 2 wk postpartum were inversely associated with Bayley MDI scores at 7, 18, or 30 mo
(change in MDI scores at age 7, 18, or 30 mo per log2 breast milk PCB increment, β=–4.19, t=–1.99,
p=.025) (Walkowiak et al., 2001). Further analysis showed that compared to infants of mothers in
the 5th percentile of breast milk PCBs (≤173 ng/g lipid), infants of mothers in the 95th percentile
(≥679 ng/g lipid) had an average MDI score deficit of 8.3 points (95% CI 0.0 to 16.5). In the U.S.
Collaborative Perinatal Project birth cohort, Bayley MDI scores at age 7–10 mo were not associated
with 3rd trimester maternal serum PCB levels (change in MDI score at age 8 mo per unit increase in
maternal serum PCB (µg/L), β=0.10 ± 0.26(SE), p=.71) (Daniels et al., 2003).
Low-level lactational exposure, inadequate evidence: In the Michigan cohort, the Fagan test of
visual recognition memory at 7 mo was not associated with lactational PCB exposure (β not stated,
F(3,80)=1.18, p > .05) (Jacobson et al., 1985). In the North Carolina cohort, Bayley MDI scores at 6
and 12 mo were not associated with cumulative lactational PCB exposure from birth to age at test
(change in MDI score at age 12 mo per 1 mg cumulative breast milk PCB intake, β=–0.06 ±
0.16(SE), p=.70; this study did not include the Fagan test (Gladen et al., 1988). In the Dutch cohort,
Bayley MDI at age 18 mo was not associated with lactational PCB exposure (result stated without
supporting data) (Koopman-Esseboom et al., 1996). The Oswego birth cohort study reported no
association between Fagan Test of Infant Intelligence scores at ages 6 and 12 mo and PCB levels in
breast milk samples collected at 1–3 mo postpartum, an index of postnatal exposure (Darvill et al.,
2000). In the German birth cohort, cumulative lactational PCB dose was not associated with Bayley
MDI scores at 7, 18 or 30 mo (result stated without supporting data) (Walkowiak et al., 2001).
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Pesticides Maternal exposure, DDT/DDE, inadequate evidence: In a North Carolina birth
cohort study, there was a favorable relationship between Bayley MDI scores at age 6 mo and
breast milk DDE levels (β=0.65 ± 0.22, p=.004); Bayley MDI scores at age 12 mo were not
associated with breast milk DDE levels (Gladen et al., 1988). A birth cohort study in New York
State (offspring of women who consumed Lake Ontario fish) reported an inverse association
between cord serum DDE and Bayley MDI scores at age 12 mo (r=–.14, p=.3) but not at age
6 mo (r=–.09, p=.18); there was no adjustment for potential confounders (Darvill et al., 2000).
In a small Spanish birth cohort study in the region of an organochlorine production plant, Bayley
MDI scores at age 13 mo were inversely associated with cord serum DDE levels (β=–3.44 ± 1.39,
p < .05) (Ribas-Fito et al., 2003). A birth cohort study of Mexican farm workers in California
revealed no association between prenatal serum DDE and Bayley MDI scores at ages 6 or 12
mo and an inverse association of borderline statistical significance at age 24 mo (β=–2.44, 95% CI
–4.92 to 0.05) (Eskenazi et al., 2006).
Lactational exposure, DDT/DDE, inadequate evidence: The North Carolina birth cohort study
reported no association between Bayley MDI scores at ages 6 or 12 mo and cumulative lactational
DDE exposure (Gladen et al., 1988).
Summary Epidemiologic evidence for the role of environmental toxicants in cognitive deficits
among children age 0–2 includes: (a) sufficient evidence—high-level prenatal exposure to MeHg or
PCBs, PCDFs and related toxicants; (b) limited evidence—low-level prenatal exposure to lead or
PCBs, low-level childhood lead exposure.
Cognitive Function: Children Age 3 or Older
Lead It has been known for over 60 yr that high-level childhood lead exposure causes severe
neurotoxicity including persistent cognitive deficits (Byers & Lord, 1943). This section focuses on
the relationship between cognitive function and relatively low-level childhood lead exposure that
does not produce clinically obvious signs or symptoms.
Low-level prenatal exposure, limited evidence: Several birth cohort studies found that cognitive
scores among children age 3 yr or older were inversely associated with prenatal or cord blood lead
levels (Dietrich et al., 1990; Schnaas et al., 2006; Wasserman et al., 2000a). In the Mexico City
birth cohort study, IQ at age 6–10 yr was inversely associated with 3rd trimester maternal blood
lead (per natural log blood lead increment, β=–3.90, –6.45 to –1.36) but not with childhood
blood lead at age 1–5 yr (β=0.10, –3.88 to 4.06) or 6–10 yr (β=0.17, –1.41 to 1.76), in analyses
that adjusted for blood lead levels at other time periods (Schnaas et al., 2006). Other studies
reported no association between childhood cognitive function scores and prenatal or cord blood
lead levels (Tong et al., 1996).
Low-level childhood exposure, sufficient evidence: Several meta-analyses and literature reviews
covering studies published before 1996 found suggestive but inconclusive evidence for cognitive
deficits in children at relatively low childhood blood or tooth dentin lead levels (Agency for Toxic
Substances and Disease Registry, 1999b; Banks et al., 1997; Needleman & Gatsonis, 1990; Pocock
et al., 1994; Schwartz, 1994; Thacker et al., 1992). The estimated average full-scale IQ deficit for a
blood lead increment of 10 µg/dl was 2–3 points (Pocock et al., 1994; Schwartz, 1994). In a reanalysis
of two birth cohort studies (Bellinger et al., 1991; Needleman et al., 1979), there were inverse
dose-response relationships in nonparametric smoothing models between Bayley MDI scores and
blood lead levels and between WISC full-scale IQ scores and dentin lead levels cross the observed
ranges with no apparent thresholds (Schwartz, 1993).
Subsequent to the reviews noted above, longitudinal cohort studies in Port Pirie (Baghurst et al.,
1992; Tong et al., 1996, 1998), Yugoslavia/Kosovo (Factor-Litvak et al., 1999; Wasserman et al.,
1997, 2000a, 2003), Cincinnati (Ris et al., 2004), Mexico City (Schnaas et al., 2000), Rochester, NY
(Canfield et al., 2003), Boston (Bellinger & Needleman, 2003; Bellinger et al., 1992), Detroit
(Chiodo et al., 2004), and 4 U.S. cities (Chen et al., 2005) found significant inverse associations
between full-scale IQ among children age 3 yr or older and relatively low-level childhood blood
lead levels (current, previous or lifetime average). Findings included dose-response relationships
and independence from several potential confounders.
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The Rochester cohort study found relatively strong inverse associations between full-scale IQ at
ages 3 and 5 yr and blood lead at levels below 10 µg/dl, independent of several potential confounders
(Canfield et al., 2003). The IQ deficit per unit current blood lead increment (µg/dl) was greater
among the subgroup whose peak lifetime blood lead level was less than 10 µg/dl (β=–1.37 ± 0.60,
p=.03) compared to the whole group (β=–0.46 ± 0.15, p=.004). In a pooled analysis of 7 longitudinal cohort studies (involving 1333 children), there were inverse dose-response relationships
between IQ and early-childhood, peak, lifetime average, and current blood lead levels (Lanphear
et al., 2005). Over the blood lead range 2.4–30 µg/dl, the average adjusted IQ decrement
estimated using a log-linear model was 6.9 points (95% CI 4.2–9.4).
An important exception was the Mexico City birth cohort study in which IQ at age 6–10 yr was
inversely associated with 3rd trimester maternal blood lead (see earlier discussion) but not with
blood lead at age 1–5 yr (per natural log blood lead increment, β=0.10, –3.88 to 4.06) or 6–10 yr
(β=0.17, –1.41 to 1.76), in analyses that adjusted for blood lead levels at other time periods
(Schnaas et al., 2006). Recent reviews concluded that: (1) lead impairs behavioral and cognitive
development of children at blood lead levels below 10 µg/dl, (2) no blood lead threshold for such
effects has been demonstrated, (3) there appears to be a steeper slope for the inverse association
between IQ and blood lead below 10 µg/dl compared to higher levels, (4) lead accounts for 1–4%
of variance in cognitive ability whereas social and parenting factors account for at least 40%
(Bellinger, 2004; Koller et al., 2004; Lidsky & Schneider, 2003), and (5) the best fit for the relationship between childhood cognitive scores and blood lead concentrations is inverse log-linear (using
the natural log of blood lead concentration) (Rothenberg & Rothenberg, 2005).
A large cross-sectional study, based on almost 5000 children in NHANES III, found inverse
associations between scores on 4 cognitive test subscales (Arithmetic and Reading subtests of the
Wide Range Achievement Test-Revised and the Block Design and Digit Span subtests of WISC-R)
and current blood lead level, independent of several potential confounders (Lanphear et al.,
2000). Further analysis of children enrolled in NHANES III revealed inverse dose-response relationships between current blood lead and Wide Range Achievement Test-Revised subtest scores
for math (β=–0.57 ± 0.17) and reading (β=–0.80 ± 0.21) and WISC-III block design subtest
scores (β=–0.08 ± 0.03) after adjustment for several potential confounders including serum cotinine (Yolton et al., 2005). In the Kosovo birth cohort study, IQ at age 3–7 yr was inversely associated with both prenatal blood lead (β=–6.05 ± 1.35) and postnatal increases of at least 50% in
blood lead (Wasserman et al., 2000a). In a subsequent report of this study, there were inverse
dose-response relationships between full-scale IQ at age 10–12 yr and current blood and tibial
bone lead levels and lifetime average blood lead levels when modeled separately (Wasserman et
al., 2003). In models that included both bone and blood lead levels, full-scale IQ at age 10–12 yr
was inversely associated with bone but not blood lead levels.
Methylmercury Low-level prenatal MeHg exposure, limited evidence: Major birth cohort
studies have found inconsistent evidence of low-dose effects of MeHg on cognitive function among
children age 3 yr or older. Maternal hair mercury levels in Iraq ranged up to 674 µg/g but were
much lower among populations exposed mainly via consumption of fish with background MeHg
levels including the Faroe Islands (0.6–39.1 µg/g), the Seychelles (0.5–27 µg/g), Madeira (1.1–54 µg/
g), Massachusetts (0.02–2.38 µg/g), and Poland (0.1–3.40 µg/g). In the Seychelles Islands birth
cohort study, full-scale IQ at age 5 yr was not associated with prenatal or current child hair mercury
levels (range 0.9–26 µg/g) (Davidson et al., 1998). Also, there were no unfavorable associations
between mercury exposure indices and language, visual-spatial function or applied problem solving
scores. A re-analysis of the Seychelles Islands cohort data suggested a nonlinear dose-response relationship; McCarthy GCI scores at age 5–6 yr increased with prenatal hair mercury concentrations
up to 10 µg/g and then decreased (Axtell et al., 2000). Among 46 neuropsychological outcomes
measured in the Seychelles Islands study, only one (a test of fine motor function among boys using
their nonpreferred hand) was unfavorably associated with prenatal MeHg exposure; there were
favorable associations between maternal hair mercury levels and 2 outcomes (language function at
age 5 yr, attention deficit index at age 9 yr) (Davidson et al., 2004). Based on reanalyses of the
relationships between maternal hair mercury and scores on 26 neuropsychological tests in the
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Seychelles Islands cohort at age 9 yr, the estimated average benchmark dose 95% lower confidence
limit for maternal hair mercury was 20.1 µg/g (range 17.2–22.5) (van Wijngaarden et al., 2006).
In the Faroe Islands birth cohort study, cord blood mercury levels were unfavorably associated with test scores at age 7 yr for digit spans (average change in score per log cord blood
mercury increment, β=–0.27, p=.05), language (β=–2.03, p=.007), long-term memory (β=–0.99,
p = .03), and attention ( β = 39.3 ms, p < .001) (Grandjean et al., 1997). Further analyses of
maternal hair, cord blood, and current child hair mercury levels revealed that associations with
attention, language, and memory deficits were virtually limited to the prenatal exposure indices
(see also childhood exposure discussion later). After adjustment for cord tissue PCB levels, associations persisted between cord blood mercury and scores for memory (average change in longterm memory score per unit change in cord blood mercury (µg/L), β = –0.94, p = .04), attention
(average change in CPT reaction time (ms) per unit change in cord blood mercury, β = 40.3,
p = .0002) and language (Boston Naming Test, β = –1.94, p = .009) (Grandjean et al., 2001).
A reanalysis of the Faroe Islands birth cohort study, excluding children with vision deficits and
adjusting for visual contrast sensitivity, found that there was still no association between WISC
subscale scores at age 7yr and cord blood mercury levels (e.g., average change in WISC block
design score per unit change in cord blood mercury, β = –0.12, p = .25) (Grandjean et al., 2001).
Among the subgroup of children for whom mercury levels in 2 prenatal hair segments were available and consistent, there were inverse associations between cord blood mercury levels and test
scores at age 7 yr for WISC block design (average change in WISC block design score per unit
change in cord blood mercury, β = –0.28, p = .06), CPT reaction time (ms) (β = 47.0, p = .01),
long-term memory (β = –0.81, p < .04), Bender Gestalt copying errors (β = 1.53, p < .03), and
language (β = –1.59, p < .02) (Grandjean et al., 2003). When Faroese children were examined at
age 14 yr, there were inverse associations between maternal hair mercury and verbal (change per
doubling of mercury level, β = –6.87, p = .05) and attention (β = –9.54, p = .017) but not memory
scores (β = –3.05, p = .38) (Debes et al., 2006).
In the New Zealand birth cohort study, there were inverse associations (of borderline statistical
significance) between average prenatal hair mercury levels and test scores at age 6–7 yr including
WISC full-scale IQ (β=–0.42, 95% CI –1.1 to 0.18), McCarthy perceptual scale (β=–0.50, 95% CI
–0.92 to –0.077) and spoken language (β=–0.42, 95% CI –0.98 to 0.13) (Crump et al., 1998).
These authors estimated the benchmark doses and its 95% lower limit for average prenatal hair
mercury versus abnormal WISC full-scale IQ scores to be, respectively, 15 and 10 µg/g. Bench mark
dose was defined as the average prenatal hair concentration expected to cause a 10% increase in
abnormal (≤95th percentile) test scores. The Oswego birth cohort study reported an inverse association between McCarthy CGI scores at ages 38 and 54 mo and an interaction term modeled as
maternal hair mercury during the first half of pregnancy times cord blood PCB concentration
(change in CGI per unit change in mercury/PCB term, β=–0.50, p=.008); there was a weaker
nonsignificant inverse association with maternal hair mercury during the second half of pregnancy
times cord blood PCB (β=–0.31, p=.16) (Stewart et al., 2003).
Expert panels concluded that the associations between language, verbal memory, and other
subtle neuropsychological deficits and low-level prenatal MeHg exposure from fish consumption
observed in epidemiologic studies are consistent with neuropsychological deficits in experimental
animals after low-dose prenatal MeHg exposure (National Academy of Sciences, 2000c; United
Nations Environment Programme, 2002).
Childhood exposure, inadequate evidence: A cross-sectional study in Germany revealed no
association between vocabulary, block design or CPT reaction times and current urinary mercury
excretion rates (range 0.02–2.83 µg/d) (Walkowiak et al., 1998). In the Faroe Islands cohort,
attention scores (but not language or memory scores) were inversely associated with child hair
mercury levels at age 1 yr (Grandjean et al., 1999). In a cross-sectional study of children age 7–12 yr
in the Amazon basin, there were inverse associations between current hair mercury and Stanford–
Binet subscales scores for visuospatial function (change in score per unit change in log hair mercury,
β=–6.2, p < .001), memory (β=–2.9, p < .001) and attention functions (β=–0.9, p < .001)
(Grandjean et al., 1999).
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Arsenic Childhood exposure, inadequate evidence: A small cross-sectional study of children
age 6–9 yr in Mexico found an inverse association between full-scale IQ and urinary arsenic levels
among the subgroup of children living in a smelter town (partial correlation coefficient r=–0.33,
p=.04) but not in the whole group including children from an unexposed comparison town (r=–0.15,
p=.22) (Calderon et al., 2001).
Manganese Prenatal exposure, inadequate evidence: A birth cohort study in France observed
no association between cognitive function scores at ages 9 mo to 6 yr and prenatal manganese levels
in maternal blood or hair, cord blood, or placenta; attention, nonverbal memory, and hand skill
scores at age 3 yr were inversely associated with cord blood manganese levels (geometric mean
38.6 µg/L, range 22.0–67.7) (Takser et al., 2003). The geometric mean newborn hair manganese
concentration was 0.77 µg/g (range 0.22–4.25).
Childhood exposure, inadequate evidence: A small cross-sectional study of Chinese children
including some from a region with elevated drinking-water manganese levels reported that most
neuropsychological scores (manual dexterity, digit span, digit symbol, Benton visual retention, and
pursuit aiming tests) were inversely associated with current hair manganese levels (mean 1.25 µg/g)
(He et al., 1994).
PCBs High-level maternal exposure, sufficient evidence: Compared to unexposed children,
Yucheng children age 6–7 yr had lower WISC full-scale IQ scores at age 6–7 yr (mean IQ, exposed
vs. unexposed, 84 vs. 88, t=1.06, df=40, p=.29) (Rogan et al., 1988). Yucheng children also had
increased auditory event-related P300 potential latencies and reduced P300 amplitudes; such
neurophysiological changes have been linked to cognitive deficits, attention deficit disorder and
reading disability (Chen & Hsu, 1994). Reviewers concluded that prenatal high-level maternal
exposure to PCBs, PCDFs, and related toxicants were associated with persistent cognitive function
deficits (Longnecker et al., 1997) and a WHO expert group concluded that this is a causal relationship (Brouwer et al., 1998).
Background maternal exposure, limited evidence: Reviewers noted that there were inverse
associations between cognitive function scores and prenatal PCB exposure indices in four of the five
studies that were published by 1999 and assessed this relationship (Ribas-Fito et al., 2001). Other
reviewers concluded that the growing weight of evidence from epidemiologic studies supports an
inverse association between childhood cognitive scores and prenatal PCB exposure levels but noted
that there have been few attempts to assess the role of specific PCB congeners or classes of congeners (Schantz et al., 2003). In the Michigan birth cohort, McCarthy GCI scores at age 4 were not
associated with cord serum PCB (change in GCI per unit change in cord serum PCB, β=–0.11,
p=.22) (Jacobson et al., 1990a). Follow-up of the Michigan cohort at age 11 yr revealed an inverse
dose-response relationship between WISC-R full-scale IQ and breast milk PCB levels at birth (IQ vs.
5 categories of increasing breast milk PCBs, β=–0.17, p=.02) (Jacobson & Jacobson, 1996b). This
study also reported an inverse association of borderline statistical significance between executive
function scores at age 11 yr and prenatal PCB exposure (change in Stroop Color-Word test score
per unit change in prenatal PCB, β=–0.15, p < .10) (Jacobson & Jacobson, 2003). The latter study
observed no association between another test of executive function at age 11yr and prenatal PCB
exposure (change in Wisconsin card sort categories completed per unit change in prenatal
PCB, β=–0.04, p > .05). In the Dutch cohort, Kaufman ABC scores at age 42 mo were inversely
associated with maternal 3rd trimester plasma PCBs (per natural log PCB increment, β=–4.56 ± 1.62,
p=.005); this association was stronger among the subgroup of formula-fed infants (β=–8.69 ± 2.49,
p=.0006) than the breastfed group (β=–2.20 ± 2.14, p=.30 (Patandin et al., 1999). McCarthy
GCI scores at age 7 yr in the Dutch cohort were not associated with 3rd trimester maternal plasma
PCBs (per natural log PCB increment, β=–0.14 ± 1.58) (Vreugdenhil et al., 2002). This study
reported inverse associations between Tower of London test scores at age 9 yr and maternal plasma
PCBs (change in score, 4th vs. 1st quartile maternal plasma PCB, β=–1.85 ± 0.67(SE), p=.007)
(Vreugdenhil et al., 2004a). In the German birth cohort, breast milk PCB concentrations about 2 wk
postpartum were inversely associated with Kaufman ABC scores at age 42 mo (per log2 breast milk
PCB increment, β=–4.30, t=–1.93, p=.03) (Walkowiak et al., 2001). In the Oswego Newborn
and Infant Development cohort, cord blood highly chlorinated PCB concentrations were inversely
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associated with McCarthy GCI scores at 38 mo (linear trend test, F(1,165)=7.33, p=.008) but not
at 54 mo (linear trend test, F(1,166)=1.25, p > .05) (Stewart et al., 2003). In the U.S. Collaborative
Perinatal Project cohort, WISC full-scale IQ at age 7 yr was not associated with 3rd trimester maternal
serum PCB in a fully adjusted model (change in IQ per unit change in maternal serum PCB (µg/g
lipid), β=1.90 ± 1.92(SE), p > .05) (Gray et al., 2005).
Lactational exposure, limited evidence: In the Michigan birth cohort, McCarthy GCI scores at
age 4 yr were not associated with current serum PCB (stated without supporting data) (Jacobson
et al., 1990a). This study reported no association between executive function scores at age 11 yr
and postnatal PCB exposure (e.g., change in Wisconsin card sort categories completed per unit
change in serum PCB at age 4 yr, β=0.01, p > .05) (Jacobson & Jacobson, 2003). In the German
birth cohort, lactational PCB exposure was inversely associated with Kaufman ABC scores at age
42 mo (change in Kaufman-ABC index per unit change in current serum PCB, t=–2.01, p=.03;
β-coefficient not reported) (Walkowiak et al., 2001).
General findings Reviewers concluded that prenatal and lactational exposure to PCBs,
PCDDs, and PCDFs can cause neuropsychological and neuromotor deficits in humans and in
experimental animal with LOAELs in the range of background general population dioxin-TEQ body
burdens (Brouwer et al., 1995). Cord plasma and breast milk noncoplanar PCB congener levels
have also been associated with cognitive deficits in children age 6–7 yr (Vreugdenhil et al., 2002).
Such congeners were the most potent in reducing dopamine content and disrupting calcium
metabolism in neurons in vitro (Tilson & Kodavanti, 1997). Inconsistent findings in epidemiologic
studies may arise from PCB exposure intensity differences; for instance, average prenatal serum
PCB-153 levels in 10 epidemiologic studies of neurodevelopment varied from 30 to 450 ng/g lipid,
being highest in the Faroe Islands birth cohort study and lowest in two U.S. studies (Longnecker
et al., 2003).
Pesticides Maternal exposure, DDT/DDE, inadequate evidence: In the North Carolina birth
cohort, McCarthy CGI scores at ages 3, 4, and 5 yr were not obviously associated with breast milk
DDE levels (results were displayed in graphs without statistical analysis) (Gladen & Rogan, 1991).
Lactational exposure, DDT/DDE, inadequate evidence: The North Carolina study also reported
no association between McCarthy CGI scores and cumulative lactational DDE exposure (Gladen &
Rogan, 1991).
Childhood exposure, organophosphate pesticides, inadequate evidence: In a small crosssectional study of children age 2–11 yr living in homes where methyl parathion had been illegally
used for pest control in Mississippi and Ohio, cognitive function scores were similar to those in a
comparison group of unexposed children (result stated without supporting data); methyl parathion
exposure status (yes/no) was based on urinary p-nitrophenol and environmental wipe methyl
parathion levels above defined limits (Ruckart et al., 2004).
Environmental tobacco smoke Childhood exposure, limited evidence: Expert panels found
limited evidence of an inverse association between childhood cognitive function and postnatal
ETS exposure (California Environmental Protection Agency, 2005; U.S. Department of Health and
Human Services, 2006). The panel noted that this relationship has been much less studied than
prenatal active smoking and that smoking mothers tend to smoke both during and after
pregnancy—thus associations with childhood maternal smoking may partially reflect the impact of
prenatal smoking. Among children age 6–16 yr enrolled in NHANES III, there were inverse
dose-response relationships between log serum cotinine and math (β=–1.93 ± 0.70), reading
(β=–2.69 ± 0.75) and visual construction scores (β= –0.55 ± 0.12) (Yolton et al., 2005). In a
pregnancy cohort study, Bayley MDI scores at age 3 yr were inversely associated with 3rd trimester
maternal personal air PAH exposure (>4.16 vs. ≤4.16 ng/m3, β=–5.69, p < .01) (Perera et al.,
2006). However, ETS is only one source of PAH exposure (e.g., other sources include outdoor air
pollution and diet).
Solvents Maternal exposure, unspecified solvents, inadequate evidence: A small retrospective
cohort study in Canada found no association between full-scale, verbal or performance IQ scores at
ages 6–8 yr and maternal 1st trimester occupational organic solvent exposure (Laslo-Baker et al.,
2004).
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Summary Epidemiologic evidence for the role of environmental toxicants in cognitive deficits
among children age 3 yr or older includes: (a) sufficient evidence—high- or low-level childhood
lead exposure; high-level prenatal exposure to MeHg or to PCBs, PCDFs, and related toxicants;
(b) limited evidence—prenatal low-level exposure to lead, MeHg, or PCBs; low-level lactational
exposure to PCBs; childhood ETS exposure.
Problem Behaviors Problem behaviors include hyperactivity, distractibility, impulsivity, and
impersistent and aggressive behaviors.
Lead Childhood exposure, limited evidence: Several birth cohort studies found fairly consistent associations between teacher- and/or mother-reported problem behaviors (e.g., inattention,
easily distracted, aggressiveness) among school-age children and current or lifetime average blood
lead or tooth dentin lead levels, independent of potential confounders (Bellinger et al., 1994; Burns
et al., 1999; Chiodo et al., 2004; Dietrich et al., 2001; Factor-Litvak et al., 1999; Leviton et al.,
1993; Needleman et al., 1996; Silva et al., 1988; Wasserman et al., 1998). Two large U.S. crosssectional studies reported associations between problem behaviors and current blood lead levels
(Needleman et al., 1979; Schwartz & Otto, 1987). A large European cross-sectional study in which
the blood lead range was 5–60 µg/dl observed no association with problem behaviors (Winneke
et al., 1990). Two small cross-sectional studies reported associations between problem behaviors
and dentin or current blood lead levels (Hansen et al., 1989; Mendelsohn et al., 1998) but several
others did not (Harvey et al., 1988; Landrigan et al., 1975; Lansdown et al., 1986; Winneke et al.,
1983). A case-control study in Boston observed an association between court-adjudicated delinquency and current tibial bone lead levels, independent of race and other potential confounders
(OR=3.7, 95% CI 1.3–10.5) (Needleman et al., 2002). In sum, these studies provide moderately
strong evidence of an association between problem behaviors and childhood lead exposure.
Methylmercury Maternal exposure, inadequate evidence: The Seychelles Islands study
appears to be the only cohort to have assessed the relationship between aggressive and other problem
behaviors and MeHg exposure indices. Inattentive, aggressive, and other problem behaviors at age
5–6 yr (identified by parents using a checklist) were not associated with prenatal hair mercury levels
(Myers et al., 2000). When reassessed at age 9 yr, prenatal hair mercury levels were inversely associated with the Connor’s teacher rating scale hyperactivity index (β=–0.0067 ± 0.0023, p=.004)
and were not related to attention scores (Myers et al., 2003).
Childhood exposure, inadequate evidence: In the Seychelles Islands study, inattentive, aggressive and other problem behaviors at age 5–6 yr were not associated with current child hair mercury
levels (e.g., for Myers et al., 2000).
PCBs High-level maternal PCB/PCDF exposure, limited evidence: Four reviews of epidemiologic studies noted that high-level maternal exposure to PCBs, PCDFs, and related toxicants during
the Yusho and Yucheng incidents was associated with persistent problem behaviors including
hyperactivity (Brouwer et al., 1998; Guo et al., 2004; Longnecker et al., 1997; Schantz, 1996).
There appears, however, to have been no association between problem behaviors and prenatal or
postnatal PCB exposure indices (stated without supporting data) (Chen et al., 1994).
Background maternal PCB exposure, inadequate evidence: A Michigan birth cohort study
found an inverse association between freedom from distractibility at age 11 yr and prenatal but not
postnatal PCB exposure levels (change in WISC-R subscale for freedom from distractibility per unit
change in cord or maternal serum PCB level (5 categories), β=–0.17, p=.02) (Jacobson & Jacobson,
1996b). Further follow-up of this cohort revealed no association between reaction time and prenatal
or postnatal PCB levels in the whole group; however, there was an inverse association between
mother-reported attentiveness and prenatal PCB levels among the subgroup of infants who were
breastfed less than 6 wk (change in score per unit change in prenatal PCB level, β=–0.39, p < .05)
(Jacobson & Jacobson, 2003). The Faroe Islands birth cohort study found an inverse association
between attention scores (measured using the continous performance test) at age 7 yr and umbilical
cord tissue PCB levels (change in reaction time per log increment of cord tissue levels of 3 noncoplanar PCBs, β=–7.2, p=.45, adjusted for cord blood mercury) (Grandjean et al., 2001). In the
Rotterdam component of the Dutch birth cohort, sustained attention scores at age 9 yr were
inversely associated with maternal plasma PCB levels (change in reaction time, 4th vs. 1st quartile
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D. T. WIGLE ET AL.
maternal plasma levels of 4 noncoplanar PCBs, β=20.4 ± 14.0 ms, p=.15) (Vreugdenhil et al.,
2004a).
Lactational PCB exposure, inadequate evidence: The Michigan birth cohort study found no
association between freedom from distractibility at age 11 yr and serum PCB at age 4 yr (an index
of lactational exposure) when adjusted for prenatal serum PCB levels; the report does not give data
for β related to freedom from distractibility at age 11 yr vs. serum PCB at age 4 yr (Jacobson &
Jacobson, 1996b). In the Rotterdam component of the Dutch birth cohort, sustained attention
scores at age 9 yr were not associated with breast-feeding duration (change in reaction time, breastfed long vs. short, β=1.53 ± 15.7 ms, p=.92) (Vreugdenhil et al., 2004a).
Pesticides Paternal occupational exposure, other herbicides, inadequate evidence: A crosssectional study of offspring of licensed pesticide applicators reported an association between parent-reported attention deficit hyperactivity disorder and paternal use of the herbicide glyphosate
(OR=3.6, 95% CI 1.35–9.65) (Garry, 2002).
Environmental tobacco smoke Maternal exposure, inadequate evidence: A WHO expert
group found inadequate evidence for an association between childhood behavioral problems and
prenatal ETS exposure (World Health Organization, 1999).
Childhood exposure, limited evidence: A WHO expert group found inadequate evidence for
an association between childhood behavioral problems and childhood ETS exposure (World Health
Organization, 1999). A more recent expert panel review concluded that there is limited evidence of
an association between childhood problem behaviors and childhood ETS exposure but noted that
this may partially reflect the impact of prenatal active smoking (California Environmental Protection
Agency, 2005).
Summary Epidemiologic evidence for the role of environmental toxicants in problem behaviors includes limited evidence for high-level prenatal exposure to PCBs, PCDFs, and related toxicants and childhood exposure to lead or ETS.
Motor Function: Children Age 0–2 The level of epidemiologic evidence for associations
between motor and sensory function and environmental factors is summarized in Table 4.
Lead Prenatal or early childhood exposure, inadequate evidence: Several birth cohort studies
found no association between PDI scores at ages 6 mo to 2 yr and cord blood or childhood blood
lead levels (Bellinger et al., 1984; Cooney et al., 1989a, 1989b; Ernhart et al., 1988; Wigg et al.,
1988). In a Mexico City birth cohort, PDI scores were inversely associated with current blood lead
(β= –1.18, p < .01) at age 2 yr but not at age 1 yr (Tellez-Rojo et al., 2006).
TABLE 4. Role of Environmental Toxicants in Motor and Sensory Function
Toxicant
Lead
Methylmercury
Exposure
Prenatal
Childhood
Prenatal
Childhood
Cadmium
PCBs
DDT/DDE
Organophosphate
insecticides
HCB
j
Childhood
Prenatal
Lactational
Prenatal
Lactational
Childhood
Prenatal
Childhood
Motor function,
age 0–2
Motor function,
age ≥3
I
I
High-level—S
Low-level—L
High-level—L
I
L
L
High-level—Sj
Low-level—L
I
I
I
I
I
I
I
I
I
I
Neonatal hypotonia, delayed psychomotor development during infancy.
Auditory function
L
L
High-level—S
Low-level—L
I
High-level—L
Low-level—L
I
Visual function
I
High-level—S
Low-level—I
High-level—S
Low-level—I
High-level—I
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Methylmercury High-level prenatal exposure, sufficient evidence: A review of congenital
MeHg poisoning at Minamata noted that most infants age 1–2 yr had abnormal reflexes, increased
muscle tone, and displayed involuntary movements (Harada, 1977). Delayed motor development
of Iraqi infants was associated with high prenatal hair and childhood blood mercury levels
(Amin-Zaki et al., 1981; Marsh et al., 1980, 1981). Extrapolation of Iraqi data suggested increased
risk of delayed motor development at prenatal hair mercury levels as low as 10–20 µg/g (World
Health Organization, 1990), a view shared by recent reviewers (Clarkson, 2002; Newland, 2002).
Low-level prenatal exposure, limited evidence: Among Seychellois infants age 6 mo, there was
a statistically nonsignificant elevated risk of abnormal or questionable neurologic signs (deep tendon
reflexes, limb tone) among infants of women with hair mercury exceeding 9 µg/g (compared to
≤6 µg/g, crude OR=1.67, 95% CI 0.49–5.63, calculated from data in paper) (Myers et al., 1995).
Among infants of women who consumed Great Lakes fish, abnormal reflex scores on postnatal day
2 were not associated with maternal hair mercury levels during early or later pregnancy (e.g.,
change in score per unit change in hair mercury during gestation month 5–9, β=–0.019, p=.76)
(Stewart et al., 2000). A pregnancy cohort study of nonsmoking women in Poland found inverse
associations between Bayley MDI or PDI scores below 85 at age 1 and cord blood mercury
(≥median vs. <median, OR=3.58, 95% CI 1.40–9.14) and maternal blood mercury levels
(OR=2.82, 95% CI 1.17–6.79); this study did not report data for Bayley MDI and PDI separately
(Jedrychowski et al., 2006).
Childhood high-level exposure, limited evidence: Among children exposed at age 0–14 yr
during the Iraq MeHg poisoning episodes and followed for 2 yr, the severity and persistence of
ataxia and muscle weakness were associated with baseline blood mercury levels (Amin-Zaki et al.,
1978). Children with mild/moderate poisoning improved slowly but all had persistent hyperreflexia;
7 of 18 children with severe poisoning had persistent physical disability.
PCBs Neonatal hypotonia High-level exposure, sufficient evidence: A WHO working group
concluded that high-level prenatal exposure to PCBs, PCDFs, and related compounds caused
neonatal hypotonia (Brouwer et al., 1998).
Background exposure: A recent review concluded that the evidence from birth cohort studies
suggests an association between neonatal hypotonia and hyporeflexia and background PCB
exposure levels (Longnecker et al., 1997). A review of seven cohort studies noted that abnormal
neonatal reflexes were associated with PCB exposure in all four studies that assessed it (Ribas-Fito
et al., 2001).
Motor function, age 0–2 High-level maternal PCB/PCDF exposure, limited evidence: Compared to unexposed children, Yucheng children age 6–30 mo had lower Bayley Scale psychomotor
development index (PDI) scores (mean PDI score, exposed vs. unexposed children, 101 ± 2.7(SE)
vs. 108 ± 2.1, mean difference t=2.05, df=88, p=.04 (t and p calculated from data in paper)
(Rogan et al., 1988). Further investigation showed that Yucheng children age 7 yr had small and
statistically nonsignificant gross (exposed vs. unexposed, Chinese Child Developmental Inventory
(CCDI) scores, 27.4 ± 0.6(SE) vs. 28.2 ± 0.6, t=–1.5, p=.14) and fine motor function deficits
(exposed vs. unexposed, CCDI scores, 31.5 ± 1.0(SE) vs. 32.6 ± 0.9, t=–1.3, p=.20) (Guo et al.,
1994; Lai et al., 1994, 2001)
Low-level maternal PCB exposure, limited evidenc: A WHO working group concluded that
childhood psychomotor function was inversely associated with low-level prenatal exposure to PCBs
and related compounds but noted that most of the individual neuropsychological test results were
within normal limits (Brouwer et al., 1998). Among reviewed studies, a North Carolina birth cohort
study revealed inverse associations between Bayley PDI scores at ages 6 and 12 mo and breast milk
PCBs soon after birth (an index of prenatal PCB exposure) (change in PDI score at age 6 mo per unit
change in breast milk PCB (µg/g lipid), β=–0.96 ± 0.46, p=.04; at age 12 mo, β=–1.34 ± 0.61,
p=.03) (Gladen et al., 1988; Rogan & Gladen, 1991). The PDI deficits at ages 18 and 24 mo of
infants of women with the highest breast milk PCB levels, compared to those with the lowest levels,
were not statistically significant (18 mo, deficit = –4.0 ± 3.9(SE); 24 mo,deficit –7.9 ± 4.5) (Rogan &
Gladen, 1991). In a Dutch birth cohort, suboptimal neonatal neurologic scores at 10–21 d after
birth (based in part on reflexes and postural tone) were associated with breast milk PCB-TEQ and
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PCDD/PCDF-TEQ levels (per doubling of breast milk PCB-TEQ levels (pg/g lipid), OR=3.21, 95%
CI 1.37–7.48; per doubling of breast milk PCDD/PCDF-TEQ, OR=3.12, 95% CI 1.36–7.18) but
not with maternal or cord serum individual or aggregate levels of 4 noncoplanar PCB congeners
(per doubling of maternal or cord serum noncoplanar PCBs, OR=1.11, 95% CI 0.74–1.65)
(Huisman et al., 1995a). At this age, maternal and cord serum and breast milk PCB levels all reflect prenatal
exposure. Continued follow-up of the whole Dutch birth cohort revealed a persistent inverse association between neurologic optimality scores at age 18 mo and cord plasma levels of 4 noncoplanar
congeners (change in score per log increment of cord serum PCBs, β=–0.149 ± 0.049, p=.003)
(Huisman et al., 1995b). Within the Rotterdam component of the Dutch cohort, there was an
inverse association between Bayley PDI scores at age 3 mo and maternal plasma levels of 4 noncoplanar PCBs (per natural log PCB increment, β=–4.8 ± 2.0, p=.02) but not at age 7 mo (β=2.3 ±
1.7, p=.18) (Koopman-Esseboom et al., 1996). A small German birth cohort study observed no
association between Bayley PDI scores at age 7 mo and cord plasma levels of 3 noncoplanar PCBs
(change in PDI per log increment of cord plasma PCBs, β=0.009 ± 0.63, p > .9) (Winneke et al.,
1998). A small Faroe Islands birth cohort study found no association between neonatal optimality
scores at age 2 wk and maternal serum or breast milk PCB levels (Spearman’s rank correlation coefficients, scores vs. maternal serum or breast milk PCBs, R=0.03 and –0.03, respectively, p > .05)
(Steuerwald et al., 2000). The Oswego birth cohort study observed an association between abnormal reflexes on postnatal day 2 and cord blood levels of highly chlorinated PCB congeners (F(1,262
df)=2.81, p=.095) (Stewart et al., 2000). In a German birth cohort, Bayley PDI scores at ages 7–30
mo were inversely associated with breast milk PCBs, an index of prenatal exposure (per log2 increment, β=–4.61, t=–2.22, p=.015) (Walkowiak et al., 2001). There was no association between
3rd trimester maternal serum PCBs (11 congeners) and Bayley PDI at age 8 mo in the U.S. Collaborative Perinatal Project (change in Bayley PDI per unit change in maternal serum PCBs (µg/L),
β=0.47 ± 0.32, p=.14) (Daniels et al., 2003). In a small Spanish birth cohort, there was a statistically nonsignificant inverse association between Bayley PDI at age 13 mo and cord serum PCBs
(7 congeners) (change in Bayley PDI per doubling of cord serum PCB (µg/L, β=–2.84 ± 1.72, p < .10)
(Ribas-Fito et al., 2003). The U.S. study was much larger than the Spanish cohort and included
women recruited during 1959–1966 when population serum PCB levels were substantially higher
than recently (median and 95th percentiles were 2.7 and 6.3 µg/L).
Low-level lactational exposure, inadequate evidence: A North Carolina birth cohort study
revealed no association between Bayley PDI scores at ages 6 and 12 mo and cumlulative lactational
PCB exposure (change in PDI score at age 6 mo per mg PCB ingested, β=–0.27 ± 0.20, p=.17; at
age 12 mo, β=–0.27 ± 0.18, p=.13) (Gladen et al., 1988; Rogan & Gladen, 1991). In a Dutch
birth cohort, suboptimal neonatal neurologic scores at 10–21 d after birth (based in part on reflexes
and postural tone) were associated with breast milk PCB-TEQ and PCDD/PCDF-TEQ levels (per
doubling of breast milk PCB-TEQ levels (pg/g lipid), OR=3.21, 95% CI 1.37–7.48; per doubling of
breast milk PCDD/PCDF-TEQ, OR=3.12, 95% CI 1.36–7.18) (Huisman et al., 1995a). Within the
Rotterdam component of the Dutch cohort, there was an inverse association of borderline statistical
significance between Bayley PDI scores at age 3 mo and breast milk PCB-TEQ (per natural log
increment, β=–7.4 ± 4.0, p=.07), adjusted for breastfeeding duration (Koopman-Esseboom et al.,
1996). At age 7 mo, PDI scores were inversely and significantly associated with breast milk PCB-TEQ
levels, adjusted for breastfeeding duration (medium vs. lowest category, β=–9.5 ± 3.9, p=.01;
highest vs. lowest category, β=–7.7 ± 4.9, p=.12) (Koopman-Esseboom et al., 1996). A small German
birth cohort study observed an statistically nonsignificant inverse relationship between Bayley PDI
scores at age 7 mo and breast milk levels of 3 noncoplanar PCBs (change in PDI per log increment
of breast milk PCBs, β=–0.71 ± 0.63, p=.13, adjusted for breastfeeding duration) (Winneke et al.,
1998). In the German birth cohort, Bayley PDI scores at age 7–30 mo were not associated cumulative lactational PCB dose (stated without supporting data) (Walkowiak et al., 2001).
Pesticides Maternal exposure, DDT/DDE, inadequate evidence: In the North Carolina cohort
study, neonatal hyporeflexia was associated with breast milk DDE levels (≥5 vs. <2 µg/g lipid,
OR=2.13, 95% CI 0.85–5.22) (Rogan et al., 1986). However, Bayley PDI scores at ages 6, 12, and
18 mo were not associated with breast milk DDE levels and there was a favourable association at
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age 24 mo (≥6 vs. <1 µg/g lipid, mean PDI difference = 8.0 ± 3.9 points) (Rogan & Gladen, 1991;
Rogan et al., 1986). A New York State birth cohort study involving mothers who consumed Lake
Ontario fish revealed no association between abnormal reflexes on postnatal day 2 and cord blood
DDE levels (Stewart et al., 2000). In a Faroe Islands birth cohort study, neurologic optimality test
scores at age 2 weeks were not associated with maternal serum (Spearman R=0.02) or breast milk
DDE levels (R=–0.01) (Steuerwald et al., 2000). There was an inverse association between cord
serum DDE and Bayley PDI scores among Spanish children age 13 mo (per doubling of cord serum
DDE, β=–3.83 ± 1.46, p < .05), independent of several covariates including breast-feeding
duration and cord serum PCB and HCB levels (Ribas-Fito et al., 2003). In a California birth cohort
study, PDI scores at ages 6 and 12 but not 24 mo were inversely associated with prenatal
serum DDE levels (change in PDI per 10-fold DDE increment: age 6 mos, β=–2.14, 95% CI –4.20
to –0.08; 12 mos, β=–2.14, 95% CI –4.83 to 0.56; 24 mos, β=0.59, 95% CI –1.58 to 2.77)
(Eskenazi et al., 2006).
Lactational exposure, DDT/DDE, inadequate evidence: The North Carolina cohort found no
association between cumulative lactational DDE exposure and PDI scores at ages 6, 12, 18, or 24
mo (Gladen et al., 1988; Rogan & Gladen, 1991).
Maternal exposure, HCB, inadequate evidence: The New York State birth cohort study
reported no association between abnormal reflexes on postnatal day 2 and cord blood HCB levels
(Stewart et al., 2000).
Maternal exposure, organophosphate insecticides, inadequate evidence: A birth cohort study in
California reported an association between abnormal reflexes among infants age less than 2 mo and
maternal urinary organophosphate insecticide metabolites (total dialkyl phosphates, β=0.53, 95%
CI 0.23–0.82); the association was stronger for infants with 4 or more versus 3 or fewer abnormal
reflexes (per 10-fold urinary dialkyl phosphate increment, OR=4.9, 95% CI 1.5–16.1) (Young
et al., 2005).
Summary Epidemiologic evidence for the role of environmental toxicants in motor deficits
among children age 0–2 includes: (a) sufficient evidence—high-level prenatal exposure to MeHg or
PCBs, PCDFs and related toxicants; (b) limited evidence—low-level prenatal or high-level childhood MeHg exposure; low-level prenatal PCB exposure.
Motor Function: Children Age 3 or Older
Lead Maternal exposure, inadequate evidence: In the Cincinnati birth cohort study, fine
motor function scores at age 15–17 yr were inversely associated with blood lead at age 6 but not
with prenatal or average childhood levels (Ris et al., 2004).
Childhood exposure, limited evidence: Fine motor function scores were inversely associated
with childhood blood lead levels in several birth cohorts (Needleman et al., 1990; Ris et al., 2004;
Stiles & Bellinger 1993; Stokes et al., 1998; Wasserman et al., 2000b) and cross-sectional studies
(Landrigan et al., 1975). A small cross-sectional study of children age 7–12 yr revealed no association between fine motor function and dentin lead levels (Winneke et al., 1983). In sum, there is
fairly consistent evidence for an association between childhood fine motor function deficits and low
to moderate lead exposure levels.
Methylmercury Low-level prenatal exposure, limited evidence: An expert panel review noted
that low-level prenatal MeHg exposure from prenatal fish consumption was associated with fine-motor
function deficits in two of the three large birth cohort studies and that evidence from experimental
animal studies showed motor deficits at low-dose prenatal MeHg exposure (National Academy of
Sciences, 2000c). In the Faroe Islands birth cohort study, fine motor function scores (finger tapping
speed and hand-eye coordination errors) at age 7 yr were unfavorably associated with prenatal hair
and/or cord blood mercury levels (change in score per doubling of maternal hair mercury, finger
tapping speed (nonpreferred hand), β=–0.32, p < .05; hand-eye coordination errors, β=0.018,
p < .05) (Dahl et al., 1996). Further analysis of this study revealed an inverse association between
finger tapping speed using the preferred hand (change per log cord blood mercury, β=–1.10,
p=.05) but not the other (Grandjean et al., 1997). In analyses limited to children of women with
prenatal hair mercury below 10 µg/g, the associations just described persisted with similar regression
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coefficients and p-values. Reanalysis of the Faroe Islands birth cohort study, excluding children with
strabismus or needing eye glasses and adjusting for visual contrast sensitivity, revealed inverse associations between finger tapping speed but not hand–eye coordination errors at age 7 yr and cord
blood mercury levels (Grandjean et al., 2001). Data from the Faroe Islands birth cohort study on
finger tapping speeds at age 7 yielded a benchmark dose lower limit of 4.3 µg/g for prenatal hair
mercury (using a log dose-response model and a 5% probability of an adverse response) (BudtzJorgensen et al., 2000). An analysis limited to Faroese children for whom two prenatal hair segment
mercury levels were available and consistent, revealed persistent unfavourable associations
between cord blood mercury levels and finger tapping speed and hand–eye coordination errors
(Grandjean et al., 2003). In the Faroe Islands cohort, there were inverse dose-response relationships
between maternal prenatal hair mercury levels and motor function test scores at age 14 (β=–9.37,
p=.009) (Debes et al., 2006).
In the Seychelles Islands study, prenatal hair mercury was inversely associated with grooved
pegboard time at age 9 (preferred hand, β=–1.07 ± 0.52, p=.045; nonpreferred hand, β=–1.39 ±
0.63, p=.03) (Davidson et al., 2000) but not with finger tapping speed (either hand) at age 9 yr
(Myers et al., 2003). In the New Zealand birth cohort study, there was an inverse association of
borderline statistical significance between McCarthy motor scale scores at age 6–7 and average
prenatal hair mercury levels (Crump et al., 1998). These authors estimated the bench mark dose
and its lower limit for average prenatal hair mercury versus abnormal McCarthy motor scale scores
to be, respectively, 21 and 9.8 µg/g. A cross-sectional study of children age 0–6 in French Guiana
reported no association between maternal hair mercury and finger tapping speed (β=1.27, p=.64)
or McCarthy leg coordination scores (β=–0.15, p=.62) (Cordier et al., 2002).
Low-level childhood MeHg exposure, inadequate evidence: After adjustment for cord blood
mercury, finger tapping scores at age 7 yr in the Faroe Islands birth cohort study were not associated
with current hair mercury (Grandjean et al., 1997, 1999). In a cross-sectional study of German
children age 5–7, finger tapping speed was not associated with current urinary mercury excretion
rates (µg/d) (β=0.32, p > .05) (Walkowiak et al., 1998). A larger cross-sectional study of children
age 7–12 in the Amazon Basin reported inverse associations between log current hair mercury
levels and finger tapping speed (preferred hand, β=–6.53, p < .001) and motor coordination and
dexterity score (β=–2.23, p=.005) (Grandjean et al., 1999).
Cadmium Low-level childhood exposure, inadequate evidence: A small cross-sectional study
of children age 5–16 in Maryland reported an inverse association between current hair cadmium
levels and fine motor function (finger tapping score vs. hair cadmium, partial R=–0.20, p=.05)
(Thatcher et al., 1982).
Manganese High-level childhood manganese exposure, sufficient evidence: Case reports indicate that children exposed to high levels of manganese from chronic parenteral nutrition developed
neurologic abnormalities including mild psychomotor retardation, frequent static and intention
tremor and magnetic resonance imaging (MRI) abnormalities (Fell et al., 1996; Komaki et al., 1999).
The latter report indicated that symptoms and MRI abnormalities disappeared after manganese
administration ceased (Komaki et al., 1999). Reviewers concluded that children exposed to high
manganese levels because of liver disease with impaired excretion or dependency on parenteral
nutrition developed neurotoxicity including loss of motor control of limbs and tremor (Agency for
Toxic Substances and Disease Registry, 2000b).
PCBs Background maternal PCB exposure, inadequate evidence: A Dutch birth cohort
study found no associations between neurologic optimality scores at age 42 mo and cord or
maternal levels of 4 noncoplanar PCBs (results stated without supporting data) (Lanting et al.,
1998b). Further follow-up of this cohort revealed a statistically nonsignificant association
between McCarthy motor subscale scores at age 6–7 and 3rd trimester maternal plasma levels of
4 noncoplanar PCBs (per natural log PCB increment, all children, β = –2.45 ± 1.45, p = .09,
adjusted for breastfeeding duration); this association was somewhat stronger among the formulafed subgroup (β = –3.92 ± 2.04, p = .06) (Vreugdenhil et al., 2002). In the Faroe Islands birth
cohort, fine motor function test scores at age 7 were not associated with log cord tissue levels of
3 noncoplanar PCBs, adjusted for cord blood mercury levels and other potential confounders
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(finger tapping score, preferred hand, β = –0.76, p = .30; hand-eye coordination errors, β = 0.04,
p = .26) (Grandjean et al., 2001).
Background lactational PCB exposure, inadequate evidence: A Dutch birth cohort study found
no associations between neurologic optimality scores at age 42 mo and breast milk PCB-TEQ levels
(results stated without supporting data) (Lanting et al., 1998b). Further follow-up of this cohort
revealed an association of borderline statistical significance between McCarthy motor subscale
scores at age 6–7 and 3rd trimester maternal plasma levels of 4 noncoplanar PCBs among formulafed but not among breastfed children (per natural log PCB increment, formula-fed, β=–3.92 ± 2.04,
p=.06; breastfed, β=–1.28 ± 1.84, p=.49, adjusted for breastfeeding duration) (Vreugdenhil
et al., 2002).
Pesticides Childhood exposure, organophosphate pesticides, inadequate evidence: A small
cross-sectional study of children age 2–11 living in homes where methyl parathion had been
illegally used for pest control in Mississippi and Ohio reported no association between lowest decile
motor function scores and exposure status (exposed vs. unexposed, Mississippi, OR=1.39, 95% CI
0.66–2.94; Ohio, OR=1.20, 95% CI 0.42–3.49) (Ruckart et al., 2004).
Childhood exposure, hexachlorobenzene, inadequate evidence: Follow-up of adults exposed
to hexachlorobenzene as children (after eating treated seed grain) revealed that about half had
weakness, myotonia and other neurological symptoms (Gocmen et al., 1989). There were no analyses of these traits in relation to exposure intensity.
Summary There was limited epidemiologic evidence for the role of environmental toxicants in
motor deficits among children age 3 yr or older including low-level childhood lead exposure and
low-level prenatal MeHg exposure.
Sensory Function
Maternal exposure, limited evidence: In the Cincinnati birth cohort, central auditory processing
ability at age 5 yr was inversely associated with neonatal and childhood blood lead levels (Dietrich
et al., 1992). Among neonates in Mexico City, wave III latencies (the time intervals between auditory stimuli and brainstem responses as detected on electroencephalographs, EEGs) were inversely
associated with prenatal and cord blood lead levels; wave III–V interpeak intervals were associated
with prenatal blood lead levels over the range 6–25 µg/dl (Rothenberg et al., 1994). Follow-up of this
cohort found a dose-response relationship between I-V and III-V conduction intervals at age 5 yr
and maternal 2nd trimester blood lead levels over the range 8–31 µg/dl (Rothenberg et al., 2000).
Auditory function Lead Childhood exposure, limited evidence: Evidence from two large
well-designed cross-sectional studies based on NHANES II and a similar survey of the U.S. Hispanic
population (Schwartz & Otto, 1987, 1991) and a Polish cohort study (Osman et al., 1999) suggests
that childhood lead exposure produces increased hearing thresholds. Findings included monotonic
dose-response relationships extending to blood lead levels below 10 µg/dl with no evidence of a
threshold (Schwartz, 1993). Reviewers concluded that epidemiologic and toxicologic evidence
support a relationship between increased hearing thresholds and low or moderate lead exposure
(Otto & Fox, 1993). Cross-sectional studies in Massachusetts and Denmark but not in Ecuador
observed unfavourable associations between central auditory processing indices and dentin or
current blood lead levels (Counter et al., 1997; Hansen et al., 1989; Needleman et al., 1979).
A review concluded that epidemiologic and toxicologic evidence indicate increased brainstem
auditory evoked potential latencies at low to moderate lead exposure levels (Otto & Fox, 1993).
The Mexico City cohort found inverse (i.e., apparently favorable) associations between I-V interpeak latencies at age 5 years and blood lead levels at ages 1 and 4 years; there were similar associations for III-V latencies (Rothenberg et al., 2000). The authors speculated that these results might
be explained by reduced auditory brainstem pathway length caused by lead exposure (similar to
the known inverse association between head circumference and lead exposure). Among Chinese
children age 1–6 yr, auditory evoked potential latencies were associated with blood lead over the
range 3–38 µg/dl and were significantly increased at levels above 10 µg/dl (Zou et al., 2003).
Methylmercury High-level prenatal exposure, sufficient evidence: During the 1972 Iraq
MeHg poisoning episode, many of the most affected infants had severely impaired hearing (Amin-Zaki
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et al., 1974b, 1979), a finding supported by several reviewers (Clarkson, 2002; National Academy of
Sciences, 2000c; United Nations Environment Programme, 2002).
Low-level prenatal exposure, limited evidence: In the Faroe Islands birth cohort, hearing thresholds at age 14 yr were not associated with cord blood mercury except for an inverse association for
right ear threshold at 4 kHz (p-trend<.01) (Murata et al., 2004). However, brain stem auditoryevoked potential latencies and interpeak intervals at age 7 and 14 were associated with prenatal
hair and cord blood mercury levels (e.g., age 7, change in potential III latency at 40 Hz per log cord
blood mercury increment, β=0.108, p=.02) (Murata et al., 1999, 2004). Combined analysis of the
Faroe Islands and Madeira studies yielded a benchmark dose for prenatal hair mercury of 9.5 µg/g
for a doubling of a 5% prevalence of abnormal auditory-evoked potential latencies at 40 Hz, with
similar results at 20 Hz (Murata et al., 2002). A reviewer concluded that prenatal or childhood
MeHg exposure affects auditory systems at the cortical level (Newland 2002).
Childhood (United Nations Environment Programme, 2002) exposure, inadequate evidence:
A small cross-sectional study of Ecuadorian children age 3–15 yr reported an inverse (i.e., favourable)
association between current blood mercury levels and hearing threshold at 3 kHz in the right ear (Pearson
r=.55, p=.01) but not the left ear (Counter et al., 1998). This study also found associations between
brainstem auditory-evoked potential wave V latencies and I-V interpeak intervals at age 3–15 and current child blood mercury levels (e.g., wave V latency vs. current blood mercury, Spearman R=0.38,
p=.03) but there was no adjustment for potential confounders (Counter, 2003). In the Faroe Islands
birth cohort study, brain stem auditory-evoked potential latencies and interpeak intervals at age 7 and
14 were not associated with current child hair mercury levels (e.g., age 7, change in potential III latency
at 40 Hz per log child hair mercury increment, β=0.002, p=.96) (Murata et al., 1999, 2004).
PCBs High-level maternal PCB/PCDF exposure, limited evidence: Yucheng children age 7–12
prenatally exposed to high levels of PCBs, PCDFs and related toxicants had increased auditoryevoked potential latencies compared to those of unexposed children (mean auditory-evoked
potential latencies, 356.0 ± 36.9 ms vs. 329.3 ± 25.5 ms, t=3.09, p=.003) (Chen & Hsu, 1994).
Background maternal PCB exposure, limited evidence: In the Faroe Islands birth cohort, there
was no association between auditory-evoked potential latencies at age 7 yr and umbilical cord
tissue levels of 3 noncoplanar PCBs (change in latency per log increment cord tissue levels of
3 noncoplanar PCBs, 20 Hz III β=0.02 ms, p=.68; 20 Hz V β=0.07 ms, p=.08; 40 Hz III
β=0.05 ms, p=.34; 40 Hz V β=0.03 ms, p=.54) (Grandjean et al., 2001). Among children age
9 in the Rotterdam component of the Dutch birth cohort study with 4th quartile maternal plasma
levels of 4 noncoplanar PCBs, mean auditory-evoked potential latencies were greater than those of
children with 1st quartile maternal plasma levels (adjusted mean latency difference, P300Fz 14.3 ±
9.5 ms, p=.14; P300Cz 25.6 ± 9.6 ms, p=.01; P300Pz 22.0 ± 9.4 ms, p=.02) (Vreugdenhil et al.,
2004b). In the U.S. Collaborative Perinatal Project cohort, sensorineural hearing loss at age 8 was
not associated with maternal serum PCB levels over the range <1.25 to ≥5 µg/L (p-trend=.76)
(Longnecker et al., 2004).
Lactational PCB exposure, inadequate evidence: Among children age 9 in the Rotterdam
component of the Dutch birth cohort study, those who were breast-fed for at least 17 wk had
reduced auditory-evoked potential latencies compared to those who were formula-fed (adjusted
mean latency difference, P300Fz –19.8 ± 10.8 ms, p=.07; P300Cz–20.2 ± 10.9 ms, p=.07;
P300Pz –22.5 ± 10.6 ms, p=.04) (Vreugdenhil et al., 2004b).
Summary There is sufficient epidemiologic evidence that high-level prenatal or childhood
MeHg exposure produces auditory and visual function deficits. Limited evidence supports associations
between auditory function deficits and low-level prenatal or childhood lead exposure, low-level
prenatal MeHg exposure or prenatal exposure to PCBs, PCDFs, and related toxicants can impair
auditory function in children.
Visual function Lead Childhood exposure, inadequate evidence: A German cross-sectional
study of children age 6 reported no association between visual evoked potential (VEP) latencies and
current blood lead levels over the range 1.3–19.0 µg/dl; among 6 relationships explored, only one
was related to tooth lead concentrations (F2/N2 was inversely associated with tooth lead, p=.06)
(Winneke et al., 1994).
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Methylmercury Maternal exposure, inadequate evidence: VEP latencies at age 7 were not
associated with cord blood mercury in the Faroe Islands birth cohort study (Grandjean et al., 1997).
In a cross-sectional study of children age 6–7 in Madeira, N145, N75-N145 and P100-N145 VEP
latencies were associated with maternal hair mercury levels (Murata et al., 1999). A pooled study of
children age 7–12 in Greenland, Madeira and the Faroes observed no association between VEP
latencies and prenatal hair mercury levels (Weihe et al., 2002). A Quebec birth cohort study
revealed associations between P100 VEP latencies and cord blood mercury levels (Saint-Amour
et al., 2006). It should be noted, however, that high-level prenatal or early childhood MeHg exposure in Iraq caused blindness in several children (Amin-Zaki et al., 1974a, 1974b).
Childhood exposure, inadequate evidence: A German cross-sectional study revealed no association between VEP latencies at age 7 and current urinary mercury levels (Altmann et al., 1998).
However, 4 of 10 contrast sensitivity values tested, indicators of visual cortical function, were
inversely associated with current urinary mercury levels. In the Madeira cross-sectional study, only
P100-N145 VEP latencies were associated with current child hair mercury levels (Murata et al.,
1999). The Quebec birth cohort study reported inverse associations between N75 and P100 VEP
latencies and current child blood mercury levels, independent of potential confounders including
maternal alcohol or marijuana use and current child blood PCB levels (Saint-Amour et al., 2006).
Although inverse associations between VEP latencies and child blood mercury levels might be
attributed to a favorable effect of a high-fish diet, both increased and reduced latencies might
reflect disruption of visual processing (Saint-Amour et al., 2006).
PCBs High-level maternal PCB/PCDF exposure, limited evidence: Yucheng children age 7–12 yr
prenatally exposed to high levels of PCBs, PCDFs, and related toxicants had visual-evoked
potential latencies similar to those of unexposed children (mean visual-evoked potential latencies,
148.7 ± 15.0 ms vs. 153.1 ± 19.2 ms, t=0.94, p=.35) (Chen & Hsu, 1994).
Background maternal PCB exposure, limited evidence: In the Faroe Islands birth cohort, there
was no association between visual-evoked potential latencies at age 7 yr and umbilical cord tissue
levels of 3 noncoplanar PCBs (change in latency per log increment cord tissue levels of 3 noncoplanar
PCBs, N75 β=0.74 ms, p=.26; P100 β=1.44 ms, p=.22; N145 β=3.35 ms, p=.11) (Grandjean
et al., 2001). A Quebec birth cohort study reported no association between VEP latencies and cord
plasma PCB-153 levels (result stated without supporting data) (Saint-Amour et al., 2006).
Lactational PCB exposure, inadequate evidence: The Quebec birth cohort study observed associations between P100 and N150 VEP latencies and current child blood PCB levels, independent of
potential confounders including maternal alcohol or marijuana use and current child blood mercury
levels (change in latency per natural log increment of cord plasma PCB-153, 12% contrast, N75
β=–0.44 ms, p > .05; P100 β=3.22 ms, p < .05; N150 β=5.58 ms, p < .05) (Saint-Amour et al.,
2006). Current plasma PCB levels in young children with a history of breastfeeding mainly reflect
lactational exposure.
Respiratory Diseases
Lower respiratory tract diseases (including asthma, bronchiolitis/bronchitis and pneumonia) are
the leading cause of hospitalization among U.S. children age 1–9 yr (U.S. Department of Health
and Human Services, 2003). Over 4 million U.S. children have asthma and asthma prevalence, and
morbidity and mortality rates increased substantially among children and young adults during
recent decades in Canada and other developed countries (Commission for Environmental Cooperation, 2006; Millar & Hill, 1998). Upper respiratory tract infections are major causes of illness leading
to physician visits. There were about 8 million physician visits for middle ear infections among
children age 0–12 yr in United States during 2004 (Hing et al., 2006).
The level of epidemiologic evidence for associations between major childhood respiratory
diseases and environmental chemical contaminants is summarized in Table 5. Sufficient evidence
was found for associations between: (1) new onset asthma and ETS, (2) asthma severity and ETS and
outdoor air pollution, and (3) lung and middle ear infections and ETS. Children in household with
relatively low socioeconomic status are at increased risk of exposure to the above environmental
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TABLE 5. Role of Environmental Toxicants in Common Childhood Respiratory Diseases
Toxicant
Exposure
PCBs
Prenatal
DDT/DDE
Other organochlorine
pesticides
Unspecified pesticides
Environmental
tobacco smoke
Indoor air NO2
Indoor chlorinated
swimming pools
Indoor house dust,
di(2-ethylhexyl)
phthalate
Indoor air, VOCs
Indoor air, formaldehyde
Indoor biomass smoke
Outdoor air pollution
New-onset
childhood asthma
Childhood
lung infections
Childhood middle
ear Infections
High-level—I
Low-level—I
L
I
High-level—L
Low-level—I
I
I
I
L
I
I
I
Lactational
Prenatal
Lactational or
childhood
Lactational or
childhood
Prenatal
Childhood
Childhood
I
S
Childhood
Childhood
I
L
Childhood
L
Childhood
Childhood
Childhood
Childhood
I
L
Childhood
asthma severity
I
L
S
S
S
L
S
S
L
L
L
risk factors for respiratory diseases (Almqvist et al., 2005; Perera et al., 2002; Shapiro & Stout,
2002). Relatively little is known about critical exposure time windows for the influence of environmental agents on the risk of childhood respiratory diseases. However, prenatal and early-life exposures may be especially important as lung development continues from the prenatal period through
adolescence (Dietert et al., 2000; Pinkerton & Joad, 2000).
New-Onset Asthma
Pesticides Maternal exposure, organochlorine pesticides, inadequate evidence: A prospective
study of Spanish children revealed an association between wheezing at age 4 yr and cord serum
DDE levels (4th vs. 1st quartile serum DDE, RR=2.63,95% CI 1.19–4.69) (Sunyer et al., 2005).
There were also associations between cord serum DDE and persistent wheezing (wheezing at age
4 yr and during previous year) (RR=1.26, 95% CI 1.04–1.54) and for physician-diagnosed asthma
(RR=1.46, 95% CI 0.92–2.32). No associations were observed between wheezing and cord serum
HCB quartiles.
Maternal exposure, unspecified pesticides, inadequate evidence: The Ontario Farm Family
Study found no association between physician-diagnosed childhood asthma and prenatal farm pesticide use (any vs. none, OR=1.00, 95% CI 0.71–1.40) or any specific pesticide (Weselak et al., 2007).
Paternal exposure, chlorophenoxy herbicides, inadequate evidence: Among U.S. Vietnam
veterans, the prevalence of parent-reported asthma up to age 5 yr was higher than that among
children of non-Vietnam veterans (crude OR=1.1, 95% CI 1.0–1.4) (Centers for Disease Control,
1989). Potential limitations include differential recall bias as well as a lack of data on maternal and
other environmental exposures.
Childhood exposure, DDT/DDE or HCB, inadequate evidence: In a retrospective cohort study
of German children age 7–10, physician-diagnosed asthma was associated with blood DDE (≥0.3
vs. <0.3 µg/L, OR= 3.71, 95% CI 1.10–12.6) but not with HCB levels (Karmaus et al., 2001). Further
investigation showed that an inverse association between asthma and a history of breast feeding was
strongest among children with blood DDE levels below 0.29 µg/L (Karmaus et al., 2003).
Childhood exposure, unspecified pesticides, inadequate evidence: A retrospective cohort study
of Lebanese children age 5–16 yr indicated associations between physician-diagnosed asthma and
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parent-reported residential proximity to agricultural pesticide use (OR=2.10, 95% CI 1.01–4.42),
residential pesticide use (home or garden use, yes vs. no, OR=1.99, 95% CI 1.00–3.99) or parental
occupational pesticide exposure (OR=4.61, 95% CI 2.06–10.3) (all ORs adjusted for ETS and other
risk factors) (Salameh et al., 2003). The questionnaire used in this study did not permit assessment
of age at diagnosis of asthma or pesticide exposure before the onset of asthma. Also, there was
potential for recall bias in parental completion of the questionnaire.
A case-control study nested within the Children’s Health Study revealed that onset of asthma by
age 3 and persistent to at least age 5 was associated with exposure beginning before age 1 to residential or agricultural use of any pesticide (OR=3.58, 95% CI 1.59–8.06) and herbicides in particular (OR=10.1, 95% CI 2.46–41.3) (Salam et al., 2004). The latter study found no link with
pesticide exposure beginning after age 1 yr.
Environmental tobacco smoke Early childhood exposure, sufficient evidence: Several recent
reviews concluded that incident childhood asthma is associated with childhood ETS exposure
(California Environmental Protection Agency, 2005; DiFranza et al., 2004; Jaakkola & Jaakkola,
2002; World Health Organization 1999). The California expert panel noted that the evidence is
particularly strong for young children and those whose mothers smoked during pregnancy (California
Environmental Protection Agency, 2005). A meta-analysis of 4 cohort studies indicated an association
between incident asthma before age 8 yr and maternal smoking (summary OR=1.31, 95% CI
1.22–1.41) (Strachan & Cook, 1998b). A UK birth cohort study reported that incident asthma/
wheeze by age 3 yr was associated with postnatal maternal smoking, independent of prenatal
smoking and other potential confounders (OR=1.93, 95% CI 1.10–3.38) (Murray et al., 2004).
A recent case-control study nested in the Children’s Health Study in California reported a statistically nonsignificant increased risk of asthma before age 5 among children in homes with 2 or more
smokers (OR=1.3, 95% CI 0.8–2.1, not adjusted for prenatal smoking) (Li et al., 2005).
Indoor gases Childhood exposure, inadequate evidence: At relatively high levels, nitrogen
dioxide (NO2) can precipitate asthma episodes but there is inadequate evidence to assess its role in
asthma development (National Academy of Sciences, 2000a). See also later discussion of outdoor
air pollution, since NO2 is a major ambient air pollutant and respiratory toxicant.
Other indoor contaminants Childhood exposure, indoor chlorinated swimming pools, limited
evidence: Ecologic studies in Belgium and Europe revealed associations between the prevalence of
exercise-induced asthma among children in 15 schools and cumulated frequency of indoor
swimming pool attendance (Bernard et al., 2003) and between the prevalence of self-reported
childhood asthma in 66 European cities and the per capita prevalence of indoor chlorinated swimming pools (Nickmilder & Bernard, 2007). Further investigation in Belgium showed a dose-response
relationship between childhood asthma and cumulated indoor chlorinated pool attendance among
the subset with elevated serum immunoglobulin (Ig) E levels, especially among those exposed
before age 7, and independent of other risk factors including ETS (physician-diagnosed asthma,
serum IgE >100 kIU/L, per 100 h cumulative exposure, OR=1.57, 95% CI 1.07–2.30) (Bernard
et al., 2006). Chlorinated swimming pools contain DBPs including volatile compounds that are
inhaled by swimmers; nitrogen trichloride is the most volatile and concentrated DBP in the air of
indoor pools and is a powerful respiratory tract irritant (Hery et al., 1995).
Childhood exposure, various VOCs, inadequate evidence: In an Australian case-control study,
asthma among preschool children was associated with residential indoor air concentrations of
several VOCs, particularly benzene (per 10 mg/m3 increment, OR=2.92, 95% CI 2.25–3.80),
ethylbenzene (OR=2.54, 95% CI 1.16–5.57) and toluene (OR=1.84, 95% CI 1.41–2.41)
(Rumchev et al., 2004).
Childhood exposure, formaldehyde, limited evidence: A recent review found limited evidence
for an association between childhood asthma and indoor levels of formaldehyde or the presence of
formaldehyde-emitting materials (Mendell, 2007). Because such materials also emit other VOCs
including aldehydes, it is uncertain that formaldehyde per se explains the latter association.
Childhood exposure, house dust di(2-ethylhexyl) phthalate (DEHP), limited evidence: A nested
case-control study in Sweden revealed a dose-response relationship between childhood asthma
and house dust DEHP concentrations (ORs for increasing quartiles relative to 1st quartile were 1.56,
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95% CI 0.70–3.46, 2.05, 95% CI 0.94–4.47 and 2.93, 95% CI 1.36–6.34) (Bornehag et al., 2004).
A similar study in Bulgaria showed a dose-response relationship between wheezing among children
age 2–7 and house dust DEHP levels (4th vs. 1st quartile dust DEHP levels, OR = 3.7, 95% CI 1.4–
9.9, p-trend = 0.02) (Kolarik et al., 2007).
Outdoor air pollution Childhood exposure, limited evidence: A reviewer found overwhelming
evidence that ambient air pollution from traffic increases the risk of developing asthma during childhood (Schwartz, 2004). A WHO expert panel reviewed studies published up to 2004 and concluded that there is little evidence for a causal association between childhood asthma prevalence/
incidence and air pollution in general, but there is suggestive evidence for a causal association
between the prevalence/incidence of childhood asthma symptoms and living in close proximity to
traffic (Binkova et al., 2005). Among reviewed studies, cohort studies in Japan, Germany and
California found associations, including dose-response relationships, between incident asthma in
young children and ambient air levels of one or more pollutants including ozone, NO2 and PM
(Gehring et al., 2002; McConnell et al., 2002; Shima & Adachi, 2000; Shima et al., 2002). Among
high-ozone communities in the California cohort study, there was a dose-response relationship
between incident asthma and number of team sports engaged in by subjects (OR=3.3, 95% CI
1.9–5.8) with a somewhat stronger association among children without a history of wheeze at
baseline (OR=4.4, 95% CI 2.1–9.3) (McConnell et al., 2002). The California study also observed
associations between incident asthma and team sports participation in both low- and high-PM10
communities (low PM10, OR=1.7, 95% CI 0.9–3.2; high PM10, OR=2.0, 95% CI 1.1–3.6).
A cohort study in the Netherlands found weak associations between incident asthma before age 2
yr and ambient air PM (per interquartile range, OR=1.12, 95% CI 0.84–1.50) and NO2 (per interquartile range, OR=1.18, 95% CI 0.93–1.51) levels (Brauer et al., 2002). The Japanese cohort
study reported a moderately strong association between incident asthma in young children and
residential proximity to main roads (<50 m vs. rural residence, girls, OR=4.06, 95% CI 0.91–18.1;
boys, OR=3.75, 95% CI 1.00–14.1) (Shima et al., 2003). A small case-control study in France
revealed an association between childhood incident asthma and cumulative traffic density exposure
before age 3 (3rd vs. 1st tertile, OR=2.28, 95% CI 1.14–4.56) (Zmirou et al., 2004). In a new
California cohort study, incident asthma by age 5–7 was associated with residential proximity to a
main road (<75 vs. >300 m, all children, OR=1.29, 95% CI 1.01–1.66; long-term residents,
OR=1.46, 95% CI 0.98–2.17) (McConnell et al., 2006). This relationship was limited to girls
(OR=2.51, 95% CI 1.39–4.54; boys, OR=0.94, 95% CI 0.54–1.64).
Summary Epidemiologic evidence for the role of environmental toxicants in new-onset
childhood asthma includes: (a) sufficient evidence—childhood ETS exposure; (b) limited
evidence—childhood indoor chlorinated swimming pool use, house dust di(2-ethylhexyl)phthalate,
indoor air formaldehyde and residential outdoor air pollution (mainly traffic-related). Although
limited, the evidence linking ambient air pollution to incident asthma is quite strong and suggestive
of a causal relationship.
Asthma Severity
Pesticides Childhood exposure, organophosphate and pyrethroid insecticides, inadequate
evidence: In New York City, daily asthma emergency department visits in the South Bronx did not
change significantly during the 1999 West Nile Virus mosquito control campaign in which
malathion (an organophosphate insecticide) was aerially sprayed and resmethin (a pyrethroid insecticide) was sprayed from trucks for 4 d (O’sullivan et al., 2005). This study did not measure individual exposure or control for other environmental factors. Another study assessed emergency room
visits to all 11 public hospitals in New York City during a 14-mo period including the 2000 West
Nile Virus campaign in which pyrethroid pesticides were sprayed from truck from July to September
2000 (Karpati et al., 2004). ZIP codes were used to assign spraying versus nonspraying exposure
and residence location at the time of emergency department presentation. No difference in emergency department presentation for asthma was observed among children age <15 yr during the 3 d
before and after spraying (RR=0.78, 95% CI 0.61–1.01). No association was observed among ZIP
codes considered that received the heaviest spraying. Although the New York City studies evaluated
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the effects of low levels of pesticide exposure among the general population, these studies only
considered asthma symptoms severe enough to require presentation to a hospital. It is possible that
more subtle respiratory effects may have been experienced. Additionally, an ecologic indicator of
exposure was used; therefore the precise level of personal exposure is also unknown.
Childhood exposure, unspecified pesticides, inadequate evidence: A cross-sectional study in
New Jersey revealed a slightly elevated risk of peak flow test failure among children in grades 2–3
exposed to indoor residential use of pesticide sprays or powders (OR=1.12, 95% CI 0.88–1.41)
and those in grades 3–4 (OR=1.30, 95% CI 0.97–1.74) but not those in grade 5 (OR=0.86, 95%
CI 0.50–1.49) (Schneider et al., 2004). This study experienced declining response rates and
changes in protocol and staff over the 4-year study period and lacked information on other potentially related environmental influences.
Environmental tobacco smoke Childhood exposure, sufficient evidence: Several reviews
concluded that the frequency and severity of symptoms among asthmatic children is associated
with postnatal ETS exposure (California Environmental Protection Agency, 2005; DiFranza et al.,
2004; Jaakkola & Jaakkola, 2002; World Health Organization, 1999). Three expert panel reviews
concluded that this association is causal (California Environmental Protection Agency, 2005; U.S.
Department of Health and Human Services, 2006; World Health Organization 1999).
Other indoor air pollutants Childhood exposure, biomass smoke, limited evidence: A WHO
review panel found limited evidence for an association between increased childhood asthma severity
and exposure to biomass smoke in developing countries (Bruce et al. 2002).
Outdoor air pollution Childhood exposure, sufficient evidence: Reviewers noted evidence
that summertime ozone elevations increase childhood asthma severity as indicated by reduced lung
function, more symptoms and more episodes requiring medical attention (including ER visits) (Suh
et al., 2000). A WHO expert panel reviewed studies published up to 2004 and concluded that
there is sufficient evidence for a causal association between childhood asthma exacerbation and air
pollution, mainly due to PM and ozone (Binkova et al., 2005). In an ecologic study in Vancouver,
Canada, asthma hospitalizations among girls age 6–12 yr in low household income families (based on
small area census data) were associated with 4-d average SO2 levels (OR=1.18, 95% CI 1.02–1.36);
there was no association for girls in higher household income families (OR=0.99, 95% CI 0.85–1.15)
or for boys in either group (Lin et al., 2004). In a new California cohort study, current wheeze
among children age 5–7 was associated with residential proximity to a main road (<75 vs. >300 m,
all children, OR=1.40, 95% CI 1.09–1.78; long-term residents, OR=1.67, 95% CI 1.14–2.43)
(McConnell et al., 2006).
Summary There is sufficient epidemiologic evidence for causal associations between childhood asthma severity and postnatal exposure to ETS or outdoor air pollution and limited evidence
for an association with indoor exposure to biomass smoke (mainly in developing countries).
Lung Infections
PCBs High-level maternal PCB/PCDF exposure, limited evidence: A retrospective cohort
study of Yucheng children age 6 yr revealed a substantially increased prevalence of a history of
bronchitis or pneumonia by age 6 mo (exposed vs. comparison children, OR=7.02, 95% CI 2.74–
21.0; calculated from data in paper) (Rogan et al., 1988). A WHO expert group concluded that
high-level maternal PCB/PCDF exposure was associated with increased risk of bronchitis among
infants (Brouwer et al., 1998).
Background maternal PCB exposure, inadequate evidence: Reviews of epidemiologic studies
published up to the mid–1990s found inadequate evidence for an association between risk of
infections during infancy and background maternal PCB exposure levels (Longnecker et al., 1997;
Tryphonas, 1998). A multicentre Dutch birth cohort study reported no association between a
history of pneumonia by age 42 mo and maternal plasma PCB (per natural log increment of maternal plasma PCB, OR=0.41, 95% CI 0.10–1.63, adjusted for breastfeeding history and duration)
(Weisglas-Kuperus et al., 2000). Among Inuit women and infants in northern Quebec, a history of
clinically confirmed lower respiratory tract infection by age 12 mo was not associated with maternal
plasma PCB–153 levels (4th vs. 1st quartile plasma PCB–153, OR=1.03, 95% CI 0.72–1.48,
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p-trend=.36) (Dallaire et al., 2004). In a similar study, a clinically confirmed history of lower respiratory tract infection by age 5 yr was associated with cord plasma PCB–153 (4th vs. 1st quartile,
OR=1.44, 95% CI 1.20–1.72) with evidence of a dose-response relationship (per log cord plasma
PCB–153 increment, OR=1.14, 95% CI 1.04–1.24) (Dallaire et al., 2006).
Lactational PCB exposure, inadequate evidence: The Dutch birth cohort study revealed an
inverse association between a history of pneumonia by age 42 mo and current child plasma PCB, an
index of lactational exposure (per natural log increment of child plasma PCB (µg/L), OR=0.0.01, 95%
CI 0.01–0.37, adjusted for breastfeeding vs. formula feeding and duration of breastfeeding) (WeisglasKuperus et al., 2000). In a German cross-sectional study of children age 7–10 yr, a history of physician-diagnosed pneumonia was not associated with whole blood PCB levels stratified by whole blood
DDE levels (for DDE <0.3 µg/L, PCB >0.48 vs. ≤0 0.48 µg/L, OR=1.24, 95% CI 0.53–2.92; for DDE
≥0.3 µg/L, PCB >0.48 vs. ≤0 0.48 µg/L, OR=0.68, 95% CI 0.26–1.76) (Karmaus et al., 2001).
Pesticides Maternal exposure, DDT/DDE, inadequate evidence: A birth cohort study in northern
Quebec revealed no association between a history of lower respiratory tract infection before age 6
mo (confirmed by medical chart review) and prenatal plasma DDE (4th vs. 1st quartile, OR=0.96,
95% CI 0.55–1.66, p-trend=0.89) (Dallaire et al., 2004).
Maternal exposure, other and unspecified pesticides, inadequate evidence: A retrospective
cohort study of Ontario farm families revealed no association between a history of persistent cough
or bronchitis and parent-reported prenatal farm use of any pesticide (OR=1.21, 95% CI 0.77–
1.90) or specific pesticides (Weselak et al., 2006). This study relied on parent recall of prenatal
events often 10 yr or more before interview, raising the possibility of considerable exposure misclassification and attenuation of risk estimates.
Childhood exposure, DDT/DDE or HCB, inadequate evidence: A retrospective cohort study of
German children age 7–10 revealed no association between current whole blood DDE levels and a
history of physician-diagnosed whooping cough (≥0.3 vs. <0.3 µg/L, OR=0.61, 95% CI 0.38–0.99)
or pneumonia (OR=0.79, 95% CI 0.45–1.40) (Karmaus et al., 2001). This study reported no association between these outcomes and blood HCB levels.
Environmental tobacco smoke Childhood exposure, sufficient evidence: A meta-analysis of 13
studies concluded that serious lower respiratory infection among young children age 0–6 was associated with ETS exposure (pooled OR=1.57 95% CI = 1.28–1.91) (Li et al., 1999). Four recent
reviews concluded that childhood ETS exposure is an important cause of lower respiratory infections during early childhood (California Environmental Protection Agency, 2005; Jaakkola &
Jaakkola, 2002; U.S. Department of Health and Human Services, 2006; World Health Organization, 1999; California Environmental Protection Agency, 2005; Jaakkola & Jaakkola, 2002).
Other indoor air pollutants Childhood exposure, biomass smoke, sufficient evidence: A WHO
review panel found sufficient evidence for an association between childhood pneumonia and
exposure to biomass smoke in developing countries (Bruce et al. 2002). In developed countries,
lower respiratory tract infections in children have been linked to indoor air PM10 concentrations
related to use of wood-burning stoves (PM10 ≥ 65 vs. <65 µg/m3, crude OR=7.0, 95% CI 0.9–
56.9); the authors stated that the OR changed little after adjustment for other factors including ETS
(Robin et al., 1996).
Outdoor air pollution Childhood exposure, limited evidence: Reviewers concluded that
evidence from time-series and panel studies published up to 2000 supports an association between
acute respiratory infection and outdoor air pollutants, especially PM2.5 and ozone, but it was not
clear whether ambient air pollutants affected respiratory infection incidence, severity or both
(Romieu et al., 2002). Another reviewer noted consistent evidence for an association between
childhood bronchitis and long-term exposure to PM and some evidence for reduced risk after
abatement of such exposure (Schwartz 2004). Recently, a WHO expert panel concluded that air
pollutants including PM10, PM2.5, NO2, SO2 and ozone are associated with upper and lower respiratory symptoms in children, much of which are likely related to infections (Binkova et al., 2005). In a
large California retrospective cohort study, postneonatal respiratory deaths were associated with
average PM2.5 levels within 8 km of the maternal residence (per 10 µg/m3 increment, OR=2.13,
95% CI 1.12–4.05) (Woodruff et al., 2006).
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449
Summary Epidemiologic evidence for the role of environmental toxicants in childhood lung
infections includes: (a) sufficient evidence—childhood ETS and biomass smoke exposure; (b)
limited evidence—high-level prenatal exposure to PCBs, PCDFs and related toxicants; childhood
exposure to outdoor air pollution, mainly from traffic.
Middle Ear Infections
PCBs High-level maternal PCB/PCDF exposure, inadequate evidence: Yucheng children
age 8–16 had a statistically nonsignificant increased risk of parent-reported ear infections during the
6 mo before their 1995 examination (mean frequency, exposed vs. unexposed, 0.67 ± 3.88 vs.
0.03 ± 0.22, p>.05) (Yu et al., 1998). However, Yucheng children had an increased risk of
otolaryngologically confirmed chronic otitis media compared to unexposed children (OR=3.23,
95% CI 1.70–6.23) (Chao et al., 1997). This study revealed associations between chronic otitis
media and current serum PCDF levels (cases vs. controls, proportion with serum PCDF ≥400 ng/kg
lipid, 5/15 vs. 0/15, p=.04) but not current serum PCB levels (cases vs. controls, proportion with
serum PCB ≥4 mg/kg lipid, 5/15 vs. 4/15, p=.86).
Background maternal PCB exposure, inadequate evidence: In the Dutch birth cohort, a history of at least 6 ear infections by age 42 mo was weakly associated with maternal plasma levels
of 4 noncoplanar PCBs (per natural log increment, OR = 1.37, 95% CI 0.87–2.17, adjusted for
breastfeeding history and duration) (Weisglas-Kuperus et al., 2000). Further follow-up of this
cohort showed that a history of recurrent otitis media at age 3–7 yr was not associated with
maternal plasma PCB (per natural log increment of plasma PCB (µg/L), OR = 0.98, 95% CI 0.53–
1.80) (Weisglas-Kuperus et al., 2004). In a cohort of Inuit infants, a history of otitis media by age
1 yr was not associated with prenatal plasma PCB-153 levels (4th vs. 1st quartile, OR = 0.97, 95%
CI 0.73–1.28, p-trend = .89) (Dallaire et al., 2004). In a similar study, a clinically confirmed history of acute otitis media by age 5 yr was associated with cord plasma PCB–153 (4th vs. 1st
quartile, OR = 1.37, 95% CI 1.20–1.55) with evidence of a dose-response relationship (per log
cord plasma PCB–153 increment (µg/g lipid), OR = 1.12, 95% CI 1.05–1.20) (Dallaire et al.,
2006).
Lactational PCB exposure, limited evidence: In the North Carolina birth cohort study, the
prevalence of a history of ear infections before age 1 was lower (47%) among those in the highest
cumulative lactational PCB dose category than that for formula-fed infants (58%) or those in the
lowest cumulative lactational PCB dose category (50%); the authors did not include a statistical
measure of the trend of prevalence versus PCB dose (Rogan et al., 1987). Among breastfed Quebec
Inuit infants in the highest tertile of breast milk PCB levels, there were statistically nonsignificant
elevated risks of otitis media (at least once before age 12 mo, 3rd vs. 1st tertile, OR=1.28, 95% CI
0.92–1.77; at least 3 occurrences before age 1, OR=1.65, 95% CI 0.49–5.57) (Dewailly et al.,
2000). In the Dutch birth cohort, a history of 1 or more ear infections by age 42 mo was not associated with current child plasma PCB levels (per natural log plasma PCB increment, OR=1.27, 95%
CI 0.61–2.64, adjusted for breastfeeding history and duration) but there was an association for the
occurrence of at least 6 ear infections by age 42 mo (per natural log plasma PCB increment,
OR=3.06, 95% CI 1.17–7.98) (Weisglas-Kuperus et al., 2000). This study also showed that recurrent otitis media was inversely associated with short lactational exposure (6–16 vs. ≥16 wk,
OR=0.12, 95% CI 0.01–1.07) (Weisglas-Kuperus et al., 2000). Plasma PCB levels at age 3–4 yr are
4–5 times higher among breast-fed compared to formula-fed children; thus plasma PCB levels in
young children are a proxy for lactational exposure (Lanting et al., 1998a). Further follow-up of the
Dutch birth cohort revealed that a history of recurrent otitis media at age 3–7 yr was associated with
cumulative lactational PCB dose based on breast milk PCB levels times weeks of lactation (per
natural log increment of lactational PCB dose (µg-week/g lipid), OR=1.19, 95% CI 1.01–1.41)
(Weisglas-Kuperus et al., 2004). Among German children age 7–10, a history of physician-diagnosed
ear infections was associated with whole blood PCB levels among the subgroup of children with
above-median whole blood DDE levels (whole blood PCB >0.48 vs. ≤0 0.48 µg/L, OR=3.70, 95%
CI 1.64–8.34) but not among children with below-median DDE levels, suggesting a possible interaction between these 2 exposures (Karmaus et al., 2001).
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D. T. WIGLE ET AL.
Pesticides Lactational exposure, DDT/DDE, limited evidence: Among breast-fed infants in
northern Quebec, at least 1 occurrence of otitis media before age 1 was associated with breast
milk levels of DDE (3rd vs. 1st tertile DDE, OR=1.52, 95% CI 1.05–2.22) (Dewailly et al., 2000).
This study observed stronger associations for a history of at least 3 occurrences of otitis media
before age 1 yr and breast milk levels of DDE (3rd vs. 1st tertile DDE, OR=3.48, 95% CI 0.86–
14.1). In a cross-sectional study of German children age 7–10 yr, a history of physician-diagnosed
otitis media was inversely associated with blood DDE (≥0.3 vs. <0.3 µg/L, OR=0.50, 95% CI
0.31–0.79) (Karmaus et al., 2001). The latter study did observe an association between otitis
media and above-median blood levels of both DDE and HCB (OR=2.38, 95% CI 1.08–5.25).
A birth cohort study in northern Quebec reported a dose-response relationship of borderline
statistical significance between a history of otitis media before age 6 mo (confirmed by medical
chart review) and prenatal plasma DDE (4th vs. 1st quartile, OR=1.55, 95% CI 0.90–2.68,
p-trend=0.07) (Dallaire et al., 2004).
Lactational exposure, other organochlorine pesticides, inadequate evidence: Among breastfed infants in northern Quebec, at least 1 occurrence of otitis media before age 1 yr was
associated with breast milk levels of HCB (3rd vs. 1st tertile HCB, OR = 1.49, 95% CI 1.10–2.03),
dieldrin (OR = 1.26, 95% CI 0.93–1.71) and mirex (OR = 1.36, 95% CI 0.99–1.86) (Dewailly et
al., 2000). This study observed stronger associations for a history of at least 3 occurrences of otitis
media before age 1 yr and breast milk levels of HCB (3rd vs. 1st tertile HCB, OR = 3.71, 95%
CI 1.10–12.6) and dieldrin (OR = 3.50, 95% CI 0.95–13.0) (Dewailly et al., 2000). In a crosssectional study of German children age 7–10 yr, a history of physician-diagnosed otitis media was
unrelated to HCB levels (≥0.2 vs. <0.2 µg/L, OR = 1.25, 95% CI 0.60–2.62, among the subset of
children with DDE <0.3 µg/L) (Karmaus et al., 2001). This study did observe an association
between otitis media and above-median blood levels of both DDE and HCB (OR= 2.38, 95% CI
1.08–5.25).
Environmental tobacco smoke Childhood exposure, sufficient evidence: A pooled-analysis of
seven studies found an association between recurrent middle ear infections and smoking by either
or both parents (pooled OR=1.48, 95% CI = 1.08–2.04) (Strachan & Cook, 1998a). A review of
studies published up to 1998 concluded that middle ear disease was associated with parental
smoking by either parent with stronger associations among pre-school compared to older children
(Cook & Strachan, 1999). Four recent reviews, including two by expert panels, concluded that
childhood ETS exposure can cause acute and chronic childhood middle ear infections (California
Environmental Protection Agency, 2005; Jaakkola & Jaakkola, 2002; U.S. Department of Health
and Human Services, 2006; World Health Organization, 1999).
Other indoor air pollutants Childhood exposure, biomass smoke, limited evidence: A WHO
review panel found limited evidence for an association between childhood otitis media and exposure to biomass smoke in developing countries (Bruce et al., 2002).
Outdoor air pollution Childhood exposure, limited evidence: Reviewers found limited
evidence for an association between otitis media and ambient air pollution (Bernard et al., 2006). A
large cross-sectional study in the former East Germany reported associations between childhood
middle ear infections and residential ambient air pollutant levels (per 50 µg/m3 total suspended particulate increment, OR=1.45, 95% CI 0.89–2.37; per 100 µg/m3 SO2 increment, OR=1.42, 95%
CI 0.94–2.15) (Heinrich et al., 2002). In a small Spanish birth cohort study, otitis media before age
1 yr was associated with residence in regions with high ambient SO2 levels (median annual SO2,
75 vs. 20 µg/m3, OR=2.01, 95% CI 1.05–3.84), independent of ETS and other potential confounders (Caceres Udina et al., 2004). In a large pregnancy cohort study in the Netherlands and
Germany, otitis media by age 2 yr was associated with residential ambient air pollutant levels
(Netherlands, per 3 µg/m3 PM2.5 increment, OR=1.13, 95% CI 1.00–1.27; per 10 µg/m3 NO2
increment, OR=1.14, 95% CI 1.03–1.27; similar results for Germany) (Brauer et al., 2006).
Summary Epidemiologic evidence for the role of environmental toxicants in childhood middle
ear infections includes: (a) sufficient evidence—childhood ETS exposure; (b) limited evidence—
lactational PCB exposure; lactational or childhood DDT/DDE exposure or childhood exposure to
biomass smoke or outdoor air pollution, mainly from traffic.
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
451
Childhood Cancer
About 10,000 new-onset childhood cancers were anticipated to be diagnosed among children
age 0–14 yr in the United States during 2006 (American Cancer Society, 2006). U.S. childhood
brain cancer, leukemia, and neuroblastoma incidence rates increased during the mid–1980s and
then stabilized, while Hodgkin’s disease rates decreased and leukemia rates may have declined
slightly (Linet et al., 1999). In contrast, overall childhood cancer incidence rates in Europe increased by
1.0% per year in children age 0–14 and by 1.5% in youth age 15–19 during 1970–1999 (SteliarovaFoucher et al., 2004). Cancer incidence rate trends reflect both true changes in risk over time but also
any changes in diagnostic and reporting practices/efficiencies. Although it is not clear whether there
have been sustained true increased cancer risks among children, the persistence of this devastating
disease justifies enhanced research into environmental and other preventable causes.
Reviewers concluded that epidemiologic studies published up to early 1998 provided fairly
consistent evidence of associations between parental occupational or childhood pesticide exposure
and risk of childhood cancers, including leukemia, brain cancer, neuroblastoma, non-Hodgkin’s
lymphoma (NHL), Wilms’s tumor, and Ewing’s sarcoma (Zahm & Ward, 1998). A recent review
noted that 15 case-control, 4 cohort, and 2 ecologic studies have since been published and that 15
of these 21 studies observed statistically significant increased risks between childhood cancer and
parental occupational or childhood pesticide exposure (Infante-Rivard & Weichenthal, 2006). The
authors concluded that there is an association between pesticide exposure and childhood cancer
but the epidemiologic evidence is insufficient to prove a cause–effect relationship.
Although pesticides and some other environmental exposures (e.g., ionizing radiation) may increase
overall cancer risk, discussion here focuses on specific types of childhood cancer. As discussed later, and
summarized in Table 6, there is limited epidemiologic evidence for associations between several types of
childhood cancer and parental or childhood exposure to pesticides, ETS or solvents.
Childhood Leukemia Childhood leukemia appears to have a clonal origin, developing from a
single abnormal precursor cell over a period of several months to a leukemia cell burden of about
1012 cells (Ford et al., 1998; Ma et al., 1999; Mori et al., 2002; Taub & Ge, 2004). Analysis of routinely collected neonatal blood samples has revealed leukemia clones with specific chromosomal
translocations (e.g., MLL-AF4[t(4;11)], MLL-AF4[t(12;21)]) in children who later developed acute
lymphatic leukemia (ALL), suggesting that many such cases originate in utero (Gale et al., 1997). It is
not known whether the abnormal cells detected in neonatal blood samples lead inevitably to childhood leukemia, but it appears that preleukemic clones can persist during childhood and that only a
minority progress to leukemia, suggesting that postnatal exposures may be required for such
progression (Maia et al., 2004).
Metals, metalloids Parental or childhood exposure, inadequate evidence: In a Denver
case-control study, leukemia was not associated with likely paternal occupational arsenic exposure
(based on job title) (OR=0.7, 95% CI 0.2–2.7) (Feingold et al., 1992). A large population-based
case-control study in Quebec found no association between childhood ALL and average prenatal
drinking-water arsenic (>5 vs. ≤5 µg/L, OR=0.94, 95% CI 0.49–1.81) or cumulative prenatal
exposure (>95th vs. ≤95th percentile, µg-d/L, OR=0.70, 95% CI 0.39–1.25); there was also no
association with postnatal arsenic exposure indices or with lead or cadmium indices (Infante-Rivard
et al., 2001). An ecologic study in Nevada reported no increased leukemia in areas where drinkingwater arsenic levels were 10–25 (SIR=0.61, 95% CI 0.12–1.79) or 35–90 µg/L (SIR=0.86, 95% CI
0.37–1.70) (Moore et al., 2002).
PCBs Childhood exposure, inadequate evidence: A small German case-control study of childhood leukemia found no association with bone-marrow PCBs (mean concentration, cases vs. controls, 4.21 and 3.38 µg/g lipid, p=.28) (Scheele et al., 1992).
TCDD Paternal occupational exposure, inadequate evidence: Among children of sawmill
workers in British Columbia, Canada, leukemia was not associated with paternal chlorophenate
exposure duration (≥3000 vs. <3000 h cumulated exposure, OR=0.8, 95% CI 0.2–3.6) (Heacock
et al., 2000). An expert panel found insufficient evidence for an association between paternal
phenoxy herbicide exposure and childhood cancer (National Academy of Sciences, 2003).
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D. T. WIGLE ET AL.
TABLE 6. Role of Environmental Toxicants in Childhood Cancer
Toxicant
Exposure
Lead
Prenatal
I
Paternal
Childhood I
Prenatal
As—I
Cd—I
Pb—I
Paternal
As —I
Childhood As—I
Cd—I
Pb—I
Childhood I
Paternal
I
Other metals,
metalloids
PCBs
TCDD and related
toxicants
Chlorophenate wood
preservatives
Herbicides
Insecticides
Fungicides
Soil fumigants
Unspecified pesticides
Active smoking
Environmental
tobacco Smoke
Outdoor air pollution
Drinking water DBPs
Drinking water
nitrate, nitrite
Hazardous waste
disposal sites
Various and
unspecified solvents
Leukemia
Lymphomak
I
L
Prenatal
Paternal
Childhood
Prenatal
Childhood
Prenatal
I
I
L
I
L
I
As—I
I
Childhood I
Paternal
I
Childhood
Prenatal
L
Paternal
Childhood
Prenatal
Paternal
Childhood
Brain Neuroblastoma Wilms’s tumor Other cancers
I
I
I
I
I
I
I
I
I
L
I
L
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
Paternal
I
L
L
Childhood
Prenatal
Paternall
Prenatal
Childhood
Prenatal
Paternalm
Childhood
Prenatal
Childhood
Prenatal
Childhood
Childhood
I
I
L
L
L
I
L
L
I
I
I
I
I
I
I
L
L
L
I
L
I
L
I
I
I
Prenatal
L
Paternal
I
Childhood L
I
I
I
I
I
I
I
I
I
I
I
I
I
I
Note. TCDD, 2,3,7,8-tetrachlorodibenzo-p-dioxin.
k
Most studies assessed non-Hodgkin’s lymphoma.
l
Paternal smoking could act as a source of ETS exposure and/or through paternal germ-cell mutations.
m
Paternal occupational exposure to motor vehicle exhaust emissions.
Germ cell—I
Soft tis sarc—I
Germ cell—I
Soft tis sarc—I
Germ cell—I
Germ cell—I
Germ cell—I
Bone—I
Germ cell—I Bone—I
Soft tis sarc—I
Eye tumors - I
Germ cell—I
Bone—L
Soft tis sarc—I
Eye tumors—I
Germ cell—I Bone—I
Germ cell—I
Germ cell—I
Germ cell—I
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
453
Childhood exposure, inadequate evidence: Follow-up of Seveso residents age 0–19 yr at the
time of the 1976 chlorophenol plant explosion revealed no overall excess of incident cancer by
1986 (obs=17, exp=13.6, SIR=1.2, 95% CI 0.7–2.1); there was a statistically nonsignificant
excess of leukemia and other hematopoietic cancers (SIR=1.6, 95% CI 0.7–3.4) (Pesatori et al.,
1993). Much longer follow-up is needed to assess the risk of cancer after longer latent periods.
Pesticides Although pesticides may increase overall childhood cancer risk, discussion here
focuses on specific types of cancer. A review of 18 epidemiologic studies published up to early
1998 noted that 13 studies reported increased childhood leukemia risk in relation to parental occupational or childhood pesticide exposure but did not clearly define the most critical exposures with
regard to timing (preconceptual vs. gestational vs. childhood) or parent (maternal vs. paternal)
(Zahm & Ward, 1998). A recent review noted that the 12 studies published since the Zahm and
Ward paper generally support an association between childhood leukemia and pesticide exposure
with the greatest risks being childhood exposure to household insecticides and parental exposure to
pesticides before or during pregnancy (Infante-Rivard & Weichenthal, 2006). Other reviewers
reached similar conclusions (Brown 2006; Buffler et al., 2005).
Maternal exposure, herbicides, limited evidence: As shown in Figure 3, leukemia was
associated with prenatal residential herbicide application in case-control studies in Los Angeles
(herbicides or other pesticides, OR=9.0, 95% CI 1.25–394; the high odds ratio was based on 9/1
discordant pairs3 in a matched-pairs analysis) (Lowengart et al., 1987), Quebec (any herbicide use,
OR=1.84, 95% CI 1.32–2.57; used more than 5 times vs. no use, OR=3.72, 95% CI 0.72–19.1)
(Infante-Rivard et al., 1999), California (prenatal residential herbicide use, OR=1.6, 95% CI 0.9–
3.0) (Ma et al., 2002), United States/Canada (frequency of use ≥median vs. none, OR=1.84, 95%
CI 0.91–3.73, p-trend=.13) (Alderton et al., 2006) and France (prenatal outdoor use of herbicides,
OR=5.9, 95% CI 0.7–52, 6 exposed case mothers) (Menegaux et al., 2006). The Quebec study
also reported an association between ALL and prenatal use of herbicides alone (OR=1.56, 95% CI
0.96–2.55) (Infante-Rivard et al., 1999).
Maternal exposure, insecticides, limited evidence: As shown in Figure 4, leukemia was associated with prenatal indoor insecticide use but not residential proximity to agricultural insecticide use.
There were elevated risks in case-control studies in Los Angeles (≥1 vs. 0 times/wk, OR=3.25, 95%
Study
Lowengart
Comparison
garden, 1+/mth
Yr
1987
Infante
1–5 X
1999
Infante
5+ X
1999
y/n
2002
Ma
Reynolds
nearby simazine
2005
Reynolds
nearby trifluralin
2005
Alderton
freq > median
2006
y/n
2006
Menegaux
Odds ratio and 95% CI
0.1 0.2 0.5 1
2
5 10
FIGURE 3. Childhood leukemia vs. prenatal herbicide exposure (mth=month, X=times during gestation, y/n=any exposure vs. none).
3
Ratio of the number of matched pairs in which the case was exposed but not the control to the number of matched pairs in which
the control was exposed but not the case.
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454
D. T. WIGLE ET AL.
Study
Comparison
Yr
Lowengart
home 1+ /wk
1987
Buckley
home <1/wk
1989
Odds ratio and 95% CI
home 1+/wk
Infante-Rivard
Leiss
Infante-Rivard
occup
1991
pest strip
1995
home
1999
plants
Alexander
home
2001
Ma
home
2002
nearby carb
2005
Reynolds
nearby OP
Alderton
home
2006
Menegaux
home
2006
0.1 0.2
0.5
1
2
5
10
FIGURE 4. Childhood leukemia vs. prenatal insecticide exposure (occup=occupation, carb=carbamates and OP=organophosphates).
CI 1.00–13.7) (Lowengart et al., 1987), United States/Canada (≥1/wk, OR=1.47, 95% CI 0.69–3.17,
p-trend=.05) (Buckley et al., 1989), Denver (prenatal use of insecticide pest strips in home,
OR=3.0, 95% CI 1.6–5.7) (Leiss & Savitz, 1995), Quebec (prenatal use of indoor insecticides for
cockroaches, ants, flies, bees or wasps, all cases, OR=1.79, 95% CI 1.34–2.40, cases with
CYP1A1m2 polymorphism, OR=4.73, 95% CI 1.18–18.6) (Infante-Rivard et al., 1999), a sevencountry study (leukemia before age 18 mo vs. maternal use of the carbamate insecticide propoxur,
OR=5.14, 95% CI 1.27–20.9) (Alexander et al., 2001), California (prenatal indoor insecticide use,
OR=2.1, 95% CI 1.3–3.5) (Ma et al., 2002), United States/Canada (ALL, any prenatal home extermination, OR=2.25, 95% CI 1.13–4.49; cases and controls all had Down’s syndrome) (Alderton
et al., 2006), and France (prenatal home use of insecticides, OR=1.8, 95% CI 1.2–2.8) (Menegaux
et al., 2006). Leukemia was associated with prenatal residential use of insecticides for interior or
outdoor plants (any use, OR=1.97, 95% CI 1.32–2.94; used more than 5 times, OR=4.01, 95% CI
1.12–14.3) and with prenatal professional indoor insect extermination (OR=1.68, 95% CI 0.87–
3.25) in Quebec (Infante-Rivard et al., 1999). There was a dose-response relationship between
leukemia and duration or frequency of prenatal indoor insecticide use in the first United States/Canada
study (Buckley et al., 1989). In a recent United States/Canada study, there was a nonmonotonic relationship between ALL risk and frequency of prenatal home extermination (<median vs. none,
OR=3.44, 95% CI 1.41–8.39; ≥median, OR=1.28, 95% CI 0.46–3.55) (Alderton et al., 2006).
There was no association between leukemia and prenatal occupational insecticide exposure in a
Spanish case-control study (OR=1.40, 95% CI 0.44–4.41, 7 exposed case mothers) (Infante-Rivard
et al., 1991). Childhood leukemia was associated with prenatal use of tree insecticides in a Quebec
case-control study (any use, OR=1.70, 95% CI 1.12–2.59; used more than 5 times, OR=3.27,
95% CI 0.64–16.7) (Infante-Rivard et al., 1999). A French case-control study observed a statistically
nonsignificant elevated leukemia risk related to prenatal garden insecticide use (OR=1.9, 95% CI
0.6–6.5) (Menegaux et al., 2006).
Maternal exposure, unspecified pesticides, inadequate evidence: Elevated leukemia risks were
related to prenatal occupational pesticide exposure in case-control studies in Shanghai (ALL,
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455
OR=3.5, 95% CI 1.1–11.2; ANLL, OR=2.4, 95% CI 0.5–11.0) (Shu et al., 1988), United States/
Canada (crude OR=2.85, 95% CI 0.82–10.8, calculated from data in paper) (Buckley et al., 1989),
Spain (occupation in farming, OR=1.80, 95% CI 0.60–6.64 (Infante-Rivard et al., 1991), Germany
(regional study, OR=2.59, 95% CI 0.45–20.4, only 4 exposed case mothers) (Meinert et al., 1996),
Greece (home or occupational exposure, OR=3.6, 95% CI 1.2–10.8) (Petridou and Dessypris
2000), Germany (nation-wide, OR=3.6, 95% CI 1.5–8.8) (Meinert et al., 2000), a seven-country
study (leukemia before age 18 mos, OR=3.67, 95% CI 1.54–8.74) (Alexander et al., 2001), and
Israel (occupational exposure of either parent, OR=2.35, 95% CI 1.10–5.0) (Abadi-Korek et al.,
2006). Elevated leukemia risk was also related to maternal outdoor pesticide use in case-control
studies in Germany (prenatal or postnatal use, OR=2.76, 95% CI 1.26–6.30) (Meinert et al., 1996)
and France (prenatal, OR=2.5, 95% CI 0.8–7.2) (Menegaux et al., 2006).
There was no association in case-control studies in the Netherlands (prenatal occupational pesticide exposure, OR=0.7, 95% CI 0.2–2.5, only 4 exposed case mothers) (van Steensel-Moll et al.,
1985), Denver (prenatal outdoor herbicide or insecticide use, OR=1.1, 95% CI 0.6–1.9) (Leiss &
Savitz, 1995) or United Kingdom (periconceptual maternal occupations potentially exposed to
agrochemicals, OR=0.81, 95% CI 0.31–2.12) (McKinney et al., 2003). The consistency of findings
are suggestive of an association, but the heterogeneity of exposure indices, the lack of demonstrated dose-response relationships, and the relatively small numbers of exposed case mothers in
most studies preclude firm conclusions.
Paternal occupational exposure, major pesticide classes, inadequate evidence: In the Quebec
study, leukemia risk was elevated in relation to preconceptual paternal occupational exposure to
herbicides (OR=2.05, 95% CI 0.93–4.56), insecticides (OR=1.38, 95% CI 0.87–2.18) and fungicides (OR=5.11, 95% CI 1.46–17.8) (Infante-Rivard & Sinnett 1999).
Paternal occupational exposure, unspecified pesticides, inadequate evidence: Elevated leukemia
risk was associated with paternal occupational pesticide exposure in case-control studies in Italy
(OR=5.6, 95% CI 1.3–24.3, 5 case fathers worked in farming) (Magnani et al., 1990), United
States/Canada (1–1000 d cumulated exposure vs. none during period from 1 yr before birth to date
of diagnosis, OR=1.0, 95% CI 0.4–2.4); >1000 d vs. none, OR=2.7, 95% CI 1.0–7.0,
p-trend=.04) (Buckley et al., 1989), United Kingdom (paternal occupation in farming, OR=1.98,
95% CI 0.66–5.96) (Gardner et al., 1990), and Germany (occupational pesticide exposure in year
before conception, OR=1.5, 95% CI 1.1–2.2) (Meinert et al., 2000).
Leukemia was not associated with paternal occupational pesticide exposure in case-control
studies in Quebec (paternal work as farmer based on birth certificate information, OR=0.70, 95%
CI 0.39–1.21) (Fabia & Thuy, 1974), the Netherlands (paternal work in agriculture, forestry or horticulture, OR=0.9, 95% CI 0.5–1.5; self-reported occupational pesticide exposure, OR=1.0, 95%
CI 0.6–1.7) (van Steensel-Moll et al., 1985), Germany (paternal occupational pesticide exposure in
year before pregnancy, OR=1.29, 95% CI 0.46–3.62) (Meinert et al., 1996), Sweden (OR=0.90,
95% CI 0.37–2.19) (Feychting et al., 2001), or the United Kingdom (periconceptual paternal
occupations potentially exposed to agrochemicals, OR=0.83, 95% CI 0.58–1.19) (McKinney et al.,
2003). In a Norwegian retrospective cohort study, leukemia risk among offspring of farmers (84% of
whom were males) was not elevated compared to nonfarm children (all farm children, SIR=0.96,
95% CI 0.82–1.11; children on farms reporting pesticide expenditures, SIR=1.06, 95% CI 0.75–
1.49) (Kristensen et al., 1996a). Two cohort studies reported normal leukemia risks among offspring
of licensed pesticide applicators: a Swedish study (OR=0.43, 95% CI 0.19–0.86) (Rodvall et al.,
2003) and a U.S. study (OR=0.91, 95% CI 0.47–1.75) (Flower et al., 2004). Despite the
dose-response relationship observed in one study, the inconsistent findings, heterogeneous exposure indices and negative results of two cohort studies of licensed pesticide applicators provide
inadequate evidence of an association.
Childhood exposure, herbicides, inadequate evidence: The Quebec study reported an association between leukemia and childhood residential herbicide use (with or without residential use of
other pesticides, OR=1.41, 95% CI 1.06–1.86) but not with residential use of herbicides alone
(OR=0.88, 95% CI 0.58–1.33) (Infante-Rivard et al., 1999). An ecologic study in Maryland
observed borderline associations between leukemia and residence at diagnosis <3.2 km from well
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D. T. WIGLE ET AL.
water containing detectable levels of the herbicides atrazine (OR=1.43, 95% CI 0.89–2.30) and
metolachlor (OR=1.48, 95% CI 0.93–2.36) but not simazine (OR=0.97, 95% CI 0.48–1.96)
(Thorpe & Shirmohammadi, 2005). There was a statistically nonsignificant elevated leukemia risk
related to childhood residential herbicide use in a French case-control study (OR=1.4, 95% CI
0.8–2.4) (Menegaux et al., 2006). An ecologic study in California found no association between
total leukemia before age 15 yr and childhood residential proximity to agricultural use (≥90th vs.
<1st percentile, lb/mi2) of the herbicides simazine (a triazine herbicide, OR=0.79, 95% CI 0.45–
1.40) or trifluralin (a dinitroaniline herbicide, OR=0.87, 95% CI 0.46–1.63) (Reynolds et al.,
2002a). Further analysis restricted to ALL before age 15 yr revealed a slightly elevated risk related to
simazine use (≥75th vs. <1st percentile, lb/mi2, OR=1.21, 95% CI 0.86–1.71) but not trifluralin
(OR=0.81, 95% CI 0.49–1.34) (Reynolds et al., 2005a). A California case-control study restricted
to ALL before age 5 revealed elevated risks related to childhood residential proximity to agricultural
use (≥50th vs. <1st percentile, lb/mi2) of simazine (OR=1.29, 95% CI 0.81–2.05) but not trifluralin
(OR=0.92, 95% CI 0.51–1.65) (Reynolds et al., 2005b).
Leukemia was not associated with childhood residential herbicide use in Denver (OR=0.9,
95% CI 0.5–1.8) (Leiss & Savitz, 1995), California (birth to age 1, OR=0.7, 95% CI 0.4–1.2; age 1–
2, OR=1.1, 95% CI 0.7–2.0; age 2–3, OR=1.1, 95% CI 0.6–2.1) (Ma et al., 2002) or in a United
States/Canada study (≥median frequency of use vs. none, OR=1.07, 95% CI 0.42–2.67,
p-trend=.66) (Alderton et al., 2006).
Childhood exposure, insecticides, limited evidence: As shown in Figure 5, leukemia was associated with childhood indoor residential use of insecticides but not residential proximity to agricultural insecticide use. Leukemia risk was elevated in relation to childhood indoor insecticide use in
case-control studies in United States/Canada (≥1×/wk, OR=2.02, 95% CI 0.91–4.57, p-trend=.04)
(Buckley et al., 1989), Denver (insecticide pest strips4 in home, birth to 2 yr before diagnosis,
Study
Comparison
Yr
Buckley
home <1/wk
1989
Odds ratio and 95% CI
home 1+/wk
Leiss
Infante-Rivard
pest strip
1995
home
1999
plants
Meinert
home 11+/yr
2000
home 6–10/yr
Ma
Reynolds
home, age <4
2002
nearby carb
2005
nearby OP
Alderton
home
2006
Menegaux
home
2006
0.1 0.2
0.5
1
2
5
10
FIGURE 5. Childhood leukemia vs. childhood insecticide exposure (carb=carbamates, OP=organophosphates).
4
Pest strips contained dichlorvos, a highly volatile organophosphate insecticide; it is a known mutagen and animal carcinogen
causing leukemia and lung and mammary-gland tumors.
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
457
OR=1.7, 95% CI 1.2–2.4; <2 yr before diagnosis, OR=2.6, 95% CI 1.7–3.9) (Leiss & Savitz, 1995),
Quebec (indoor insecticide use for cockroaches etc., all ALL cases, OR=1.38, 95% CI 1.07–1.77; ALL
cases with CYP1A1m2 polymorphism, OR=3.95, 95% CI 0.81–19.2) (Infante-Rivard et al., 1999),
Germany (indoor insecticide use >10 vs. <1×/yr, OR=1.8, 95% CI 1.0–3.3, p-trend=.12)
(Meinert et al., 2000), California (indoor insecticide use, birth to age 1, OR=1.7, 95% CI 1.0–2.9;
age 1–2, OR=1.6, 95% CI 1.0–2.7; age 2–3, OR=1.2, 95% CI 0.7–2.1) (Ma et al., 2002), United
States/Canada (frequency of use ≥median vs. none, OR=1.63, 95% CI 0.84–3.30, p=.21) (Alderton
et al., 2006) and France (indoor insecticide use, OR=1.7, 95% CI 1.1–2.4; childhood use of insecticide shampoo, OR=1.9, 95% CI 1.1–3.2) (Menegaux et al., 2006). The French study also
reported an association between leukemia and childhood use of insecticidal shampoos (OR=1.9,
95% CI 1.1–3.2) (Menegaux et al., 2006). Dose-response relationships between leukemia and
frequency of childhood residential indoor insecticide use were apparent in Germany (p-trend=.12)
(Meinert et al., 2000) and California (per increment in frequency of use, birth to age 1, OR=1.2,
95% CI 1.0–1.4; age 1–2, OR=1.1, 95% CI 1.0–1.3) (Ma et al., 2002). Childhood indoor insecticide use was not associated with leukemia in the Denver case-control study (home extermination,
<2 yr before diagnosis, OR=0.9, 95% CI 0.5–1.4) (Leiss & Savitz, 1995). The apparent interaction
between indoor insecticide exposure and the CYP1A1m2 polymorphism in the Quebec study may
indicate a role for P-450 enzyme activation; however, this appears to be the only study of interactions between genetic polymorphisms and pesticide exposure in childhood cancer and more
research is urgently needed.
ALL was associated with childhood outdoor residential insecticide use in case-control studies in
Quebec (tree insecticides, OR=1.41, 95% CI 1.01–1.97) (Infante-Rivard et al., 1999), California
(ALL age <5, childhood residential proximity, agricultural use of organophosphate insecticide,
≥50th vs. <1st percentile, lb/mi2, OR=1.22, 95% CI 0.96–1.56) (Reynolds et al., 2005b), and
France (OR=2.4, 95% CI 1.3–4.3) (Menegaux et al., 2006). The California study also observed
elevated risks of leukemia before age 5 yr related to childhood residential proximity to agricultural
use of the organochlorine miticide dicofol (OR=1.83, 95% CI 1.05–3.22) but not propargite
(OR=0.96, 95% CI 0.62–1.49) or the broad classes of organochlorine (OR=1.29, 95% CI 0.78–
2.13) or carbamate insecticides (OR=1.08, 95% CI 0.80–1.47) (Reynolds et al., 2005b). An
ecologic study in California found no association between total leukemia before age 15 yr and childhood residential proximity to agricultural use (≥90th vs. <1st percentile, lb/mi2) of organochlorine (OR=0.70,
95% CI 0.39–1.23), organophosphate (OR=0.91, 95% CI 0.70–1.18) or carbamate insecticides (OR=1.03,
95% CI 0.75–1.41) (Reynolds et al., 2002a). Further analysis restricted to ALL before age 15
revealed no association with agricultural use of organochlorine (≥75th vs. <1st percentile, lb/mi2,
OR=0.73, 95% CI 0.47–1.15), organophosphate (OR=0.94, 95% CI 0.77–1.14), or carbamate
insecticides (OR=0.87, 95% CI 0.67–1.13) or propargite (OR=1.03, 95% CI 0.76–1.39) (Reynolds
et al., 2005a). The consistency of findings and evidence of dose-response relationships provide
relatively strong evidence of an association.
Childhood exposure, fungicides, inadequate evidence: Leukemia was associated with
childhood residential garden fungicide use in a French case-control study (OR = 2.5, 95% CI
1.0–6.2) (Menegaux et al., 2006). An ecologic study in California found no association
between leukemia and childhood residential proximity to agricultural use (≥90th vs. <1st
percentile, lb/mi2) of dithiocarbamate fungicides (OR = 0.89, 95% CI 0.61–1.30) (Reynolds et
al., 2002a). This study did report elevated leukemia risk related to agricultural use of the chlorinated isophthalic acid fungicide chlorothalonil (OR = 1.27, 95% CI 0.90–1.80). Further analysis, restricted to ALL before age 15 yr and using different cut points of pesticide use intensity,
revealed no association with agricultural use of dithiocarbamate fungicides (≥75th vs. <1st
percentile, lb/mi2, OR=0.92, 95% CI 0.70–1.19) (Reynolds et al., 2005a). A California case-control
study restricted to ALL before age 5 yr revealed no association with childhood residential proximity
to agricultural use (≥50th vs. <1st percentile, lb/mi2) of dithiocarbamate fungicides (OR = 1.01,
95% CI 0.71–1.42) and a statistically nonsignificant elevated risk related to childhood residential proximity to agricultural use of chlorothalonil (OR = 1.33, 95% CI 0.88- 2.01) (Reynolds
et al., 2005b).
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D. T. WIGLE ET AL.
Childhood exposure, soil fumigants, inadequate evidence: An ecologic study in California
found no association between total leukemia and childhood residential proximity to agricultural use
(≥90th vs. <1st percentile, lb/mi2) of the soil fumigant metam sodium (OR=0.92, 95% CI 0.48–
1.73) (Reynolds et al., 2002a). A subsequent case-control study restricted to ALL before age 5
revealed an association with childhood residential proximity to agricultural use (≥50th vs. <1st
percentile, lb/mi2) of metam sodium (OR=2.05, 95% CI 1.01–4.17) but not methyl bromide
(OR=0.89, 95% CI 0.60–1.33) (Reynolds et al., 2005b).
Childhood exposure, residential use of unspecified pesticides, inadequate evidence: A
case-control study in northwestern Germany reported an association between leukemia and prenatal
or childhood garden pesticide use (crude OR=2.76, 95% CI 1.26–6.30, calculated from data in
paper) but did not distinguish between prenatal or childhood exposure (Meinert et al., 1996).
Childhood outdoor residential pesticide use was not associated with childhood leukemia in
case-control studies in Denver (yard insecticide or herbicide use, OR=0.9, 95% CI 0.5–1.8) (Leiss &
Savitz, 1995) or a nationwide German study (residential garden pesticide exposure use, OR=1.0,
95% CI 0.8–1.2) (Meinert et al., 2000).
Environmental tobacco smoke Prenatal active smoking, inadequate evidence: Prenatal active
smoking was not associated with childhood leukemia in a meta-analysis of 8 studies (summary
OR=1.05, 95% CI 0.82–1.34) (Boffetta et al., 2000). A Quebec study reported elevated risks in
relation to 2nd or 3rd trimester maternal active smoking among subgroups of childhood leukemia
based on CYP1A1 polymorphisms (e.g., 3rd trimester smoking >20 vs. 0 cigarettes/d, OR=2.8
(95% CI 0.8–9.8) for CYP1A1*2A and 5.4 (95% CI 0.8–37.3) for CYP1A1*4 polymorphisms
(Infante-Rivard et al., 2000). A case-control study in France revealed no overall association between
childhood leukemia and prenatal active smoking (OR=1.2, 95% CI 0.7–2.1); case-only analyses
showed associations with prenatal smoking among infants with CYP1A1 (*1/2A or *2A/2A)
(OR=2.2, 95% CI 1.0–4.9) or GSTM1 null polymorphisms (OR=2.3, 95% CI 1.2–4.4) (Clavel
et al., 2005). A subsequent report of this study revealed no association between ALL or ANLL and
postnatal maternal or paternal smoking (e.g., ALL, maternal smoking birth to interview, ≥20
cigarettes/d, OR=0.6, 95% CI 0.5–1.6) (Menegaux et al., 2005). A population-based case-control
study in California revealed no association between preconceptual or prenatal active smoking and
ALL or AML (e.g., ALL, prenatal smoking, OR=0.93, 95% CI 0.58–1.51) (Chang et al., 2006).
Prenatal ETS exposure, limited evidence: Few studies have assessed the independent effect of
prenatal ETS exposure and childhood cancer risk, i.e., independent from maternal or paternal
active smoking. An expert group convened by the State of California found inadequate evidence of
a causal association between childhood leukemia and prenatal ETS exposure (California Environmental Protection Agency, 2005). The U.S. Surgeon General concluded that there is suggestive evidence of a causal relationship between childhood leukemia and prenatal or postnatal ETS exposure
but not sufficient for a firm conclusion (U.S. Department of Health and Human Services, 2006).
Paternal active smoking, limited evidence: A meta-analysis of 4 epidemiologic studies indicated a weak association between childhood ALL and paternal smoking (summary OR = 1.17,
95% CI 0.96–1.42) (Boffetta et al., 2000). An expert group convened by the State of California
found limited evidence of a causal association between childhood leukemia and preconceptual
paternal smoking (California Environmental Protection Agency, 2005). A large UK case-control
study reported no association between childhood leukemia and paternal preconceptual smoking
(≥20 vs. 0 cigarettes/d, OR = 1.01, 95% CI0.87–1.17) (Pang et al., 2003). In a population-based
case-control study in California, preconceptual paternal smoking was associated with AML
(OR = 3.84, 95% CI 1.04–14.2) and to a lesser degree with ALL (OR= 1.32, 95% CI 0.86–2.04)
(Chang et al., 2006).
Childhood exposure, limited evidence: The California expert panel found inadequate evidence
of a causal association between childhood leukemia and childhood ETS exposure (California Environmental Protection Agency, 2005). The U.S. Surgeon General concluded that there is suggestive
evidence of a causal relationship between childhood leukemia and prenatal or postnatal ETS exposure but insufficient for a firm conclusion (U.S. Department of Health and Human Services, 2006).
In a population-based case-control study in California, ALL was associated with postnatal maternal
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459
active smoking among the subgroup of children whose fathers smoked preconceptually (OR=3.94,
95% CI 1.25–12.4) (Chang et al., 2006).
Outdoor air pollution Based on studies of experimental animals and occupationally exposed
adults, diesel and gasoline engine emissions, respectively, are classified as probable and possible
human carcinogens (International Agency for Research on Cancer, 1989). However, a WHO expert
panel noted that the few available studies of childhood cancer and outdoor air pollution had inconsistent results and there was inadequate evidence for a causal association (Binkova et al., 2005).
A systematic review of eight case-control and seven ecologic studies of all childhood cancers
combined and air pollution noted that four relatively small case-control studies (including two in
Denver) reported associations and these studies all had methodologic limitations (Raaschou-Nielsen &
Reynolds, 2006). Two case-control studies with high statistical power, validated exposure assessment
methods, substantial exposure gradient and low potential for bias or confounding, reported no
association (Raaschou-Nielsen et al., 2001; Reynolds et al., 2004). However, some of the studies
that reported no association between air pollution exposure indices and all childhood cancer types
combined did observe associations with specific types of childhood cancer (see later discussion).
Maternal exposure, inadequate evidence: A large Danish case-control study reported no association between childhood leukemia and cumulative prenatal exposure to benzene based on
outdoor air sampling near the maternal residence (≥90th vs. <50th percentile of ppb-days,
OR=0.8, 95% CI 0.4–1.9) or NO2 (Raaschou-Nielsen et al., 2001). A large California-wide
case-control study revealed no association between leukemia and traffic density near the maternal
residence at birth (≥90th vs. <29th percentile of vehicle-miles/mi2, OR=0.92, 95% CI 0.73–1.15)
(Reynolds et al., 2004). A case-control study in France reported an association between leukemia
and prenatal residential proximity to car repair garages or gasoline stations (adjoining property,
OR=2.2, 95% CI 0.9–5.7) but not heavy-traffic roads (<50 vs. ≥50 m, OR=1.3, 95% CI 0.5–3.2)
(Steffen et al., 2004).
Paternal exposure, motor vehicle emissions, limited evidence: Reviewers concluded that the
preponderance of epidemiological evidence suggests an association between childhood leukemia/
lymphoma and paternal vehicle-related occupations likely exposed to motor vehicle emissions or
solvents (Olshan & van Wijngaarden, 2003).
Childhood residential exposure, limited evidence: A case-control study in Denver, CO,
revealed a dose-response relationship between leukemia and traffic density on the street of residence at diagnosis (≥10,000 vs. <500 vehicles/d, OR = 4.7, 95% CI 1.6–13.5) (Savitz & Feingold,
1989). This association occurred among cases age 0–4 (OR = 5.6, 95% CI 1.9–16.7) but not older
children. In a re-analysis of the Denver study, leukemia was associated with the highest traffic
density on any street within 0.25 km of residence at diagnosis (≥500 vs. <500 vehicles/d,
OR = 2.08, 95% CI 1.06–4.07) (Pearson et al., 2000). In a Swedish ecologic study, motor vehicle
density in the municipality of residence was associated with AML (≥20 vs. <5 vehicles/km2,
OR = 1.62, 95% CI 0.91–2.91) but not ALL (OR = 0.88, 95% CI 0.69–1.12) among persons age
0–24 (Nordlinder & Jarvholm, 1997). A small Swedish case-control study (only 39 cases) revealed
a statistically nonsignificant and imprecise association between childhood leukemia and average
ambient air NO2 levels (an indicator of motor vehicle emissions) near the residence during the
year of diagnosis (≥50 vs. <40 µg/m3, OR = 2.7, 95% CI 0.3–20.6) (Feychting et al., 1998). In a
UK case-control study, there were elevated leukemia risks for children living <100 m from a
main road (OR= 1.61, 95% CI 0.90–2.87) or gasoline station (OR = 1.99, 95% CI 0.73–5.43)
(Harrison et al., 1999). In a California ecologic study, leukemia was weakly associated with traffic
density near the childhood residence (≥90th vs. <25th percentile of vehicles/d/mi2, OR = 1.15,
95% CI 0.97–1.37) (Reynolds et al., 2002b). An extension of this study reported a similar association between leukemia and airborne levels, in the census tract of childhood residence at diagnosis, of 25 hazardous air pollutants (HAP) classified as known, probable or possible human
carcinogens by the U.S. versus <25th percentile of carcinogenic HAP index, OR = 1.21, 95% CI
1.03–1.42, p-trend<.05) (Reynolds et al., 2003). A small case-control study of leukemia in Los
Angeles County revealed an association with traffic density near the childhood residence of
longest duration (5th vs. 1st quintile, OR = 1.9, 95% CI 0.9–3.7) which weakened after
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D. T. WIGLE ET AL.
adjustment for electric wire code (OR = 1.4, 95% CI 0.7–3.0) (Langholz et al., 2002). A casecontrol study in France reported an association between leukemia and childhood residential proximity to car repair garages or gasoline stations (adjoining property, OR=4.0, 95% CI 1.5–10.3) but
not heavy-traffic roads (<50 vs. ≥50 m, OR = 1.3, 95% CI 0.6–2.9) (Steffen et al., 2004). The latter study also noted a dose-response relationship between leukemia and duration of childhood
residence near a car repair garage or gasoline station (≥36 vs. 0 mo, OR = 4.7, 95% CI 1.2–18.5,
p-trend<.05). An Italian case-control study reported an association between leukemia and ambient air benzene levels, estimated from modeling of vehicle density on roads <0.3 km from childhood residence (>10 vs. <0.1 µg/m3, OR = 3.91, 95% CI 1.36–11.3, p-trend = .005) (Crosignani
et al., 2004). In a Taiwanese case-control study, residential proximity to 4 petrochemical production complexes (involved in petroleum cracking and production of many products including vinyl
chloride monomer, polyethylene, acrylates, MTBE) was associated with leukemia among persons
age 20–29 yr (per log exposure score increment, OR = 1.54, 95% CI 1.14–2.09) but not with
childhood leukemia (OR = 1.04, 95% CI 0.79–1.38) (Yu et al., 2006). In a large Danish casecontrol study, leukemia was not associated with cumulative childhood exposure to ambient air
benzene (90th vs. <50th percentile of ppb-days, OR=0.4, 95% CI 0.1–1.6) or NO2 (RaaschouNielsen et al., 2001) near the residence at diagnosis.
Drinking-water disinfection by-products contaminants MX comprised 49% of mutagenic
activity in Massachusetts municipal water supplies (Wright et al., 2002). Among DBPs studied to
date, known animal carcinogens include chloroform, bromoform, BDCM, HAAs, haloacetonitriles,
bromate and MX (Komulainen, 2004).
Prenatal or childhood exposure, inadequate evidence: A large population-based case-control
study of ALL in Quebec reported no association with prenatal THM exposure indices (>95th vs.
≤95th percentile average THM, OR=1.05, 95% CI 0.46–2.39; >95th vs. ≤95th percentile cumulative THM, OR=0.83, 95% CI 0.39–1.76) (Infante-Rivard et al., 2001). There was a statistically
nonsignificant elevated ALL risk among children above the 95th percentile of cumulative childhood
(OR=1.54, 95% CI 0.78–3.03) but not average childhood exposure (OR=1.08, 95% CI 0.55–
2.13). Case-only analyses within this study revealed strong but imprecise associations between ALL
and average childhood THM levels above 100 µg/L among GSTT1-null (OR=9.1, 95% CI 1.4–58)
and CYP2E1*5 children (OR=4.1, 95% CI 0.8–22) (Infante-Rivard et al., 2002). There were similar
associations with other THM exposure indices among children with such polymorphisms but the
imprecise odds ratios (because of small numbers) preclude strong inferences.
Drinking water nitrate Prenatal or childhood exposure, inadequate evidence: The Quebec
study case-control study described above revealed no association between childhood ALL and
time-weighted average drinking water nitrate level (>2 vs. ≤2 mg/L, prenatal, OR=0.7, 95% CI
0.3–1.7; postnatal, OR=0.6, 95% CI 0.2–1.6); there were similar results when exposure was
modeled as cumulative exposure (Infante-Rivard et al., 2001).
Hazardous waste disposal sites Childhood residential proximity, inadequate evidence: A
small case-control study in Woburn, MA, revealed a statistically nonsignificant elevated risk of
childhood leukemia related to prenatal or childhood consumption of contaminated drinking water
(main contaminant was trichloroethylene), OR=2.4, 95% CI 0.5–11) (Massachusetts Department
of Public Health, 1997). An ecologic study in the United Kingdom found no association between
leukemia and childhood residential proximity to hazardous waste disposal sites (<2 vs. ≥2 km,
OR=0.93, 95% CI 0.81–1.06) (Jarup et al., 2002). In a large U.S. case-control study, Wilms’s
tumor risk was reduced among offspring of women living within 1.6 km of a NPL site (OR=0.35,
95% CI 0.12–0.99); there was a similar relationship with childhood residential proximity (Tsai
et al., 2006).
Solvents Maternal exposure, specified solvents, inadequate evidence: A large Quebec casecontrol study revealed associations between ALL and prenatal occupational exposure to specific
solvents or solvent categories including methylene chloride (probable or definite exposure,
OR=3.22, 95% CI 0.88–11.7), toluene (OR=1.98, 95% CI 1.06–3.72), mineral spirits (OR=1.74,
95% CI 0.99–3.06), alkanes (OR=1.78, 95% CI 1.09–2.91), and mononuclear aromatic hydrocarbons (OR=1.67, 95% CI 1.13–2.48) (Infante-Rivard et al., 2005).
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Maternal exposure, unspecified solvents, limited evidence: The Children’s Cancer Study Group
(United States and Canada) found a dose-response relationship between childhood acute nonlymphoblastic leukemia and cumulative prenatal occupational solvent exposure (>1000 vs. 0 days,
OR=2.2, 95% CI 0.9–5.4, p-trend=0.05) (Buckley et al., 1989). In a much larger United States–
Canada study, childhood acute lymphocytic leukemia was associated with prenatal occupational
exposure to solvents or products containing solvents (solvents, OR=1.6, 95% CI 1.1–2.3; paints or
thinners, OR=1.7, 95% CI 1.2–2.3) (Shu et al., 1999). A pooled analysis of three German populationbased case-control studies indicated that acute lymphocytic leukemia was associated with prenatal
occupations likely exposed to paints or lacquers (OR=2.0, 95% CI 1.2–3.3) but not with solvents
per se (OR=1.3, 95% CI 0.8–1.9) (Schuz et al., 2000). A U.S. case-control study revealed an
association between ALL and periconceptual household painting (≥5 vs. 0 rooms, OR=1.7, 95% CI
1.1–2.7, p-trend=0.01) (Freedman et al., 2001). A large Danish case-control study found no
association between childhood leukemia and cumulative prenatal residential exposure to ambient
air benzene (90th vs. <50th percentile of ppb-days, OR=0.8, 95% CI 0.4–1.9) (Raaschou-Nielsen
et al., 2001). In a large case-control study, the subgroup of childhood ALL cases with K-ras
mutations was associated with prenatal occupational exposure to solvents, degreasers or cleaning
agents (OR=3.1, 95% CI 1.0–9.7) but not paints or thinners (OR=1.0, 95% CI 0.2–4.4) (Shu et al.,
2004). A large Quebec case-control study revealed no association between ALL and prenatal occupational exposure to the broad category of solvents (OR=1.00, 95% CI 0.78–1.28) (Infante-Rivard
et al., 2005).
Paternal occupational exposure, unspecified solvents, inadequate evidence: Reviewers noted
that all five epidemiologic studies of childhood leukemia and paternal occupational solvent exposure published up to 1993 found positive associations; significantly associated solvents included
chlorinated solvents and benzene (Colt & Blair, 1998). A review noted that studies published since
the 1998 Colt and Blair review generally did not support an association between childhood leukemia/
lymphoma and paternal occupational solvent exposure (Olshan & van Wijngaarden, 2003). Among
subsequently reported studies, there was no association between childhood leukemia and paternal
occupational exposure to solvents or products containing solvents in a United States–Canada study
(solvents, OR=1.1, 95% CI 0.9–1.3; paints or thinners, OR=1.0, 95% CI 0.8–1.2) (Shu et al.,
1999), a pooled analysis of three German population-based case-control studies (solvents,
OR=1.0, 95% CI 0.8–1.2; paints or lacquers, OR=1.1, 95% CI 0.9–1.4) (Schuz et al., 2000), or a
Swedish study (OR=1.3, 95% CI 0.8–2.0) (Feychting et al., 2001). In a large case-control study, the
subgroup of childhood ALL cases with K-ras mutations was not associated with preconceptual
paternal occupational exposure to solvents, degreasers or cleaning agents or paints/thinners
(OR=1.0, 95% CI 0.2–4.4) (Shu et al., 2004).
Childhood exposure, unspecified solvents, limited evidence: The Children’s Cancer Study Group
(United States and Canada) reported a nonmonotonic dose-response relationship between childhood acute nonlymphoblastic leukemia and frequency of household use of petroleum products (≥4
vs. 0 times/month, OR=1.8, 95% CI 0.7–4.3, p-trend=0.02) (Buckley et al., 1989). A record-based
case-control study in the United Kingdom found borderline associations between childhood leukemia and residential proximity to main roads (<100 vs. ≥100 m, OR=1.6, 95% CI 0.9–2.9) and
gasoline stations (OR=2.0, 95% CI 0.7–5.4) (Harrison et al., 1999). A large Danish case-control
study found no association between childhood leukemia and cumulative childhood residential exposure to ambient air benzene (90th vs. <50th percentile of ppb-days, OR=0.4, 95% CI 0.1–1.6)
(Raaschou-Nielsen et al., 2001). A large U.S. case-control study revealed an association between
childhood ALL and childhood artwork involving solvent exposure during the year before diagnosis
(high frequency vs. unexposed, OR=4.1, 95% CI 1.1–15.1, p-trend=0.07) (Freedman et al., 2001).
Summary Limited epidemiologic evidence supports a role for environmental toxicants in
childhood leukemia, including (a) prenatal residential or occupational exposure to herbicides,
insecticides, ETS or unspecified solvents; (b) paternal occupational exposure to motor vehicle
exhaust emissions or active smoking (not clear if paternal smoking acts through germ-cell mutations
or fetal/childhood ETS exposure); (c) childhood residential exposure to insecticides, ETS, outdoor
air pollution (mainly traffic-related) or unspecified solvents.
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D. T. WIGLE ET AL.
Childhood Lymphoma
Pesticides Among seven studies published by early 1998, three reported increased childhood
nonHodgkin’s lymphoma (NHL) risks related to indoor insecticide use during pregnancy or
childhood (two studies) or residence on farms with relatively high pesticide exposures (one study);
the authors noted that each study had few exposed cases or case parents (Zahm & Ward, 1998). A
recent review of the four studies published since Zahm and Ward’s report concluded that they
provided further evidence for an association between NHL and pesticide exposure, especially residential indoor insecticide use during pregnancy or childhood for which there were dose-response
relationships (Infante-Rivard & Weichenthal, 2006).
Maternal exposure, insecticides, inadequate evidence: The United States/Canada case-control
study reported a dose-response relationship between NHL and frequency of prenatal indoor insecticide use (1–2 vs. 0×/wk, OR=2.62, 95% CI 0.96–7.18; ≥3×/wk, OR=7.33, 95% CI 0.84–64,
p-trend=.05) (Buckley et al., 2000). This study also reported a dose-response relationship between
NHL and frequency of prenatal outdoor insecticide use (any use vs. none, OR=2.98, 95% CI 1.44–
6.16, p-trend=.002) (Buckley et al., 2000). In the Denver case-control study, lymphoma was not
associated with 3rd trimester home extermination (OR=1.2, 95% CI 0.4–3.9) or indoor use of
insecticide pest strips (OR=1.4, 95% CI 0.7–2.5) (Leiss & Savitz, 1995). Although suggestive, the
findings of Buckley et al. require replication.
Maternal exposure, unspecified pesticides, inadequate evidence: The United States/Canada
case-control study reported statistically nonsignificant elevated NHL risks related to occupational
pesticide exposure of either parent (OR=1.74, 95% CI 0.82–3.69) and prenatal use of garden
pesticide sprays (≥1/mo vs. never, OR=1.71, 95% CI 0.67–4.37) (Buckley et al., 2000). A German
case-control study observed an association between NHL and self-reported prenatal occupational
pesticide exposure (OR=11.8, 95% CI 2.2–64, 4 exposed case mothers) (Meinert et al., 2000). In
the Denver case-control study, lymphoma was not associated with 3rd trimester outdoor use of
herbicides and/or insecticides (OR=0.5, 95% CI 0.2–1.2, 6 exposed case mothers) (Leiss & Savitz,
1995). The small number of studies, heterogeneity of exposure indices, and weak findings preclude
firm conclusions.
Paternal occupational exposure, unspecified pesticides, limited evidence: As shown in Figure 6,
there was a slightly elevated risk of Hodgkin’s disease among Norwegian farm children (parents
worked on farm at least 500 hr/yr, compared to nonfarm families, SIR=1.17, 95% CI 0.85–1.56)
Study
Kristensen
Comparison
Hodg, pestic spray $
Yr
Odds ratio and 95% CI
1996
NHL, pestic $ 1
NHL, pestic $ 2
NHL, pestic $ 3
Meinert
occup expos, y/n
2000
Buckley
occup expos, y/n, m/p
2000
Rodvall
applicator
2003
Flower
applicator
2004
0.1 0.2 0.5 1
2
5 10
FIGURE 6. Childhood lymphoma vs. paternal exposure to unspecified pesticides (Hodg=Hodgkin’s Disease, pestic spray
$=expenditures on pesticide spraying equipment, pestic $ 1=low expenditures on pesticides, 2=medium, 3=high, m/p=prenatal or
paternal).
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463
(Kristensen et al., 1996a). There was also a dose-response relationship between NHL risk and pesticide expenditures on Norwegian farms (level 1 vs. none, OR=1.30, 95% CI 0.49–3.42; level 2,
OR=1.57, 95% CI 0.75–3.30; level 3, OR=2.50, 95% CI 1.02–6.15). In a United States/Canada
case-control study, there was a statistically nonsignificant elevated risk of NHL related to selfreported occupational pesticide exposure of either parent (OR=1.74, 95% CI 0.82–3.69) (Buckley
et al., 2000). A case-control study in Germany found a statistically nonsignificant elevated risk of
NHL related to paternal occupational pesticide exposure in the year before conception (OR=1.5,
95% CI 0.7–3.1) (Meinert et al., 2000). Among offspring of licensed pesticide applicators,
lymphoma risk was not elevated in a Swedish study (SIR=0.94, 95% CI 0.44–1.79, 8 observed
cases) (Rodvall et al., 2003) but was increased in the AHS cohort (SIR=2.18, 95% CI 1.13–4.19, 9
observed cases) (Flower et al., 2004). The relatively consistent findings, including the results of the
AHS cohort (likely the best cohort study of pesticide risks ever done) and dose-response relationship provide limited evidence of an association.
Childhood exposure, residential use or proximity to agricultural use of herbicides, inadequate
evidence: In the California ecologic study, there was no association between NHL and childhood
residential proximity to agricultural use of the herbicides trifluralin (≥75th vs. <1st percentile,
OR=0.47, 95% CI 0.08–2.78) or simazine (OR=0.76, 95% CI 0.14–4.06) (Reynolds et al., 2005a).
Similarly, there was no association between Hodgkin’s disease and childhood residential proximity
to agricultural use of trifluralin (OR=1.13, 95% CI 0.40–3.17) or simazine (OR=0.59, 95%
CI 0.25–1.41).
Childhood exposure, insecticides, inadequate evidence: In a case-control study in Denver,
lymphoma was associated with childhood house extermination (OR=1.8, 95% CI 1.1–2.9) but not
with indoor use of insecticide pest strips (OR=1.3, 95% CI 0.4–2.7) (Leiss & Savitz, 1995). A German
case-control study revealed a dose-response relationship between NHL and childhood indoor insecticide use (>10 vs. <1 time/yr, OR=2.8, 95% CI 1.1–7.2, p-trend=.02) (Meinert et al., 2000).
In a California-wide ecologic study, there was no association between childhood NHL and
childhood residential proximity to agricultural use of organochlorine (75th vs. <1st percentile,
lb/mi2, OR=1.11, 95% CI 0.50–2.47), organophosphate (OR=0.83, 95% CI 0.50–1.37) or carbamate insecticides (OR=0.84, 95% CI 0.49–1.43) (Reynolds et al., 2005a). This study found statistically
nonsignificant elevated risks of Hodgkin’s disease related to organochlorine (OR=1.31, 95% CI
0.70–2.46) and carbamate insecticides (OR=1.30, 95% CI 0.77–2.20) and dicofol (1.43, 95% CI
0.70–2.95) but not organophosphate insecticides (OR=1.12, 95% CI 0.71–1.76) or propargite
(OR=0.85, 95% CI 0.39–1.86).
Childhood exposure, fungicides, inadequate evidence: In the California ecologic study, there
was no association between childhood residential proximity to agricultural use of dithiocarbamate
fungicides and NHL (OR=0.52, 95% CI 0.23–1.19) but Hodgkin’s disease risk was elevated
(OR=1.38, 95% CI 0.78–2.44) (Reynolds et al., 2005a). This study found no association between
NHL or Hodgkin’s disease and childhood residential proximity to agricultural use of soil fumigants
(methyl bromide, metam sodium) or the fungicide chlorothalonil.
Childhood exposure, soil fumigants, inadequate evidence: The California ecologic study
observed no association between NHL or Hodgkin’s disease and childhood residential proximity to
agricultural use of the soil fumigants methyl bromide or metam sodium (Reynolds et al., 2005a).
Childhood exposure, unspecified pesticides, inadequate evidence: As noted earlier, children
on Norwegian farms reporting pesticide expenditures had an increased risk of NHL (highest vs. no
expenditure, OR=2.50, 95% CI 1.02–6.15) with evidence of a dose-response relationship
(Kristensen et al., 1996a). Given the design of the Norwegian study, farm pesticide exposures were
indices of both prenatal parental and childhood exposures. NHL was associated with direct
childhood exposure to herbicides or insecticides in a United States/Canada case-control study
(yes vs. no, OR=2.35, 95% CI 1.37–4.03, p-trend=.001 based on exposure frequency); the report
did not distinguish between indoor and outdoor exposure (Buckley et al., 2000). In the Denver
case-control study, lymphoma was not associated with postnatal outdoor use of herbicides and/or
insecticides (OR=0.8, 95% CI 0.3–1.8) (Leiss & Savitz, 1995). Although suggestive, the small
number of studies and heterogeneous exposure indices preclude firm conclusions.
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D. T. WIGLE ET AL.
Tobacco smoke Prenatal active smoking, inadequate evidence: A meta-analysis of 6 epidemiologic studies found no association between childhood NHL and prenatal active smoking (summary
OR=1.13, 95% CI 0.85–1.49) (Boffetta et al., 2000).
Prenatal ETS exposure, limited evidence: An expert group noted inadequate evidence for an
association between childhood lymphomas and prenatal and postnatal ETS exposure (California
Environmental Protection Agency, 2005). The U.S. Surgeon General concluded that there is suggestive evidence of a causal relationship between childhood lymphoma and prenatal or postnatal ETS
exposure (U.S. Department of Health and Human Services, 2006).
Paternal active smoking, limited evidence: A meta-analysis of 4 epidemiologic studies observed
an association between childhood NHL and paternal smoking (summary OR=2.08, 95% CI 1.08–
3.98) (Boffetta et al., 2000). An expert group convened by the State of California found limited
evidence of a causal association between childhood lymphomas and preconceptual paternal smoking;
their report noted that paternal smoking might act through preconceptual paternal germ-cell mutations or by increasing prenatal ETS exposure (California Environmental Protection Agency, 2005).
Childhood ETS exposure, limited evidence: An expert group noted inadequate evidence for
this association (California Environmental Protection Agency, 2005). The U.S. Surgeon General
concluded that there is suggestive evidence of a causal relationship between childhood lymphoma
and prenatal or postnatal ETS exposure (U.S. Department of Health and Human Services, 2006).
Outdoor air pollution Maternal exposure, major ambient air pollutants, inadequate evidence:
A large Danish case-control study revealed associations between childhood Hodgkin’s disease and
cumulative prenatal residential exposure to ambient air NO2 (90th vs. <50th percentile of
ppb-days, OR=6.7, 95% CI 1.7–26.0, p-trend=0.02) and benzene (90th vs. <50th percentile of
ppb-days, OR=4.3, 95% CI 1.5–12.4, p-trend=0.005) (Raaschou-Nielsen et al., 2001).
Childhood exposure, major ambient air pollutants, inadequate evidence: In a Swedish ecologic
study, motor vehicle density in the municipality of residence was not associated with NHL among
persons age 0–24 yr (≥20 vs. <5 vehicles/km2, OR=1.09, 95% CI 0.59–2.05) (Nordlinder &
Jarvholm, 1997). The Danish case-control study reported no association between childhood
lymphomas and cumulative childhood residential exposure to ambient air benzene (90th vs. <50th
percentile of ppb-days, OR=0.4, 95% CI 0.1–2.0) or NO2 (Raaschou-Nielsen et al., 2001).
Solvents Maternal exposure, benzene, inadequate evidence: As noted earlier, cumulative
prenatal ambient air benzene exposure was associated with childhood Hodgkin’s disease (≥90th vs.
<50th percentile of ppb-days, OR=4.3, 95% CI 1.5–12.4, p-trend=.005) (Raaschou-Nielsen
et al., 2001).
Paternal occupational exposure, inadequate evidence: Reviewers noted that all 5 reviewed
studies of childhood leukemia/lymphoma and paternal occupational solvent exposure found ORs
exceeding 2 but there were no data for lymphoma alone (Colt & Blair, 1998).
Childhood exposure, benzene, inadequate evidence: As noted earlier, childhood lymphomas
were not associated with cumulative childhood exposure to ambient air benzene (≥90th vs. <50th
percentile of ppb-days, OR=0.4, 95% CI 0.1–2.0) (Raaschou-Nielsen et al., 2001).
Summary There is limited epidemiologic evidence for the role of environmental toxicants in
childhood lymphomas including prenatal ETS exposure, paternal occupational exposure to unspecified pesticides, paternal active smoking (not clear whether paternal smoking acts through germ-cell
mutations or fetal/childhood ETS exposure) and childhood ETS exposure.
Childhood Brain Cancer
Arsenic Paternal occupational exposure, inadequate evidence: In a Denver case-control
study, brain cancer was not associated with likely paternal occupational arsenic exposure (based on
job title) (OR=1.3, 95% CI 0.4–4.1) (Feingold et al., 1992).
Pesticides A review noted that 12 of 16 case-control studies published by early 1998 revealed
increased childhood brain cancer risks related to pesticide exposure with statistically significant relationships in 7 studies (Zahm & Ward, 1998). Risks were generally highest for parental pesticide use
in the home, garden, or on pets. The only cohort study observed an association with paternal
employment in farming and one of the case-control studies found an association with prenatal
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household insecticide use or at least 1 yr of farm residence. A review of the nine studies published
since the Zahm and Ward report found general support for an association between childhood brain
cancer and pesticide exposure, especially prenatal and indoor residential insecticide use, but there
were no statistically significant dose-response relationships (Infante-Rivard & Weichenthal, 2006).
Maternal exposure, herbicides, inadequate evidence: There was no association between brain
cancer and prenatal residential herbicide use in Missouri (lawn herbicides, OR=1.1, 95% CI 0.5–
2.5) (Davis et al., 1993) or Denver (outdoor insecticide or herbicide use, OR=0.6, 95% CI 0.3–1.1)
(Leiss & Savitz, 1995). A large United States/Canada case-control study reported no association
between maternal occupational herbicide exposure and astrocytoma (OR=1.3, 95% CI 0.5–3.7) or
PNET (OR=0.5, 95% CI 0.2–1.5) (van Wijngaarden et al., 2003). In a California case-control study,
brain tumors before age 5 yr were not associated with prenatal residential proximity to agricultural
use of the herbicides simazine (≥50th vs. <1st percentile of lb/mi2, OR=1.06, 95% CI 0.49–2.26)
or trifluralin (OR=0.88, 95% CI 0.26–2.98) (Reynolds et al., 2005b).
Maternal exposure, insecticides, limited evidence: As shown in Figure 7, prenatal residential
insecticide use was related to elevated childhood brain tumor risks in Missouri (OR=1.8, 95% CI
0.8–4.0; insecticide pest strips in home, OR=5.2, 95% CI 1.2–22.2) (Davis et al., 1993), France
(OR=1.8, 95% CI 0.8–4.1) (Cordier et al., 1994), the United States/Canada (astrocytoma, any
prenatal use, OR=1.5, 95% CI 0.8–2.7; used at least weekly, OR=2.2, 95% CI 0.6–7.4) (Bunin
et al., 1994), Australia (home extermination, OR=2.0, 95% CI 1.0–3.9) (McCredie et al., 1994a),
Denver (insecticide pest strips in home, OR=1.5, 95% CI 0.9–2.4; house extermination, OR=1.3,
95% CI 0.7–2.1) (Leiss & Savitz, 1995), and the western United States (direct prenatal use of flea/tick
insecticides, all cases, OR=2.2, 95% CI 1.1–4.2; cases age <5, OR=5.4, 95% CI 1.3–22.3)
(Pogoda and Preston-Martin, 1997). The western United States study also revealed associations
between brain cancer and prenatal number of pets treated with insecticides for fleas/ticks (1 pet,
OR=1.4, 95% CI 0.9–2.4; >1 pet, OR=2.0, 95% CI 1.0–4.0, p-trend=.04). Among children in
households with indoor insecticide use during pregnancy or childhood, brain cancer risk was
Study
Comparison
Davis
garden
Yr
Odds ratio and 95% CI
1993
home
pest strip
Cordier
Bunin
home
1994
home, astro
1994
home, PNET
McCready
home
1994
pest strip
1995
Pogoda
pet insectic
1997
van Wijn
occup
2003
Reynolds
nearby carb
2005
Leiss
nearby OP
0.1 0.2
0.5
1
2
5
10
FIGURE 7. Childhood brain cancer vs. prenatal insecticide exposure (astro=astrocytoma, PNET=primitive neuroectodermal tumor,
carb=carbamates, OP=organophosphates).
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D. T. WIGLE ET AL.
strongly associated with inefficient polymorphisms at C–108 in PON1, the gene that encodes
paraoxonase, an organophosphate detoxifying enzyme (CC, OR=1.0 (referent); CT, OR=2.6, 95%
CI 1.2–5.5; TT, OR=6.6, 95% CI 1.5–29.7) (Nielsen et al., 2005). There was no association
between PNET and prenatal indoor insecticide use in the United States/Canada study (any prenatal
use, OR=0.7, 95% CI 0.4–1.4; used at least weekly, OR=1.0, 95% CI 0.2–4.9) (Bunin et al., 1994).
Prenatal outdoor residential insecticide use was related to elevated childhood brain tumor risks
in Missouri (carbaryl, OR=1.5, 95% CI 0.7–3.3; diazinon, OR=4.6, 95% CI 1.2–17.9) (Davis
et al., 1993). A large United States/Canada case-control study reported an association between
maternal occupational insecticide exposure and astrocytoma (OR=1.9, 95% CI 1.1–3.3) but not
PNET (OR=1.0, 95% CI 0.6–1.7) (van Wijngaarden et al., 2003). In a California case-control study,
brain tumors before age 5 yr were not associated with prenatal residential proximity to agricultural
use of organochlorine (≥50th vs. <1st percentile of lb/mi2, OR=0.73, 95% CI 0.32–1.65), organophosphate (OR=1.10, 95% CI 0.74–1.66) or carbamate insecticides (OR=0.93, 95% CI 0.56–
1.57), propargite (OR=1.06, 95% CI 0.50–2.28), or dicofol (OR=0.65, 95% CI 0.27–1.61)
(Reynolds et al., 2005b).
In sum, there were elevated brain cancer risks in six of seven case-control studies of residential
indoor insecticide use with statistical significance apparent in four studies, a dose-response relationship in one study, and increased risk among children with inefficient polymorphisms at C–108 in
PON1, the gene that encodes paraoxonase, an organophosphate detoxifying enzyme.
Maternal exposure, fungicides, inadequate evidence: A large United States/Canada case-control
study reported an association between maternal occupational fungicide exposure and astrocytoma
(OR=1.6, 95% CI 0.9–2.7) but not PNET (OR=0.7, 95% CI 0.4–1.2) (van Wijngaarden et al.,
2003). In a California case-control study, brain tumors before age 5 yr were not associated with
prenatal residential proximity to agricultural use of dithiocarbamate fungicides (≥50th vs. <1st
percentile of lb/mi2, OR=0.89, 0.49–1.64) or chlorothalonil (OR=1.18, 95% CI 0.58–2.38)
(Reynolds et al., 2005b).
Maternal exposure, other specified pesticides, inadequate evidence: In a California case-control
study, brain tumors before age 5 yr were not associated with prenatal residential proximity to
agricultural use of the soil fumigants dimethyl bromide (OR=0.80, 95% CI 0.44–1.46) or metam
sodium (OR=0.91, 95% CI 0.27–3.08) (Reynolds et al., 2005b).
Maternal exposure, unspecified pesticides, inadequate evidence: Elevated brain tumor risk was
related to prenatal pesticide exposure in case-control studies in France (prenatal farm residence,
OR=2.5, 95% CI 0.4–16.1, 4 exposed case mothers) (Cordier et al., 1994), the United States/
Canada (farm residence, PNET, OR=3.7, 95% CI 0.8–23.9) (Bunin et al., 1994), western U.S.
states (prenatal agricultural pesticide use, OR=1.8, 95% CI 0.77–4.2) (Holly et al., 1998), and a
seven-country study (prenatal farm residence, OR=1.3, 95% CI 1.0–1.8; maternal agricultural
pesticide use, OR=2.0, 95% CI 1.2–3.2) (Efird et al., 2003).
In the United States/Canada study, astrocytomas were not associated with prenatal farm residence (OR=0.5, 95% CI 0.1–1.8, 4 exposed case mothers) (Bunin et al., 1994). Brain cancer was
not associated with prenatal pesticide exposure in Australia (maternal farm residence or farm work,
OR=0.9, 95% CI 0.3–2.6) (McCredie et al., 1994a), Europe (OR=0.5, 95% CI 0.2–1.4, 5 exposed
case mothers) (Cordier et al., 1997), or a seven-country study (maternal occupation in agriculture,
during 5 yr before birth, OR=1.1, 95% CI 0.7–1.9; during pregnancy, OR=1.4, 95% CI 0.6–3.0)
(Cordier et al., 2001). The inconsistent findings, lack of statistical significance in most of the positive
studies and the lack of demonstrated dose-response relationships comprise inadequate evidence
for an association.
Paternal exposure, chlorophenate wood preservatives, inadequate evidence: Among children
of sawmill workers in British Columbia, Canada, there was a statistically nonsignificant elevated
brain cancer risk related to paternal occupational chlorophenate exposure duration (≥3560 vs.
<3560 h cumulated exposure, OR=1.5, 95% CI 0.4–6.9) (Heacock et al., 2000).
Paternal exposure, broad pesticide classes, inadequate evidence: A large United States/Canada
case-control study reported associations between astrocytoma brain tumors and paternal occupational exposure to herbicides (astrocytoma, OR=1.6, 95% CI 1.0–2.7), insecticides (OR=1.5, 95%
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
467
CI 0.9–2.4) and fungicides (OR=1.6, 95% CI 1.0–2.6); PNET brain tumors were associated with
paternal exposure to herbicides (OR=1.5, 95% CI 0.9–2.6) but not insecticides or fungicides (van
Wijngaarden et al., 2003). These findings require replication and exploration of dose-response
relationships.
Paternal occupational exposure, unspecified pesticides, limited evidence: There was a
dose-response relationship between nonastrocytic neuroepithelial brain tumors and pesticide
expenditures in a Norwegian record-based cohort study of farm families (level 1 expenditures,
OR=2.00, 95% CI 0.85–4.74; level 2, OR=2.93, 95% CI 1.54–5.60; level 3, OR=3.28, 95% CI
1.39–7.76) (Kristensen et al., 1996a). This study noted a particularly high risk of such tumors among
children age <5 yr on grain farms with pesticide purchases (OR=8.01, 95% CI 1.62–39.7).
Because of the design of the Norwegian study, pesticide exposure indices reflect both prenatal
parental and childhood exposures. Brain cancer risk was elevated among children of licensed
pesticide applicators in the AHS cohort (SIR=1.60, 95% CI 0.89–2.89) (Flower et al., 2004) but not
in a Swedish record-based cohort (SIR=1.03, 95% CI 0.60–1.65) (Rodvall et al., 2003).
When paternal pesticide exposure was inferred from job titles, without other evidence of exposure, increased childhood brain cancer risk was noted in studies in Ohio (preconceptual paternal
occupation in farming, brain cancer deaths, OR=2.0, 95% CI 1.0–4.1) (Wilkins & Koutras, 1988),
another Ohio study (preconceptual paternal occupation in farming, brain cancer incident cases,
OR=2.7, 95% CI 0.8–9.1) (Wilkins & Sinks, 1990), 3 U.S. states (preconceptual paternal occupation in agriculture, OR=1.8, 95% CI 0.6–6.0) (Kuijten et al., 1992), France (preconceptual paternal
occupation in agriculture (OR=2.0, 95% CI 1.0–4.1) (Cordier et al., 1997), a seven-country study
(paternal occupation in agriculture during 5 yr before birth, OR=1.3, 95% CI 1.0–1.8) (Cordier
et al., 2001), and Sweden (potential paternal preconceptual occupational pesticide exposure,
RR=2.36, 95% CI 1.27–4.39) (Feychting et al., 2001). There was no association in case-control
studies in Quebec (occupation as farmer, OR=0.56, 95% CI 0.22–1.26) (Fabia and Thuy 1974) or
the United Kingdom (occupation in farming, OR=0.70, 95% CI 0.35–1.38) (McKinney et al.,
2003). The relatively consistent findings and dose-response relationship are suggestive of an
association.
Childhood exposure, herbicides, inadequate evidence: Brain cancer was associated with childhood residential herbicide use in a case-control study in Missouri (OR=2.4, 95% CI 1.0–5.7) (Davis
et al., 1993) but not in Denver (use of herbicides or insecticides in yard, OR=0.5, 95% CI 0.2–0.9)
(Leiss & Savitz, 1995) or the western United States (OR=1.2, 95% CI 0.3–4.9, 4 exposed case
mothers) (Pogoda and Preston-Martin, 1997). An ecologic study in California found no association
between brain gliomas and childhood residential proximity to agricultural use of the herbicides
simazine (≥90th vs. <1st percentile, OR=1.12, 95% CI 0.69–1.82) and trifluralin (OR=0.58, 95%
CI 0.27–1.25) (Reynolds et al., 2002a).
Childhood exposure, insecticides, limited evidence: As shown in Figure 8, indoor residential
insecticide use was associated with elevated brain cancer risks in case-control studies in Baltimore
(household insect extermination, OR=2.29, 95% CI 0.96–5.95) (Gold et al., 1979), Missouri
(indoor use of insecticides, OR=3.4, 95% CI 1.1–10.6; indoor insecticidal pest strips, OR=3.7,
95% CI 1.0–13.7; Kwell insecticidal shampoo, OR=4.6, 95% CI 1.0–21.3) (Davis et al., 1993),
France (home extermination, OR=2.0, 95% CI 1.0–4.1) (Cordier et al., 1994), Denver (insecticide
pest strips, OR=1.4, 95% CI 0.7–2.9; house extermination, OR=1.4, 95% CI 0.6–2.7) (Leiss &
Savitz, 1995), and Los Angeles (failure to evacuate house after indoor insecticide use, OR=1.6,
95% CI 1.0–2.6) (Pogoda & Preston-Martin, 1997). A German case-control study reported
statistically nonsignificant and non-dose-related elevated brain cancer risks related to indoor insecticide use (1×/yr vs. 0, OR=1.38, 95% CI 0.84–2.25; >1/yr, OR=1.19, 95% CI 0.81–1.77) (Schuz
et al., 2001).
The western United States study found no association with any indoor insecticide use
(OR=1.2, 95% CI 0.8–2.0) or flea/tick insecticide use (OR=1.0, 95% CI 0.7–1.4) (Pogoda &
Preston-Martin, 1997). An ecologic study in California found no association between brain gliomas
and childhood residential proximity to agricultural use of organochlorine (≥90th vs. <1st percentile,
OR=0.86, 95% CI 0.44–1.67), organophosphate (OR=0.71, 95% CI 0.50–1.02) or carbamate
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D. T. WIGLE ET AL.
Study
Comparison
Yr
Gold
home
1979
Davis
garden
1993
Odds ratio and 95% CI
home
pest strip
Cordier
Leiss
Pogoda
home
1994
pest strip
1995
home
1997
no evac
Schuz
>1/yr
2001
1/yr
0.1 0.2
0.5
1
2
5
10
FIGURE 8. Brain cancer versus childhood insecticide exposure (no evac=indoor extermination without home evacuation).
insecticides (OR=0.76, 95% CI 0.48–1.19) (Reynolds et al., 2002a). The consistency of findings is
suggestive of an association but further research is needed to explore dose-response relationships
and specific insecticides or groups of toxicologically related insecticides.
Childhood exposure, fungicides, inadequate evidence: An ecologic study in California found no
association between brain gliomas and childhood residential proximity to agricultural use of dithiocarbamate fungicides (≥90th vs. <1st percentile, OR=0.59, 95% CI 0.33–1.04) or chlorothalonil
(OR=0.47, 95% CI 0.23–0.97) (Reynolds et al., 2002a).
Childhood exposure, other specified pesticides, inadequate evidence: There was no association
between brain cancer and childhood residential use of wood preservatives (OR=1.26, 95% CI
1.00–1.59) (Schuz et al., 2001). An ecologic study in California found no association between brain
gliomas and childhood residential proximity to agricultural use of the soil fumigants methyl bromide
(≥90th vs. <1st percentile, OR=0.63, 95% CI 0.38–1.05) and metam sodium (OR=0.37, 95% CI
0.09–1.41) (Reynolds et al., 2002a).
Childhood exposure, unspecified pesticides, limited evidence: Brain tumors were associated
with outdoor residential pesticide use or proximity to agricultural pesticide use in case-control
studies in Baltimore (childhood farm residence, OR=4.00, 95% CI 1.21–17.7) (Gold et al., 1979),
France (farm residence, OR=6.7, 95% CI 1.2–38.0) (Cordier et al., 1994), the United States/Canada
(residence on farm for at least 1 yr, PNET, OR=5.0, 95% CI 1.1–46.8) (Bunin et al., 1994), the
western states in the United States (farm residence before age 6 mo, OR=1.9, 95% CI 0.96–3.8;
farm residence for over 1 yr, OR=1.7, 95% CI 0.88–3.1) (Holly et al., 1998), and a seven-country
study (childhood farm residence, OR=1.3, 95% CI 1.0–1.7; beginning before age 6 mo, OR=1.6,
95% CI 1.1–2.2) (Efird et al., 2003). As noted earlier, there was a dose-response relationship
between nonastrocytic neuroepithelial brain tumors and pesticide expenditures in a Norwegian
record-based cohort study of farm families, especially among children age <5 yrliving on grain
farms (Kristensen et al., 1996a). The design of the Norwegian study precluded clear distinction
between prenatal parental and childhood pesticide exposures.
No association was apparent in studies in Ontario (childhood pesticide exposure, OR=0.94, 95%
CI 0.47–1.90) (Howe et al., 1989), United States/Canada (childhood farm residence, OR=0.4,
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95% CI 0.1–1.6) (Bunin et al., 1994) or Australia (lived or worked on farm, OR=0.6, 95% CI 0.2–
1.9, only 4 exposed cases) (McCredie et al., 1994b). Although there was uncertainty about the
timing of pesticide exposure in the Norwegian cohort, the observation of associations in the larger
studies, the strength of associations (odds ratios in 3 studies exceeded 4) and the fact that 2 of the 3
studies with negative findings had less than 100 cases is suggestive of an association.
Environmental tobacco smoke The U.S. Surgeon General concluded that there is suggestive
evidence of a causal relationship between childhood brain cancer and prenatal and postnatal ETS
exposure (U.S. Department of Health and Human Services, 2006).
Prenatal active smoking, inadequate evidence: A meta-analysis of 12 epidemiologic studies
indicated no association between childhood brain and other central nervous system cancers and
prenatal active smoking (summary OR=1.04, 95% CI 0.92–1.18) (Boffetta et al., 2000).
Prenatal ETS exposure, limited evidence: The California expert panel found no substantial
evidence for an association between childhood brain tumors and prenatal active smoking or ETS
exposure (California Environmental Protection Agency, 2005). The U.S. Surgeon General
concluded that there is suggestive evidence of a causal relationship between childhood brain cancer
and prenatal or postnatal ETS exposure but insufficient for a firm conclusion (U.S. Department of
Health and Human Services, 2006).
Paternal smoking, limited evidence. A meta-analysis of 10 epidemiologic studies indicated a weak
association between childhood brain cancer and prenatal paternal smoking (summary OR=1.22,
95% CI 1.05–1.40) (Boffetta et al., 2000). In an international case-control study, brain cancer before
age 1 yr was weakly associated with preconceptual paternal smoking (OR=1.4, 95% CI 0.9–2.1);
there was no such association among older children (Filippini et al., 2002). A recent very large UK
case-control study reported no association between central nervous system tumors and preconceptual
paternal smoking (OR=1.03, 95% CI 0.82–1.28) (Pang et al., 2003). See also next section.
Childhood exposure, limited evidence. The California expert panel found limited evidence of a
causal association between childhood brain cancer and postnatal ETS exposure, mainly related to
paternal smoking but stated that this association may reflect an effect of preconceptual paternal
smoking (California Environmental Protection Agency, 2005). The U.S. Surgeon General concluded
that there is suggestive evidence of a causal relationship between childhood brain cancer and
prenatal or postnatal ETS exposure but insufficient for a firm conclusion (U.S. Department of Health
and Human Services, 2006).
Outdoor air pollution Maternal exposure, inadequate evidence: In a large Danish case-control
study, childhood CNS tumors were not associated with cumulative prenatal exposure to ambient air
benzene (≥90th vs. <50th percentile of ppb-days, OR=0.4, 95% CI 0.1–1.3) or NO2 (Raaschou-Nielsen
et al., 2001). A large California-wide case-control study revealed a weak association between
CNS tumors and traffic density near the maternal residence at birth (≥90th vs. <29th percentile of
vehicle-miles/mi2, OR=1.22, 95% CI 0.87–1.70) (Reynolds et al., 2004).
Childhood exposure, inadequate evidence: A case-control study in Denver, Colorado revealed
a dose-response relationship between traffic density on the street of residence at diagnosis and
brain tumors among children age 0–4 yr (≥10,000 vs. <500 vehicles/d, OR=5.2, 95% CI 1.4–19.6)
but not those in older children (OR=1.5, 95% CI 0.5–5.9) (Savitz and Feingold, 1989). A small
Swedish case-control study (only 33 cases) revealed a statistically nonsignificant and imprecise
association between childhood CNS tumors and average NO2 levels (an indicator of motor vehicle
emissions) near the residence during the year of diagnosis (≥50 vs. <40 µg/m3, OR=5.1, 95% CI
0.4–61) (Feychting et al., 1998). Childhood CNS tumors were not associated with cumulative
childhood exposure to ambient air benzene (≥90th vs. <50th percentile of ppb-days, OR=0.6,
95% CI 0.2–1.7) or NO2 (Raaschou-Nielsen et al., 2001). In a California ecologic study, brain
tumors were weakly associated with traffic density near the childhood residence (≥90th vs. <25th
percentile of vehicles/d/mi2, OR=1.14, 95% CI 0.90–1.45) (Reynolds et al., 2002b). An extension
of this study reported no association between brain gliomas and airborne levels, in the census tract
of childhood residence at diagnosis, of 25 hazardous air pollutants (HAP) classified as known, probable, or possible human carcinogens by the U.S. EPA (≥90th vs. <25th percentile of carcinogenic
HAP index, OR=0.98, 95% CI 0.80–1.21, p-trend>.05) (Reynolds et al., 2003).
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D. T. WIGLE ET AL.
Drinking-water nitrate/nitrite Maternal exposure, nitrate, inadequate evidence: In a California/
Washington State case-control study, childhood brain tumors were associated with drinking water
nitrite levels at the maternal residence (≥1 vs. <1 mg/L, OR=8.8, 95% CI 2.1–46) but not nitrate
levels (≥10 vs. <10 mg/L, OR=0.6, 95% CI 0.3–1.1) (Mueller et al., 2001). A large international
case-control study revealed no association between childhood brain tumors and drinking water
nitrate or nitrite levels at the residence where the mother lived during the relevant pregnancy
(nitrate ≥50 vs. <10 mg/L, OR=1.0, 95% CI 0.4–2.2; nitrite ≥5 vs. <1 mg/L, OR=2.1, 95% CI
0.6–7.4) (Mueller et al., 2004). However, there was an association between astroglial brain tumors
and tap water nitrite levels (OR=5.7, 95% CI 1.2–27).
Solvents Prenatal occupational exposure, unspecified solvents, inadequate evidence: A populationbased case-control study in France reported an association between childhood brain tumors and
maternal employment in occupations likely exposed to solvents during the 5 yr before the child’s
birth (high vs. no exposure, OR=2.4, 95% CI 1.2–4.9) (Cordier et al., 1997). A large Danish casecontrol study found no association between childhood CNS tumors and cumulative prenatal residential exposure to ambient air benzene (90th vs. <50th percentile of ppb-days, OR=0.4, 95% CI
0.1–1.3) (Raaschou-Nielsen et al., 2001).
Paternal occupational exposure, unspecified solvents, inadequate evidence: In a Texan casecontrol study, childhood CNS tumor deaths were associated with paternal occupations with likely
solvent exposure including printing (OR=4.5, 95% CI 1.4–14.7) and petroleum refining (OR=2.7,
95% CI 0.9–7.8) but not painting (OR=1.0, 95% CI 0.3–3.3) (Johnson et al., 1987). The authors
reported a summary odds ratio of 5.0 (p < 0.05) for this association based on 7 studies published
up to 1982. During recent years, childhood CNS tumors were not associated with paternal occupational solvent exposure in studies in Denver (preconceptual exposure, yes vs. no, OR=1.2, 95% CI
0.2–8.5) (Feingold et al., 1992), France (paternal employment in occupations likely exposed to
solvents during 5 yr before child’s birth, high vs. no exposure, OR=1.2, 95% CI 0.7–1.9) (Cordier
et al., 1997), or Sweden (OR=1.2, 95% CI 0.7–1.9) (Feychting et al., 2001).
Childhood exposure, benzene, inadequate evidence: A large Danish case-control study found
no association between childhood CNS tumors and cumulative childhood residential exposure to
ambient air benzene (90th vs. <50th percentile of ppb-days, OR=0.6, 95% CI 0.2–1.7)
(Raaschou-Nielsen et al., 2001).
Summary There is limited epidemiologic evidence for the role of environmental toxicants in
childhood brain cancer including: (a) prenatal occupational or residential exposure to insecticides;
(b) paternal occupational exposure to unspecified pesticides or active smoking (not clear if paternal
smoking acts through germ-cell mutations or fetal/childhood ETS exposure); (c) childhood residential exposure to insecticides or unspecified pesticides.
Neuroblastoma
Lead Maternal occupational exposure, inadequate evidence: One case-control study reported an
association between childhood neuroblastomas and self-reported prenatal occupational lead exposure
(OR=4.7, 95% CI 1.3–18.2); the association was somewhat attenuated when exposure was defined as
self-reported plus expert-rated potential for such exposure (OR=3.5, 95% CI 0.7–22.6) (Kerr et al., 2000).
Paternal occupational exposure, limited evidence: Associations between childhood neuroblastomas and self-reported paternal occupational lead exposure were observed in case-control studies
in New York State (OR=2.4, 95% CI 1.2–4.8) (Kerr et al., 2000) and the United States/Canada
(OR=2.6, 95% CI 0.9–7.1) (De Roos et al., 2001). The association in the New York study persisted
when exposure was defined as self-reported plus expert-rated potential for such exposure
(OR=2.2, 95% CI 0.9–5.4) (Kerr et al., 2000).
TCDD Paternal occupational exposure, TCDD, inadequate evidence: A case-control study of
neuroblastoma found an association with self-reported paternal occupational dioxin exposure but
the OR was imprecise (OR=6.9, 95% CI 1.3–68, 7 exposed case fathers) (Kerr et al., 2000).
Pesticides Among five studies of neuroblastoma published by early 1998, two reported associations with parental pesticide exposure through employment in agriculture and one with parental
residential garden pesticide use during childhood (Zahm & Ward, 1998). A review of neuroblastoma
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471
studies published since the Zahm and Ward paper noted that all four found associations with
parental occupational pesticide exposure and one observed an association with residential use of
pesticides including herbicides (Infante-Rivard & Weichenthal, 2006).
Maternal exposure, insecticides, inadequate evidence: A case-control study in New York State
reported an association between neuroblastoma and maternal occupational insecticide exposure
(OR=2.6, 95% CI 1.5–4.5) (Kerr et al., 2000). In the United States/Canada case-control study,
neuroblastoma was associated with prenatal or childhood indoor insecticide use for ants or
cockroaches (use confirmed by both parents, OR=1.8, 95% CI 1.0–3.1) and, among cases age 1 yr
or older, with any indoor insecticide use (OR=1.9, 95% CI 1.1–3.2) (Daniels et al., 2001).
Maternal exposure, unspecified pesticides, inadequate evidence: A large United States/Canada
case-control study reported statistically nonsignificant elevated risks of neuroblastoma related to
maternal occupation in farming (OR=2.2, 95% CI 0.6–8.8, 7 exposed case mothers) and florist/
garden stores (OR=2.4, 95% CI 0.6–9.9, 6 exposed case mothers) (Olshan et al., 1999). A subsequent report of this study revealed that neuroblastomas were associated with maternal application
of garden pesticides (OR=2.2, 95% CI 1.3–3.8) (Daniels et al., 2001). The few studies, limited
sample size and heterogeneity of exposure indices preclude firm conclusions.
Paternal occupational exposure, insecticides, inadequate evidence: A case-control study in
New York State reported an association between neuroblastoma and paternal occupational insecticide exposure (OR=1.8, 95% CI 1.1–3.1) (Kerr et al., 2000).
Paternal occupational exposure, unspecified pesticides, inadequate evidence: In a Norwegian
cohort study, neuroblastoma risk was increased among offspring of farmers (mainly male) who
reported field vegetable farming and purchased pesticides (compared to farm families with neither
trait, RR=2.51, 95% CI 1.03–6.13) (Kristensen et al., 1996a). A large United States/Canada casecontrol study reported that neuroblastoma was associated with certain potentially exposed paternal
occupations (landscaping or grounds keeping, OR=2.3, 95% CI 1.0–5.2) but not farming (OR=0.9,
95% CI 0.4–1.8) (Olshan et al., 1999). A later report of this study indicated that neuroblastomas were
not associated with paternal application of garden pesticides (OR=1.1, 95% CI 0.8–1.5) (Daniels
et al., 2001). The AHS cohort revealed no association between neuroblastoma and paternal occupation as licensed agricultural pesticide applicators (SIR=1.26, 95% CI 0.40–3.89) but there were only
three cases (Flower et al., 2004). The findings of the Norwegian cohort and the large United States/
Canada case-control study are suggestive of an association but further research is needed to clarify
the role of specific pesticides or related groups of pesticides, critical exposure windows and doseresponse relationships. Further follow-up of the AHS cohort would also be valuable.
Childhood exposure, herbicides, inadequate evidence: The United States/Canada case-control
study also reported elevated neuroblastoma risk related to prenatal or childhood herbicide use (use
confirmed by both parents, OR=2.2, 95% CI 1.1–4.3); there was no association among infants
below age 1 yr (Daniels et al., 2001).
Childhood exposure, insecticides, inadequate evidence: A recent report of the United States/
Canada case-control study noted elevated neuroblastoma risk among children age 1 or older
related to prenatal or childhood insecticide use indoors (OR=1.9, 95% CI 1.1–3.2) and in gardens
(OR=1.7, 95% CI 0.8–3.6) (Daniels et al., 2001).
Solvents Parental occupational exposure, unspecified solvent, inadequate evidence: A large
United States–Canada case-control study of neuroblastoma found associations with likely occupational solvent exposure during the 2 yr before the child’s birth among fathers (OR=1.5, 95% CI
1.0–2.1) but not mothers (OR=1.2, 95% CI 0.7–2.1) (De Roos et al., 2001). For paternal solvent
exposure, associations were stronger (and statistically significant) for lacquer thinner, mineral spirits,
and turpentine (ORs of 1.9 to 3.5).
Summary There is limited epidemiologic evidence that childhood neuroblastoma is associated
with paternal occupational lead exposure.
Wilms’s Tumor
Pesticides A review of the six studies of Wilms’ tumor published by early 1998 noted that
none of the three that assessed postnatal parental pesticide exposure (occupational use or residential
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D. T. WIGLE ET AL.
gardening) found an association (Zahm & Ward, 1998). The only study that examined household
extermination (mainly insecticide use) observed an association and the two studies that focused on
parental occupational exposures before birth both found associations. A review of the two studies
published since Zahm and Ward’s report noted that both observed statistically nonsignificant
increased risks related to parental occupational pesticide exposure based on only two or three
exposed case parents (Infante-Rivard & Weichenthal, 2006).
Maternal exposure, herbicides, inadequate evidence: In a United States/Canada case-control
study, Wilms’s tumor was not associated with residential herbicide use at any time during pregnancy or childhood (OR=1.0, 95% CI 0.7–1.4) (Cooney et al., 2007).
Maternal exposure, insecticides, inadequate evidence: A German case-control study reported a
statistically nonsignificant elevated risk of Wilms’s tumor related to prenatal or childhood indoor
residential insecticide use (≥1 vs. <1×/yr, OR=1.27, 95% CI 0.78–2.08) (Schuz et al., 2001). In the
United States/Canada case-control study, Wilms’s tumor was associated with indoor residential
insecticide use at any time during pregnancy or childhood (OR=1.4, 95% CI 1.0–1.8) (Cooney
et al., 2007).
Maternal exposure, unspecified pesticides, inadequate evidence: A case-control study in Brazil,
a country with relatively high Wilms’s tumor incidence rates, revealed a dose-response relationship
between this tumor and frequency of preconceptual/prenatal occupational exposure to agricultural
pesticides (≥10 vs. 0 times, OR=129, 95% CI 6.4–2570, 6 exposed case and 1 exposed control
mothers, p-trend=.03) (Sharpe et al., 1995). There were statistically nonsignificant elevated risks of
Wilms’s tumor in case-control studies in Germany (maternal occupational pesticide exposure,
OR=2.52, 95% CI 0.50–12.6, 2 exposed case mothers) (Schuz et al., 2001) and the United States
(prenatal residential or occupational pesticide exposure, OR=1.32, 95% CI 0.83–2.09) (Tsai et al.,
2006). In the United States/Canada case-control study, Wilms’s tumor was associated with indoor
residential use of any pesticide at any time during pregnancy or childhood (OR=1.3, 95% CI 1.0–
1.7) (Cooney et al., 2007). Larger studies are needed to clarify the role of specific pesticides or
related groups of pesticides, critical exposure windows and dose-response relationships.
Paternal occupational exposure, unspecified pesticides, inadequate evidence: The Brazil study
revealed a dose-response relationship between Wilms’ tumor and frequency of preconceptual
paternal occupational exposure to agricultural pesticides (≥10 vs. 0 times, OR=3.2, 95% CI 1.2–
9.0, p-trend=.02) (Sharpe et al., 1995). A Norwegian retrospective cohort study reported elevated
risks of Wilms’ tumor among offspring of farmers (mainly male) who reported ownership of pesticide spraying equipment (OR=2.54, 95% CI 0.98–6.58) and those with orchards or greenhouses
and pesticide spraying equipment (OR=8.87, 95% CI 2.67–29.5) (Kristensen et al., 1996a).
Although this study lacked detailed pesticide use information, it had the advantage of not relying on
potentially biased parental recall of exposures. A mortality study in England found no association
between Wilms’s tumor and paternal occupation in agriculture (PMR=0.88, 95% CI 0.20–3.84)
(Pearce & Parker 2000). The AHS cohort revealed a statistically nonsignificant elevated risk of
Wilms’s tumor related to paternal occupation as licensed agricultural pesticide applicators
(SIR=1.56, 95% CI 0.50–4.84, only 3 cases) (Flower et al., 2004). The two studies with negative
findings are not compelling because one was a proportional mortality study limited to information
on death records and the other was a cohort study with only three observed cases.
Childhood exposure, herbicides, inadequate evidence: In the U.S./Canada case-control study,
Wilms’s tumor was not associated with residential herbicide use at any time during pregnancy or
childhood (OR=1.0, 95% CI 0.7–1.4) (Cooney et al., 2007).
Childhood exposure, insecticides, inadequate evidence: The German study reported a statistically nonsignificant elevated risk of Wilms’ tumor related to prenatal or childhood indoor residential
insecticide use (≥1 vs. <1×/yr, OR=1.27, 95% CI 0.78–2.08) (Schuz et al., 2001). In a United
States/Canada case-control study, Wilms’ tumor was associated with indoor residential insecticide
use at any time during pregnancy or childhood (OR=1.4, 95% CI 1.0–1.8) (Cooney et al., 2007).
Childhood exposure, unspecified pesticides, inadequate evidence: The German study reported
no association between Wilms’s tumor and childhood residence on farms (OR=0.8, 95% CI 0.3–2.3),
residential garden pesticide use (OR=0.8, 95% CI 0.4–1.5) or childhood paternal occupational
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pesticide exposure (OR=0.97, 95% CI 0.39–2.37) (Schuz et al., 2001). In the United States/Canada
case-control study, Wilms’s tumor was associated with indoor residential use of any pesticide at any
time during pregnancy or childhood (OR=1.3, 95% CI 1.0–1.7) (Cooney et al., 2007).
Other toxicants Paternal exposure, inadequate evidence: Reviewers found inadequate
evidence for an association between Wilms’s tumor and paternal occupational lead exposure (Colt &
Blair, 1998).
Summary There was inadequate epidemiologic evidence for an association between Wilms’
tumor and any of the environmental contaminants examined here.
Bone Tumors
Pesticides A review of the three studies of Ewing’s sarcoma published by early 1998 noted that
all found elevated risks related to parental pesticide exposure (statistically significant in two studies)
(Zahm & Ward, 1998). Reviewers noted that the only study published since the Zahm and Ward
review found increased Ewing’s sarcoma risk related to agricultural employment of at least 1 parent
(Infante-Rivard & Weichenthal, 2006).
Maternal exposure, unspecified pesticides, inadequate evidence: Maternal occupation in farming
was associated with increased risk of childhood bone cancer in Ontario (2.7, 95% CI 0.8–9.3),
especially among women with at least 5 yr exposure (OR=2.9, 95% CI 0.5–15.3) and the subgroup
of Ewing’s sarcoma (OR=7.8, 95% CI 1.9–32) (Hum et al., 1998). An Australian case-control study
reported a statistically nonsignificant elevated risk of Ewing’s sarcoma among children and young
adults in relation to periconceptual maternal occupation in farming (OR=2.8, 95% CI 0.5–15.8)
and ever-handling of pesticides (OR=2.3, 95% CI 0.5–12) (Valery et al., 2002).
Paternal exposure, unspecified pesticides, limited evidence: Ewing’s sarcoma was associated
with paternal occupational pesticide exposure in a California case-control study (OR = 8.8, 95%
CI 1.8–42.7) (Holly et al., 1992). The Ontario study showed that osteosarcoma risk was elevated
in relation to paternal occupation in farming (OR = 2.1, 95% CI 0.8–5.7) and bone cancer risk
(any histologic type) was increased in relation to paternal work in farming for at least 5 yr
(OR = 2.4, 95% CI 0.9–6.4) (Hum et al., 1998). The Australian case-control study reported statistically nonsignificant elevated risks of Ewing’s sarcoma related to periconceptual paternal occupation in farming (OR = 3.5, 95% CI 1.0–11.9) and ever-handling of pesticides (OR = 2.0, 95% CI
0.8–4.9) (Valery et al., 2002). In the AHS cohort, there was a statistically nonsignificant increased
risk of bone cancer among offspring of licensed pesticide applicators (SIR = 2.19, 95% CI 0.82–
5.84, based on 4 cases) (Flower et al., 2004). A U.S. case-control study of Ewing’s sarcoma
observed no association with likely prenatal paternal pesticide exposure (OR = 0.7, 95% CI 0.2–
2.9) (Moore et al., 2005).
Childhood exposure, insecticides, inadequate evidence: In a California case-control study,
Ewing’s sarcoma was not associated with childhood indoor insecticide use (OR=0.6, 95% CI 0.3–
1.2) or residence on or next to a farm (OR=1.0, 95% CI 0.3–4.0) (OR=0.6, 95% CI 0.3–1.2)
(Holly et al., 1992). A U.S. case-control study of Ewing’s sarcoma observed an association between
childhood residential indoor insecticide use and Ewing’s sarcoma among boys (OR=3.0, 95% CI
1.1–8.1) but not girls (OR=1.1, 95% CI 0.4–3.2) (Moore et al., 2005).
Childhood exposure, unspecified pesticides, inadequate evidence: A U.S. case-control study
reported statistically nonsignificant elevated risks of Ewing’s sarcoma among girls who had ever lived
on farms (OR=6.4, 95% CI 0.7–58.4) but not among boys (OR=0.9, 95% CI 0.4–2.2) and but not
in relation to postnatal parental occupational pesticide exposure (OR=0.9, 95% CI 0.2–3.7)
(Moore et al., 2005).
Germ-Cell Tumors Germ-cell tumors arise from male or female germ cells and can occur at
almost any anatomic site and be benign or malignant.
Pesticides Maternal exposure, herbicides, insecticides, inadequate evidence: A U.S. case-control
study revealed no association between germ-cell tumors and prenatal residential use of indoor
insecticides (used ≥2 vs. 0 types of insecticides, OR=1.2, 95% CI 0.8–1.6, p-trend=.48) and a
statistically nonsignificant, modestly elevated risk related to prenatal residential herbicide use (ever
vs. never, OR=1.3, 95% CI 0.9–1.7) (Chen et al., 2006).
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D. T. WIGLE ET AL.
Maternal exposure, unspecified pesticides, inadequate evidence: A United States/Canada casecontrol reported elevated risks of germ-cell tumors (ovarian, testicular, nongonadal) related to
maternal occupational or residential pesticide exposure (OR=2.4, 95% CI 0.9–6.9) (Shu et al.,
1995). In the AHS cohort study, children of licensed pesticide applicator-farmers (almost entirely
men) had an increased risk of germ-cell tumors (OR=2.34, 95% CI 0.88–6.24; 5 exposed case
fathers); there was potential prenatal pesticide exposure because the women lived on farms and
58% of them reported mixing or applying pesticides (Flower et al., 2004). In another U.S. case-control
study, prenatal pesticide exposure at work was not associated with germ-cell tumors (>50th
percentile of cumulative exposure vs. unexposed, OR=0.9, 95% CI 0.5–1.7, p-trend=0.93) but
was related to an elevated risk of dysgerminoma, a specific histological type of germ-cell tumor
(ever exposed vs. never, OR=1.9, 95% CI 0.9–4.2) (Chen et al., 2005b). Three studies observed
elevated risks of germ-cell tumors related to prenatal occupational or residential pesticide exposure.
However, none of the relationships were statistically significant, one was limited to a histologic
subtype and the role of maternal versus paternal exposure in the AHS cohort is not clear.
Paternal exposure, herbicides, insecticides, inadequate evidence: A U.S. case-control study
revealed no association between germ-cell tumors and preconceptual/perinatal paternal residential
use of herbicides (ever vs. never, OR=1.0, 95% CI 0.7–1.3) or insecticides (used ≥2 vs. 0 types of
insecticides, OR=1.1, 95% CI 0.8–1.6, p-trend=.64) (Chen et al., 2006).
Paternal exposure, unspecified pesticides, inadequate evidence: The United States/Canada
case-control reported elevated risks of germ-cell tumors (ovarian, testicular, nongonadal) related to
paternal occupational pesticide exposure (OR=1.8, 95% CI 0.7–5.0) (Shu et al., 1995). In the AHS
cohort study, children of licensed pesticide applicator-farmers (almost entirely men) had an
increased risk of germ-cell tumors (OR=2.34, 95% CI 0.88–6.24; 5 exposed case fathers) (Flower
et al., 2004). In a U.S. case-control study, preconceptual paternal pesticide exposure at work was
not associated with germ-cell tumors (>50th percentile of cumulative exposure vs. unexposed,
OR=0.9, 95% CI 0.6–1.3, p-trend=0.49) or any histological subtype (Chen et al., 2005b). Two of
the three studies observed statistically nonsignificant elevated risks of germ-cell tumors related to
self-reported paternal occupational pesticide exposure.
Childhood exposure, insecticides, inadequate evidence: A U.S. case-control study revealed no
association between germ-cell tumors and childhood residential indoor insecticide use (used ≥2 vs.
0 types of insecticides, OR=1.0, 95% CI 0.7–1.5, p-trend=.99) (Chen et al., 2006).
Childhood exposure, unspecified pesticides, inadequate evidence: In the AHS cohort study,
children of licensed pesticide applicator-farmers (almost entirely men) had a statistically nonsignificant increased risk of germ-cell tumors (OR=2.34, 95% CI 0.88–6.24; 5 exposed case fathers)
(Flower et al., 2004). Given their residence on farms known to use pesticides, and evidence that
children of pesticide applicators have substantially higher urinary organophosphate insecticide
metabolites compared to other children (Fenske et al., 2005), there clearly was the potential for
childhood pesticide exposure.
Tobacco smoke Maternal active smoking, inadequate evidence: In a large United States-Canada
case-control study, childhood germ-cell tumors were not associated with prenatal active smoking
(≥20 cigarettes/d, OR=1.1, 95% CI 0.6–2.1) or ETS exposure (OR=1.0, 95% CI 0.6–1.3) (Chen
et al., 2005a).
Paternal smoking, inadequate evidence: The United States–Canada case-control study revealed
no association between childhood germ-cell tumors and paternal active smoking (smoked 16 or
more years before pregnancy, OR=1.3, 95% CI 0.7–2.2, p-trend=.13; smoked 16 or more years
before pregnancy, OR=1.3, 95% CI 0.6–2.7, p-trend=.21) (Chen et al., 2005a).
Eye Tumors Most childhood eye tumors are retinoblastomas. There was an increased risk of
adult eye tumor death among farmers licensed to apply pesticides in Italy (SMR=2.38, 95% CI
0.65–6.09) (Torchio et al., 1994), but most adult eye tumors are melanomas and the relevance to
childhood eye tumors, apart from demonstrating the potential vulnerability of the eye to environmental carcinogens, is not clear.
Pesticides Maternal exposure, unspecified pesticides, inadequate evidence: A U.S. case-control
study reported an association between nonheritable retinoblastoma and maternal grandparent
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
475
occupation in farming (OR=10.0, 95% CI 1.4–433; based on 10 case-only and 1 control-only
exposed matched pairs) (Bunin et al., 1990).
Paternal exposure, unspecified pesticides, inadequate evidence: In a Norwegian retrospective
cohort study, childhood eye tumors were associated with parental (mainly paternal) pesticide
purchases for field vegetable farming (compared to no field vegetable farming and no pesticide
purchases, RR=3.17, 95% CI 0.93–10.9) (Kristensen et al., 1996a). The AHS cohort revealed a
statistically nonsignificant elevated retinoblastoma risk among children of licensed agricultural pesticide applicators (SIR=1.63, 95% CI 0.41–6.53, only 2 cases) (Flower et al., 2004).
Soft-Tissue Sarcoma
Pesticides Soft tissue sarcomas comprise relatively rare and diverse histological subtypes of
largely unknown causation (the only proven cause of adult soft-tissue sarcomas is high-dose radiation). Among adults, soft-tissue sarcomas have been linked to occupational exposure to herbicides
or multiple pesticides in several studies conducted in various countries (Briggs et al., 2003; Hoar
Zahm et al., 1988) (Hansen et al., 1992; Hoppin et al., 1999; Kogevinas et al., 1997; Lynge 1998;
Vineis et al., 1987), but other studies found no association (Fleming et al., 1999; Johnson et al.,
1990; Pahwa et al., 2003; Wiklund et al., 1988).
Maternal exposure, herbicides, unspecified pesticides, inadequate evidence: A case-control
study in Denver reported that childhood soft-tissue sarcoma was not associated with prenatal
residential herbicide use (OR=0.8, 95% CI 0.5–1.3) (Leiss & Savitz, 1995). In a relatively large
case-control study in Germany, childhood soft-tissue sarcomas were associated with maternal occupational pesticide exposure (result stated without supporting data in paper) (Meinert et al., 2000).
Paternal exposure, unspecified pesticides, inadequate evidence: The AHS cohort revealed no
association between soft-tissue sarcoma and paternal occupation as licensed agricultural pesticide
applicators (SIR=1.17, 95% CI 0.38–3.62, only 3 cases) (Flower et al., 2004).
Childhood exposure, herbicides, inadequate evidence: A case-control study in Denver
reported that childhood soft-tissue sarcoma was associated with postnatal lawn herbicide use
(OR=4.1, 95% CI 1.0–16.0) (Leiss & Savitz, 1995).
Adult Cancers
This section describes adult cancer risks linked to prenatal or childhood environmental exposures, with level of epidemiologic evidence summarized in Table 7. About 130,000 new cancer
cases annually occur among persons age 15–49 in the United States (Wu et al., 2005). Cancer incidence rates for this age range increased during 1969–1996 in Canada for lung (women), testicular
and thyroid cancers, melanoma, and NHL (Marrett et al., 2002).
Testicular Cancer Reviewers concluded that available evidence from human and animal
studies supports a hypothesis that testicular cancer is associated with a prenatal exposure to exogenous
estrogens (Storgaard et al., 2006). However, there have been few epidemiologic studies of testicular
cancer and prenatal exposure to environmental toxicants.
PCBS Maternal exposure, inadequate evidence: A recent Swedish case-control study
observed an association between testicular cancer and maternal plasma PCB levels (>median vs.
≤median, OR=3.8, 95% CI 1.4–10.0) (Hardell et al., 2003). Further investigation revealed an
TABLE 7. Role of Early-Life Exposure to Environmental Toxicants in Adult Cancer
Toxicant
Exposure
Arsenic
Prenatal
Childhood
Prenatal
Prenatal
Prenatal
Paternal
Childhood
Childhood
PCBs
Organochlorine pesticides
Unspecified pesticides
Environmental tobacco smoke
Drinking-water nitrate
Testicular
Breast
Lung
Other
I
I
I
I
I
I
L
I
S
Pancreas—I Chronic lymphocytic leukemia—I
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D. T. WIGLE ET AL.
association between testicular cancer and maternal plasma PCB-TEQ (>median vs. ≤median,
OR=3.3, 95% CI 1.3–8.4) (Hardell et al., 2006).
Pesticides Maternal exposure, organochlorine pesticides, inadequate evidence: A small casecontrol study reported that testicular cancer was marginally associated with the subjects’ plasma
HCB (≥median vs. <median plasma levels, OR=1.7, 95% CI 0.8–3.6) and DDE levels (OR=1.7,
95% CI 0.8–3.7) and strongly associated with their mothers’ plasma HCB levels (OR=4.4, 95% CI
1.7–12) (Hardell et al., 2003). This study found no association with maternal plasma DDE
(OR=1.3, 95% CI 0.5–3.0) and a statistically nonsignificant relationship with sum of chlordanes
(OR=1.9, 95% CI 0.7–5.0).
Maternal occupational exposure, unspecified pesticides, inadequate evidence: Testicular
cancer in Denmark was not associated with prenatal employment in agriculture (OR=1.2, 95% CI
0.6–2.7) (Moller, 1997).
Paternal occupational exposure, unspecified pesticides, inadequate evidence: In a retrospective
cohort study of Norwegian farm families, there was a slightly elevated testicular cancer risk among
offspring (relative to national rates, SIR=1.24, 95% CI 1.01–1.52), but risk was not related to
expenditures on farm pesticides (any expenditure vs. none, OR=0.89, 95% CI 0.60–1.32)
(Kristensen et al., 1996a).
Drinking water nitrate Childhood exposure, inadequate evidence: In a Norwegian retrospective cohort study, testicular cancer among offspring in farm families was associated with parental farm
use of high nitrogen fertilizers (OR=2.0, 95% CI 1.5–2.6) (Kristensen et al., 1996b). A populationbased case-control study in Denmark reported an association between testicular cancer and a
history of having lived for most of childhood in 3 counties with ground water nitrate levels exceeding
25 mg/L (OR=1.4, 95% CI 1.1–1.8) (Moller, 1997). However, the excess risk was largely limited to
men who grew up in urbanized regions of the three counties served by low-nitrate communal water
supplies; thus, nitrate per se probably was not responsible for the observed association.
Summary There is inadequate epidemiologic evidence for a role of prenatal or childhood
exposure to environmental toxicants in adult testicular cancer.
Breast Cancer
Environmental tobacco smoke Childhood exposure, evidence: An expert group convened by
the State of California found sufficient evidence of a causal association between breast cancer and
ETS exposure, particularly among premenopausal women and those exposed early in life (California
Environmental Protection Agency, 2005). The California report also noted that ETS contains these
carcinogens known to cause breast cancer in experimental animals: benzene, dibenz[a,h]anthracene,
four dibenzopyrenes, two nitrosamines, eight aliphatic compounds, and three arylamines/nitroarenes.
The U.S. Surgeon General reviewed findings from 7 prospective cohort and 14 case-control studies
of breast cancer and ETS exposure and concluded that the evidence was suggestive of a causal relationship but not conclusive (U.S. Department of Health and Human Services, 2006). The strongest
associations were generally based on case-control studies and premenopausal breast cancer. The
Surgeon General noted that there was inconsistent evidence for increased breast cancer risk related
to ETS exposure during childhood or adolescence.
A pooled analysis of data from 53 studies that compared ever and never smokers showed no
association between breast cancer and active smoking (pooled OR=1.03, 95% CI 0.98–1.07)
(Hamajima et al., 2002). However, smoking has effects that may increase or decrease breast cancer
risk. For instance, active smokers have lower urinary estrogen levels and increased estradiol
2-hydroxylation compared to nonsmokers, characteristics associated with reduced breast cancer
risk (U.S. Department of Health and Human Services, 2006). However, tobacco smoke mutagens
and aromatic DNA adducts are detectable in breast fluid or tissues. In active smokers, but not
ETS-exposed nonsmokers, the anti-estrogenic effects of smoking may offset any increased breastcancer
risk from tobacco smoke carcinogens (U.S. Department of Health and Human Services, 2006).
Summary There is limited epidemiologic evidence for an association between adult breast
cancer and lifetime ETS exposure; thus, childhood ETS exposure may contribute to lifetime breast
cancer risk in women.
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CHILD HEALTH AND ENVIRONMENTAL CONTAMINANTS
477
Lung Cancer
Arsenic Maternal or childhood exposure, inadequate evidence: An ecologic study in Chile
noted increased lung cancer death rates among persons who were likely exposed to high arsenic
levels in drinking water prenatally (SMR=7.0, 95% CI 5.4–8.9) or during early childhood
(SMR=6.1, 95% CI 3.5–9.9) (Smith et al., 2006).
Environmental tobacco smoke Childhood exposure, sufficient evidence: The U.S. Surgeon
General reviewed 8 cohort and over 40 case-control studies and concluded that there is sufficient
evidence that ETS causes adult lung cancer (U.S. Department of Health and Human Services,
2006). Thus, childhood ETS exposure contributes to lifetime exposure and risk of adult lung cancer.
Summary There is sufficient evidence that childhood ETS exposure contributes to lifetime ETS
exposure and increased lung cancer risk.
Other Adult Cancers
Environmental tobacco smoke Childhood exposure, inadequate evidence: A large Canadian
case-control study observed a slightly elevated risk of pancreatic cancer among persons exposed to
ETS during both childhood and adulthood (OR=1.21, 95% CI 0.60–2.44) (Villeneuve et al., 2004).
In a large case-control study restricted to nonsmokers, lifetime duration of residential and occupational ETS exposure was associated with adult chronic lymphocytic leukemia (≥53 vs. 0 yr,
OR=2.18, 95% CI 1.10–4.32, p-trend=0.001) and to a lesser degree with acute myelogenous
leukemia (≥53 vs. 0 yr, OR= 1.74, 95% CI 0.83–3.66, p-trend=0.27) (Kasim et al., 2005).
Summary There is inadequate evidence for associations between other adult cancers and
prenatal or childhood exposure to environmental chemical contaminants.
Other Health Outcomes
Table 8 summarizes the level of epidemiologic evidence for associations between environmental contaminants and postnatal growth and pubertal development, and Table 9 addresses sudden infant death
syndrome (SIDS), tooth abnormalities, chloracne, renal tubular damage, and chromosomal abnormalities.
Postnatal Growth in Height
Lead Maternal exposure, inadequate evidence: In the Cleveland birth cohort, height at age 4
was not associated with average maternal and cord blood lead levels (change in height (% of 1 SD)
per unit change in blood lead (µg/dl), β=–0.26 ± 3.37, p=.94) (Greene & Ernhart, 1991).
TABLE 8. Role of Environmental Toxicants in Child Growth and Pubertal Development
Toxicant
Exposure
Lead
Prenatal
Childhood
Methylmercury Prenatal
Lactational
PCBs
Prenatal
PBBs
DDT/DDE
Environmental
tobacco
smoke
Phthalates
Lactational or
childhood
Prenatal
Prenatal
Lactational or
childhood
Prenatal
Childhood
Childhood
Delayed
pubic hair
Delayed breast
Postnatal
Delayed
growth in height menarche development, girls development
I
L
I
I
High-level: L
Low-level: I
I
I
I
I
L
Delayed male
external genital
development
Delayed male
pubic hair
development
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
I
Early breast
I
development: I
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D. T. WIGLE ET AL.
TABLE 9. Role of Environmental Toxicants in other Child Health Outcomes
Toxicant
Exposure
Renal tubular damage
Lead
Mercury
Childhood
Childhood
S
Organic Hg—L
Elemental Hg—I
Arsenic
Prenatal
Childhood
Childhood
Prenatal
Cadmium
PCBs, PCDFs, related
toxicants
Chromosomal
abnormalities
Dental caries –L
I
L
I
Allergies/low-level
PCBs—I
Chloracne/high-level
PCBs—S
Allergies/low-level
PCBs—I
Chloracne/high-level
TCDD—S
Allergies—I
SIDS—S
Childhood
TCDD
Childhood
Unspecified pesticides
Active smoking
Environmental tobacco
smoke
Outdoor air pollution
Prenatal
Prenatal
Prenatal
Infancy
Prenatal
Infancy
Prenatal
Prenatal
Drinking water DBPs
Hazardous waste
disposal sites
Incinerators
Unspecified solvents
Other diseases
Prenatal
Prenatal
Paternal
I
I
I
SIDS—S
SIDS—L
I
I
I
I
I
Note. TCDD, 2,3,7,8-tetrachlorodibenzo-p-dioxin.
Childhood exposure, limited evidence: In the Cincinnati birth cohort, there was an inverse
dose-response relationship between infant blood lead at age 3 months and growth in stature from
age 3 to 15 months (regression slope b = −0.015 cm per µg/dL, p = .013). (Shukla et al., 1989).
Further follow-up revealed an inverse association between stature at age 33 months and the interaction term between average blood lead from age 3 to 15 months times that for age 18 to 33 months
(β = −1.81 ± 0.80 cm, p = .025) (Shukla et al., 1991). In the Cleveland birth cohort, height at age
4 yr was inversely associated with blood lead at age 6 mo (change in height (% of 1 SD) per unit
change in blood lead (µg/dl), β = –3.91 ± 2.07, p = .06) but not current blood lead (β = 1.62 ±
1.40, p = .25) (Greene & Ernhart, 1991). Three large cross-sectional studies based on NHANES II
and III found inverse dose-response relationships between height and current blood lead levels
extending below 10 µg/dl with no evidence of a threshold. Among children age 1–7, height was
inversely associated with current blood lead levels in NHANES II (change in height per unit
change in current blood lead (µg/dL), β = −0.12 ± 0.0005 cm, p < 0.0001) (Schwartz et al.,
1986). Similarly, height was inversely associated with current blood lead level among all children
in NHANES III (boys and girls age 1–7, β = −0.157 ± 0.032 cm, p < 0.0001) (Ballew et al., 1999)
and among the subset of girls age 8–18 (difference in height, current blood lead ≥ 3 vs. <1.0 µg/dL,
regression slope r = −0.51 cm, p < 0.001) (Selevan et al., 2003). A small birth cohort study in
Massachusetts revealed no association between height at age 6–8 and log tooth dentin lead levels
(β = −0.9 ± 1.1 (SE), p > 0.05) (Kim et al., 1995). A cross-sectional study of Italian children age
11–13 revealed an inverse association between height and log current blood lead levels (boys, β
= −27.4 ± 11.5 (SE) cm, p = 0.02) (Vivoli et al., 1993). Similarly, a cross-sectional study of Greek
children age 6–9 reported an inverse association between height and current blood lead (µg/dL)
(β = −0.086 ± 0.037 (SE) cm, p = 0.02) (Kafourou et al., 1997).
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Methylmercury Maternal exposure, inadequate evidence: A small birth cohort study in the
Faroe Islands reported an inverse association between height at age 18 mo and cord blood mercury
(change in height per log cord blood mercury increment, β=–0.88 cm, 95% CI –1.85 to 0.09); this
association persisted at age 42 mo but was not statistically significant (–0.97 cm, 95% CI –2.42 to
0.49) (Grandjean et al., 2003).
Lactational exposure, inadequate evidence: The Faroe Islands cohort observed inverse associations between cumulative lactational mercury exposure (cord blood mercury times breast-feeding
duration in weeks) and height at age 18 mo (change in height per log lactational mercury
increment, β=–0.73 cm, 95% CI –1.30 to –0.12) but not height at age 42 mo (β=–0.34 cm, 95%
CI –1.16 to 0.48) (Grandjean et al., 2003).
PCBs and related compounds High-level prenatal PCB exposure, limited evidence: Compard to
unexposed children, Yucheng children up to age 8 were 3% shorter (95% CI –4 to –1) (Rogan et al.,
1988). Further follow-up revealed a persistent height deficit up at ages 6–13 yr (exposed vs. unexposed, difference –3.1 cm, p < .01) (Guo et al., 1994). Reviewers concluded that growth in height
during childhood was reduced among offspring of women prenatally exposed to cooking oil highly
contaminated by PCBs, PCDFs, and related toxicants (Guo et al., 1995; Longnecker et al., 1997).
Low-level maternal exposure, inadequate evidence: The Dutch birth cohort study reported that
cord blood PCB was inversely associated with growth in height from birth to age 3 mo (change in
height per unit change in natural log PCB, β=–0.28 ± 0.12 cm, p=.03) but not with height changes
during mo 3–7, 7–18, or 18–42 (Patandin et al., 1998). A Michigan birth cohort study found no association between height at age 4 yr and prenatal PCB exposure (stated without supporting data) (Jacobson
et al., 1990b). The North Carolina birth cohort study observed no association between prenatal PCB
exposure and height at age 12–14 yr in girls or boys (average height vs. increasing prenatal maternal
PCB level, p-trend was 0.75 for girls and 0.24 for boys) (Gladen et al., 2000). In another Michigan
birth cohort, there was no association between prenatal serum PCB and height of daughters age 5–24
(PCB ≥9.0 vs. ≤5.0 µg/L, β=0.2 inches, 95% CI –0.8 to 1.3) (Blanck et al., 2002). A birth cohort study
in New York City revealed no association between height at intervals up to age 17 and 3rd trimester
maternal serum PCB concentrations (unit change in natural log of height at each age per unit change
in natural log of maternal serum PCBs, girls, –0.6 cm, 95% CI –3.2 to 2.0; boys, 0.5 cm, 95% CI –1.4
to 2.3) (Lamb et al., 2006). A retrospective cohort study of Swedish fishermen’s wives revealed no
association between maternal plasma PCB–153 concentrations and child height at age 4 or 7 (normal
birth weight children, mean difference in height at age 7, maternal plasma PCB–153 ≥250 vs. <250
ng/g lipid, β=–0.31, 95% CI –2.05 to 1.43) (Rylander et al., 2007).
Lactational or childhood PCB exposure, inadequate evidence: A Michigan birth cohort study
found no association between height at age 4 yr and lactational PCB exposure (stated without
supporting data) (Jacobson et al., 1990b). The Dutch birth cohort study reported that growth in
height from 3 to 7 mo was inversely related to cumulative lactational dioxin-TEQ based on cord
blood PCDD/PCB concentrations and breast-feeding duration (weeks) (change in height per unit
change in cumulative exposure, β = –0.21 cm, p = .04); changes in height from birth to age 3 mo,
from 7 to 18 mo, and from 18 to 42 mo were not associated with lactational PCDD/PCB-TEQ
exposure (Patandin et al., 1998). The North Carolina birth cohort study observed no association
between cumulative lactational PCB exposure and height at age 12–14 (average height vs.
increasing lactational PCB intake, p-trend was .13 for girls and 0.92 for boys) (Gladen et al.,
2000). The North Carolina birth cohort study observed no association between lactational PCB
exposure and height at age 12–14 (Gladen et al., 2000). A German cohort study found no association between growth in height from age 7–8 yr to 10–11 yr and baseline blood PCB levels (result
stated without supporting data) (Karmaus et al., 2002). In a small Faroe Islands birth cohort,
growth in height from birth to age 42 mo was inversely associated with lactational PCB exposure
(change in height per doubling of serum PCB at age 54 mo, –0.63 cm, 95% CI –1.12 to–0.13)
(Grandjean et al., 2003).
PBBs Maternal exposure, inadequate evidence: A Michigan birth cohort study reported no
association between prenatal serum PBB and height of daughters at age 5–24 yr (change in height,
PCB ≥7.0 vs. ≤1.0 µg/L, β=0.61 inches, 95% CI –0.50 to 1.7) (Blanck et al., 2002).
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Pesticides Prenatal or lactational DDT/DDE exposure, inadequate evidence: A birth cohort
study in North Carolina revealed a favourable association between height of adolescent boys and
increasing breast milk DDE concentration categories (p-trend=.05) (Gladen et al., 2000). This
study found no association between adolescent height of girls and breast milk DDE or between
height of either gender and lactational DDE exposure; there was no adjustment for maternal height
or prenatal smoking. Among girls in a German birth cohort, heights (adjusted for birth weight,
breast-feeding duration, prenatal smoking, and other potential confounders) at ages 4–6 wk, 3–4
mo, 6–7 mo, 10–12 mo, 21–24 mo, 43–48 mo, 8 yr, and 9 yr were inversely associated with blood
DDE quartiles measured at age 7–8 yr; the difference for girls at age 10 yr was not significant and
there was only one significant difference for boys (Karmaus et al., 2002). In the U.S. Collaborative
Perinatal Project (pregnant women recruited during 1959–1966), follow-up of a subsample of sons
at age 20 revealed no association between height and quintiles of prenatal serum DDE (Gladen
et al., 2004).
Environmental tobacco smoke Maternal or childhood exposure, ETS, inadequate evidence:
A longitudinal cohort study in California found no association between height at age 5 yr and prenatal ETS exposure of nonsmoking mothers (confirmed by serum cotinine levels at 1st prenatal visit)
(Eskenazi & Bergmann, 1995). The California expert panel review found inadequate evidence of a
causal association between childhood growth in height and prenatal or childhood ETS exposure
and noted that this relationship has been much less studied than postnatal growth and prenatal
active smoking (California Environmental Protection Agency, 2005).
Summary There is limited epidemiologic evidence that high-level prenatal exposure to PCBs,
PCDFs and related toxicants and childhood lead exposure can reduce childhood growth in height.
Adolescent Reproductive Development: Age at Menarche
Lead Childhood exposure, limited evidence: A large cross-sectional study of age 8–16 girls
based on NHANES III found associations between delayed onset of menarche and current blood
lead among non-Hispanic white girls (likelihood of reaching menarche, blood lead ≥ 3 vs. < 1 µg/dL,
OR = 0.74, 95% CI 0.55–1.002) and African-American girls (OR = 0.78, 95% CI 0.63–0.98) but
not among Mexican-American girls OR = 0.90, 95% CI 0.73–1.11) (Selevan et al., 2003). These
estimates were based on current blood lead levels over the range 0.7–22 µg/dL and were adjusted
for age, family size, urban residence, poverty and body mass index. An independent analysis of
NHANES III data found an inverse association between likelihood of having attained menarche and
current blood lead levels among girls of all ethnicities/races combined (OR = 0.19, 95% CI 0.08–0.43)
(Wu et al., 2003).
PCBs Prenatal or lactational exposure, inadequate evidence: The North Carolina birth cohort
study reported no association between age at menarche and indices of transplacental or lactational PCB exposure (average age at menarche, highest vs. lowest maternal or cord serum PCB
level, 12.6 vs. 12.7, p-trend over 4 exposure categories was 0.46) or cumulative lactational PCB
exposure (average age at menarche, highest vs. lowest cumulative lactational PCB dose, 12.8 vs.
12.9, p-trend over 4 cumulated exposure categories was 0.69) (Gladen et al., 2000). The Michigan
cohort observed no association between age at menarche and maternal serum PCB levels (change
in age at menarche per unit change in maternal serum PCB, β=–0.01 ± 0.04 yr, p=.76) (Vasiliu
et al., 2004).
PBBs Prenatal exposure, inadequate evidence: A Michigan birth cohort study observed an
association between maternal serum PBB levels and likelihood of being post-menarche among
breast-fed (≥7 vs. ≤1 µg/L, OR=3.4, 95% CI 1.3–9.0) but not formula-fed daughters (OR=0.8,
95% CI 0.3–1.8) (Blanck et al., 2000).
Age at Reproductive System Maturation: Pubic Hair (Girls)
Lead Childhood exposure, limited evidence: A large cross-sectional study of age 8–16 girls
based on NHANES III found associations between delayed pubic hair development and current
blood lead level among African-American girls (likelihood of reaching a successive stage of pubic hair
development, blood lead ≥ 3 vs. <1 µg/dL, OR = 0.64, 95% CI 0.41–0.97) and Mexican American
girls (OR = 0.76, 95% CI 0.63–0.91) but not among non-Hispanic white girls (OR = 0.82, 95%
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CI 0.47–1.42) (Selevan et al., 2003). An independent analysis of NHANES III data found an inverse
association between likelihood of having attained at least Tanner stage 2 pubic hair development
and current blood lead levels among girls of all ethnicities/races combined (OR = 0.27, 95%
CI 0.08–0.93) (Wu et al., 2003).
PCBs Prenatal, lactational or childhood exposure, inadequate evidence: The North Carolina
birth cohort study reported that adolescent girls with high lactational PCB exposure had earlier
onset of pubic hair of borderline statistical significance (average age at Tanner stage H3, highest vs.
lowest lactational PCB exposure [estimated from breast milk PCB levels and breastfeeding duration], 11.7 vs. 12.6, p-trend over 4 cumulated exposure categories was 0.08) (Gladen et al., 2000).
There was no association with maternal or cord serum PCB (average age at Tanner stage H3, highest vs. lowest maternal or cord serum PCB level, 10.5 vs. 12.0, p-trend over 4 exposure categories
was 0.31). In a Belgian cross-sectional study of adolescents, delayed pubic hair development was
not associated with current serum level of TCDD-like activity (per doubling of serum CALUX assay
for TCDD activity, OR=1.0, p=.97) or with serum PCBs (per doubling of sum of 3 noncoplanar
PCB congeners, OR=1.2, p=.59) (Den Hond et al., 2002).
PBBs Prenatal or lactational PBB exposure, inadequate evidence: A Michigan birth cohort
study observed an association between maternal serum PBB levels and likelihood of Tanner pubic
hair development stage H2 or greater among breast-fed (maternal serum PCB ≥5 vs. ≤1 µg/L,
OR=19.5, 95% CI 2.8–138) but not formula-fed daughters (OR=0.9, 95% CI 0.2–4.3) (Blanck
et al., 2000). The wide confidence intervals of these odds ratio reflect the small numbers of breastfed or formula-fed subjects in the highest maternal serum PBB category.
Age at Breast Development
Lead Childhood blood lead levels, inadequate evidence: A large cross-sectional study of age 8–
16 girls based on NHANES III found associations between delayed breast development and current
blood lead level among African-American girls (likelihood of reaching a successive stage of breast
development, blood lead ≥ 3 vs. <1 µg/dL, OR = 0.62, 95% CI 0.41–0.96) and Mexican-American
girls (OR = 0.70, 95% CI 0.54–0.91) but not among non-Hispanic white girls (OR = 0.75, 95% CI
0.37–1.51) (Selevan et al., 2003). An independent analysis of NHANES III data found no association
between likelihood of at least Tanner stage 2 breast development and current blood lead levels
among girls of all ethnicities/races combined (OR = 1.20, 95% CI 0.51–2.85) (Wu et al., 2003).
PCBs Prenatal or lactational exposure, inadequate evidence: The North Carolina study found
no association between age at breast development and maternal or cord serum PCB level (average
age at Tanner stage B3, highest vs. lowest maternal or cord serum PCB level, 10.1 vs. 11.1, p-trend
over 4 exposure categories was .41) or cumulative lactational PCB exposure (average age at Tanner
stage B3, highest vs. lowest cumulative lactational PCB dose, 11.6 vs. 11.9, p-trend over 4
cumulated exposure categories was .69) (Gladen et al., 2000). A Belgian cross-sectional study of
youth age 15–19 yr observed an association between delayed breast development and serum
TCDD activity (per doubling of serum CALUX assay TCDD activity, OR=2.3, p=.02) but not with
serum PCBs (per doubling of sum of 3 noncoplanar PCB congeners, OR=0.7, p=.49) (Den Hond
et al., 2002).
PBBs Prenatal or lactational PBB exposure, inadequate evidence: The Michigan birth cohort
study observed no association between maternal serum PBB levels and likelihood of Tanner breast
development stage B2 or greater among breast-fed (maternal serum PCB ≥5 vs. ≤1 µg/L, OR=1.2,
95% CI 0.2–6.4) or formula-fed daughters (OR=0.5, 95% CI 0.2–1.9) (Blanck et al., 2000).
Phthalates Childhood exposure, inadequate evidence: In a small case-control study of premature breast development among girls age less than 8 in Puerto Rico, affected girls had higher serum levels of dimethyl-, diethyl- and dibutylphthalate, DEHP and mono-(2-ethylhexyl)phthalate (metabolite
of DEHP) compared to controls (Colon et al., 2000). The presence of phthalate diesters in serum is
suggestive of contamination, as ingested diesters are rapidly metabolized to monoesters in the gastrointestinal tract (McKee, 2004). Combined with the high endemic rate of premature breast development in Puerto Rico, interpretation of these findings is difficult and confirmatory studies have not
yet been published.
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Age at Reproductive System Maturation: Male External Genitalia
PCBs, TCDD Age at external genitalia development Prenatal, lactational or childhood PCB
exposure, inadequate evidence: The North Carolina birth cohort study reported slightly earlier
attainment of Tanner male genital development stage G3 among breast-fed boys with high lactational PCB exposure (average age at G3, highest vs. lowest cumulative lactational PCB dose, 11.5
vs. 12.4, p-trend over 4 PCB dose categories was .07); there was no association with transplacental
PCB exposure (average age at G3, highest vs. lowest cord or maternal serum PCB level, 12.4 vs.
13.0, p-trend=.78) (Gladen et al., 2000). In a small birth cohort study in the Faroe Islands, external
genital development among boys age 14 yr was not associated with cord tissue PCB levels (mean
testicular volume by ascending cord tissue PCB tertile, 6.8, 8.9, 7.5 ml, p-trend=.30; average
Tanner stage of external genital development by ascending cord tissue PCB tertile, 2.1, 2.5, 2.1,
p-trend=.25) (Mol et al., 2002). A Belgian cross-sectional study of youths age 15–19 yr reported an
association between nonattainment of adult-stage male genital development and current serum
PCBs (per doubling of sum of 3 noncoplanar congeners, OR=3.8, p=.06) but not serum TCDD
activity (OR=1.3, p=.46) (Den Hond et al., 2002).
Age at pubic hair development Prenatal, lactational or childhood PCB exposure, inadequate
evidence: The North Carolina birth cohort study reported slightly earlier attainment of Tanner
pubic hair development stage H3 among breast-fed boys with high lactational PCB exposure
(average age at H3, highest vs. lowest cumulative lactational PCB dose, 12.5 vs. 12.9,
p-trend=.35); there was no association with transplacental PCB exposure (average age at H3, highest
vs. lowest cord or maternal serum PCB level, 13.1 vs. 13.1, p-trend=.93) (Gladen et al., 2000).
A Belgian cross-sectional study of youths age 15–19 yr reported an association between nonattainment of adult-stage pubic hair development and current serum PCBs (per doubling of sum of
3 noncoplanar congeners, OR=2.7, p=.06) but not serum TCDD activity (OR=1.1, p=.62) (Den
Hond et al., 2002). In a small birth cohort study in the Faroe Islands, Tanner stage of pubic hair and
genital development among boys age 14 yr was not associated with cord tissue PCB levels (average
Tanner stage by ascending cord tissue PCB tertile, 1.9, 2.4, 1.9, p-trend=.63) (Mol et al., 2002).
Phthalates Infant exposure, inadequate evidence: A small birth cohort study (13 males) found
normal testicular volume and phallic length among adolescent boys who had been neonatally
exposed to DEHP during extracorporeal membrane oxygenation with estimated cumulative doses
of 42–140 mg/kg/bw (Rais-Bahrami et al., 2004).
Summary There is limited epidemiologic evidence for an association between childhood lead
exposure and delayed menarche. There is inadequate evidence for associations between prenatal
or childhood exposures to environmental toxicants and age at breast or pubic hair development in
girls and age at genital development in boys.
Sudden Infant Death Syndrome
Tobacco smoke Maternal active smoking, sufficient evidence: The U.S. Surgeon General
concluded that there is sufficient evidence of a causal relationship between SIDS and prenatal
active smoking (U.S. Department of Health and Human Services, 2004).
Infant exposure, ETS, sufficient evidence: In a meta-analysis of 4 studies that controlled for
prenatal smoking, SIDS was associated with postnatal maternal smoking (pooled OR=1.94, 95% CI
1.55–2.43) (Anderson & Cook, 1997). These reviewers noted that 2 of the 3 available studies involving infants of nonsmoking mothers found associations between SIDS and paternal smoking with odds
ratios of 1.63 (95% CI 1.11–2.40) and 3.41 (95% CI 1.98–5.88). A WHO expert group review concluded that prenatal active smoking is a major cause of SIDS and that there is limited evidence that
childhood ETS exposure increases the risk of SIDS (World Health Organization, 1999). An expert
panel found sufficient evidence of a causal association between SIDS and childhood ETS exposure,
independent of prenatal smoking (California Environmental Protection Agency, 2005). The U.S. Surgeon General concluded that there is sufficient evidence of a causal relationship between SIDS and
ETS exposure during early infancy (U.S. Department of Health and Human Services, 2006). A large
multicentre European case-control study revealed that SIDS was associated with maternal smoking,
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especially among mothers who shared their bed with their infant (OR=17.7, 95% CI 10.3–30.3);
SIDS was also associated with the number of cigarettes smoked daily in the home (≥30 vs. 0 cigarettes/
d, OR=3.31, 95% CI 1.84–5.96) (Carpenter et al., 2004).
Outdoor air pollution Infant exposure, limited evidence: Among studies published up to
2003, reviewers found limited evidence, mainly from ecologic studies, for an association between
SIDS and ambient air pollution (Glinianaia et al., 2004b; Tong & Colditz, 2004). One review noted
little consistency in study design and pollutant measurement and the potential for prenatal smoking
to be a confounder (Tong & Colditz, 2004). Based on air quality data and a relative risk estimate
from a large cohort study, the estimated proportion of SIDS deaths in 23 U.S. metropolitan areas
attributable to PM10 levels exceeding 12 µg/m3 was 16% (95% CI 9–23) (Kaiser et al., 2004). In a
large California retrospective cohort study, SIDS was not associated with average PM2.5 levels within
8 km of the maternal residence (per 10 µg/m3 increment, OR=0.82, 95% CI 0.55–1.23) (Woodruff
et al., 2006).
Summary There is sufficient epidemiologic evidence that prenatal active smoking and ETS
exposure during infancy can cause SIDS; limited evidence supports an association between SIDS
and outdoor air pollution levels near the maternal/infant residence.
Allergies
PCBs Maternal exposure, inadequate evidence: A small Dutch nested case-control study
observed no association between prenatal plasma levels of 4 dioxin-like PCB congeners (118, 138,
153, 180) and the prevalence at age 42 mo of parent-reported eczema (OR=1.18, 95% CI 0.82–1.71)
or allergic reactions to food, pollen, dust, or household pets (OR=0.62, 95% CI 0.29–1.32)
(Weisglas-Kuperus et al., 2000). In a follow-up study at age 7 yr, there was no association between
prenatal plasma PCB levels and a history of allergic reactions (OR=0.95, 95% CI 0.59–1.52)
(Weisglas-Kuperus et al., 2004).
Childhood exposure, inadequate evidence: The Dutch case-control study reported that
children’s current plasma levels of four PCB congeners were not related to a history of eczema
(OR=0.92, 95% CI 0.41–2.08) but were inversely associated with a history of allergic reactions
(OR=0.01, 95% CI 0.01–0.37) (Weisglas-Kuperus et al., 2000). Follow-up at age 7 yr revealed no
relationship between lactational PCB exposure (estimated from breast milk PCB levels and duration
of breastfeeding) and a history of allergic reactions (OR=1.09, 95% CI 0.97–1.22) (Weisglas-Kuperus
et al., 2004).
Pesticides Prenatal exposure, unspecified pesticides, inadequate evidence: The Ontario farm
family study found several associations between childhood hay fever or other allergic conditions
and prenatal farm pesticide use (any pesticide, OR=1.58, 95% CI 1.19–2.08; herbicides,
OR=1.56, 95% CI 1.15–2.11; insecticides, OR=1.48, 95% CI 1.07–2.03; fungicides, OR=1.69,
95% CI 1.15–2.47; 2,4-D, OR=1.66, 95% CI 1.11–2.49; and organophosphate insecticides,
OR=1.55, 95% CI 1.02–2.36) (Weselak et al., 2007). When analyzed by gender, the associations
among males were stronger and those for females were not statistically significant. ORs also tended
to higher for children who were 12 yr of age or greater at the time of the survey. Potential limitations associated with this study include the retrospective data collection, no information on age of
diagnosis or the type and quantity of pesticides used, and multiple comparisons. In addition, no
information on potential postnatal confounders was collected.
Paternal exposure, unspecified herbicides, inadequate evidence Vietnam veterans reported
more often that their children had a history of allergies (crude OR = 1.6, 95% CI 1.2–2.1) but not
allergic rhinitis (crude OR = 1.3, 95% CI 0.9–1.9) compared to non-Vietnam veterans (Centers for
Disease Control. 1989).
Summary There is inadequate epidemiologic evidence for associations between allergic
conditions and prenatal or childhood exposure to environmental chemical contaminants.
Dental Caries
Lead Childhood exposure, limited evidence: A large cross-sectional study based on NHANES
III found a dose-response relationship between dental caries of primary or permanent teeth and
current relatively low blood lead levels, independent of potential confounders (>4.1 vs. ≤2.3 µg/dl,
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OR=1.66, 95% CI 1.12–2.48) (Moss et al., 1999). A small cohort study and an intervention study
found associations between current blood lead levels and deciduous but not permanent tooth
dental caries (Campbell et al., 2000; Gemmel et al., 2002).
Summary There is limited evidence for an association between childhood dental caries and
childhood lead exposure.
Chloracne
PCBs High-level childhood exposure, sufficient evidence: Chloracne occurred among children exposed to high levels of PCBs, PCDFs and related toxicants during the Yucheng incident
(prevalence of acne or acne scars at age 1 mo to 8 yr, exposed vs. unexposed children, 20/117 vs.
10/106, p=.05) (Rogan et al., 1988). When followed to age 12–14 yr, the prevalence of chloracne
scars and comedones among Yucheng children was similar to that of an unexposed comparison
group (exposed vs. unexposed, prevalence 10.0 vs. 8.1%, p > .05) (Hsu et al., 1995).
TCDD High-level childhood exposure, sufficient evidence: After the massive release of
relatively pure TCDD into air at Seveso, 88% of the 187 cases detected through screening were
children age less than 15 (Del Corno et al., 1985). A review of the Seveso incident concluded that
TCDD caused chloracne at doses that cause no other obvious health effects (Sweeney & Mocarelli,
2000). A retrospective cohort study revealed that, compared to unexposed persons, postnatally
exposed persons had greatly increased risks of physician-diagnosed chloracne (men, OR = 13.8,
95% CI 5.6–46.0; women, OR = 17.8, 95% CI 7.9–51.0) (Guo et al., 1999).
Summary There is sufficient epidemiologic evidence that high-level childhood exposure to
TCDD can cause chloracne and limited evidence that prenatal exposure to high-level PCB/PCDF
can cause chloracne.
Renal Tubular Damage This section discusses associations between increased urinary protein
excretion, a sign of renal tubular damage and environmental exposures (see also Table 9).
Lead Childhood exposure, sufficient evidence: Reviewers concluded that lead impairs renal
tubular resorptive function, causing aminoaciduria, glucosuria and hyperphosphaturia at blood lead
levels as low as 10 µg/dl (Goyer, 1990; Loghman-Adham, 1997). Four cross-sectional studies
(Bernard et al., 1995; Fels et al., 1998; Staessen et al., 2001; Verberk et al., 1996) and a birth
cohort study (Factor-Litvak et al., 1999) reported dose-response relationships between childhood
urinary protein markers of renal tubular damage and blood lead levels. These findings occurred at
low to moderate blood lead levels and were independent of potential confounders. A large
cross-sectional study in three European countries observed inverse dose-response relationships
between blood lead and serum cystatin C and β2-microglobulin levels (de Burbure et al., 2006).
This study also reported dose-response relationships between urinary proteins and an interaction
term based on the product of blood lead and urinary cadmium levels.
Mercury Childhood exposure, organic mercury, limited evidence: A cross-sectional study of
infants exposed to diapers treated with a phenyl mercuric fungicide in Argentina reported an association between urinary γ-glutamyl transpeptidase and urinary mercury excretion rates with an
apparent threshold of about 6 µg/kg body wt/day (Gotelli et al., 1985).
Childhood exposure, elemental mercury, inadequate evidence: Two cross-sectional studies
found no association between urinary protein levels among children or youth and number of amalgam tooth surfaces or urinary mercury levels (de Burbure et al., 2003; Herrstrom et al., 1995).
Cadmium Childhood exposure, inadequate evidence: Three cross-sectional studies found
no association between childhood urinary protein levels and blood cadmium (Bernard et al.,
1995; Staessen et al., 2001) or urinary cadmium levels (Noonan et al., 2002). A cross-sectional
study of children living near nonferrous smelters in France found a correlation between urinary
N-acetyl-beta-D-glucosaminidase (NAG) and log blood cadmium (partial r = .25, p = .0002)
(de Burbure et al., 2003).
Summary Epidemiologic evidence for the role of environmental toxicants in childhood renal
tubular damage (causing proteinuria) includes: (a) sufficient evidence—childhood lead exposure;
(b) limited evidence—childhood organic mercury exposure (phenylmercuric mercury used in
diaper rinses).
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Chromosomal Abnormalities The level of epidemiologic evidence for associations between
chromosomal abnormalities is summarized in Table 9.
Arsenic Maternal exposure, inadequate evidence: In a Swedish retrospective cohort study,
there was an elevated risk of chromosomal abnormalities (OR=2.58, 95% CI 0.90–6.70) among
infants of women living in parishes near a copper smelter known to emit arsenic, lead, mercury,
and cadmium (Wulff et al., 1996).
Childhood exposure, limited evidence: In a small cross-sectional study of Argentinean
children, the frequency of micronuclei per 1,000 lymphocytes was substantially higher among
children living in regions with drinking water arsenic levels averaging about 200 µg/L (high vs.
low drinking water arsenic regions, 35 ± 4.6 vs. 5.6 ± 1.6, p < 0.05) (Dulout et al., 1996). In a
Swedish retrospective cohort study, infants of women living near a copper smelter had an
increased risk of chromosomal abnormalities (RR = 2.58, 95% CI 0.90–6.70) (Wulff et al., 1996).
A cross-sectional study of young children in a Mexican mining town with high soil arsenic and
lead concentrations found associations between Comet tail length and moment and urinary
arsenic levels (note—Comet tail length and moment reflect DNA single- and double-stranded
breaks) (Yanez et al., 2003). Reviewers concluded that there is limited evidence that inorganic
arsenic causes chromosomal abnormalities in humans (Agency for Toxic Substances and Disease
Registry, 2000a).
Environmental tobacco smoke Maternal or childhood exposure, inadequate evidence: A
meta-analysis of studies published during 1980–2004 concluded that children exposed to ETS and
neonates prenatally exposed to maternal smoking, respectively, had hemoglobin adduct concentrations 1.4 and 6.7 times those of unexposed subjects but sister chromatid exchange frequency was
not associated with either exposure (Neri et al., 2006).
Outdoor air pollution Maternal exposure, inadequate evidence: A retrospective cohort study
in France observed no association between chromosomal abnormalities and traffic density near the
prenatal residence (>50,000 vs. <10,000 vehicles/d, OR=0.90, 95% CI 0.60–1.37) (Cordier et al.,
2004).
Drinking-water disinfection by-products Maternal exposure, inadequate evidence: A retrospective cohort study in Sweden found no association between chromosome abnormalities at birth
and prenatal residence in communities using hypochlorite (compared to unchlorinated water,
OR=0.8, 95% CI 0.5–1.3) or chlorine dioxide-treated drinking water (OR=0.7, 95% CI 0.4–1.2)
(Kallen and Robert 2000). A similar study in Canada found a nonmonotonic dose-response relationship between chromosomal abnormalities at birth and periconceptual community drinking water
chloroform levels (50–74 vs. <50 µg/L, OR=1.3, 95% CI 0.8–2.2; 75–99 µg/L, OR=1.9, 95% CI
1.1–3.3; ≥100 µg/L, OR=1.4, 95% CI 0.8–2.8); chromosome abnormalities were not associated
with BDCM levels (Dodds & King, 2001).
Hazardous waste disposal Maternal exposure, inadequate evidence: In a population-based
case-control study, chromosome abnormalities were weakly associated with prenatal residential
proximity to hazardous waste disposal sites (<1.6 vs. ≥1.6 km, OR = 1.18, 95% CI 0.90–1.55);
the association was stronger for the subgroup of sites containing plastics (OR = 1.46, 95% CI
1.01–2.11) (Geschwind et al., 1992). A European study case-control study revealed an association between chromosomal abnormalities and prenatal residential proximity to any of 23
hazardous waste disposal sites (≤3 vs. 3–7 km, OR= 1.49, 95% CI 1.03–2.17) (Vrijheid et al.,
2002a). However, there was a nonmonotonic dose-response relationship between hazard
categories based on expert-rated potential for toxicant exposure via air or water and structural
chromosome abnormalities (high vs. low hazard, OR= 1.65, 95% CI 0.83–3.29, p-trend = 0.31)
(Vrijheid et al., 2002b).
Incinerators Maternal exposure, inadequate evidence: In a French retrospective cohort study,
chromosome abnormalities were not associated with prenatal residence in communities with solid
waste incinerators (OR=1.01, 95% CI 0.86–1.20) (Cordier et al., 2004).
Solvents Maternal exposure, unspecified solvents, inadequate evidence: A hospital-based
case-control study in France reported no association between chromosomal abnormalities and
prenatal occupational solvent exposure (OR=0.7, 90% CI 0.2–1.7) (Cordier et al., 1992).
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Paternal occupational exposure, inadequate evidence: A Norwegian cohort study of offspring of
male printers revealed no increased risk of Down syndrome (compared to other occupations,
SIR=1.4, 95% CI 0.65–2.5) (Kristensen et al., 1993).
Summary Limited epidemiologic evidence supports an association between childhood
chromosomal abnormalities and childhood exposure to high drinking water arsenic levels.
CONCLUSION
This review identifies known and suspected relationships between environmental chemical
contaminants and adverse pregnancy and child health outcomes as well as many supported by
inadequate epidemiologic evidence. Known causes of adverse pregnancy and childhood health
outcomes include:
Maternal exposures:
• methylmercury
• high-level exposure (delayed developmental milestones and cognitive, motor, auditory and visual
deficits)
• PCBs, PCDFs, and related toxicants
• high-level exposure (neonatal tooth abnormalities, cognitive and motor deficits)
• residential environment
• active smoking (delayed conception, preterm birth, FGD, SIDS)
• ETS (preterm birth)
Childhood exposures:
•
•
•
•
•
•
•
•
lead
low-level exposure (cognitive deficits and renal tubular damage)
methylmercury
high-level exposure (visual deficits)
TCDD
high-level exposure (chloracne)
indoor air contaminants
ETS (SIDS, new-onset asthma, increased asthma severity, lung infections, middle ear infections,
adult lung and breast cancers)
• biomass smoke (lung infections)
• outdoor air pollutants (increased asthma severity)
Among proven environmental hazards, several are supported by epidemiologic studies of highly
exposed populations (e.g., MeHg, PCBs/PCDFs). Lead and ETS, however, produce adverse effects
after relatively low level prenatal or childhood exposure. For instance, there are dose-response relationships between full-scale IQ deficits and blood lead levels below 10 µg/dl, the current U.S.
Centers for Disease Control and Prevention (CDC) action level. The CDC stated these reasons for not
reducing the current lead action level : (1) No effective clinical interventions are known to lower the
blood lead levels for children with levels less than 10 µg/dl or to reduce the risk for adverse developmental effects, (2) children cannot be accurately classified as having blood lead levels above or
below a value less than 10 µg/dl because of the inaccuracy inherent in laboratory testing, and (3) no
evidence exists of a threshold below which adverse effects are not experienced; thus, any decision to
establish a new level of concern would be arbitrary and provide uncertain benefits (Centers for
Disease Control and Prevention, 2006). There appears to be no lead exposure threshold for neurotoxicity in children, suggesting a need for public health policies aimed at virtual elimination of lead
exposure related to housing (paint, plumbing), ambient air, consumer products and drinking water.
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Many relationships between relatively low-level environmental exposures and adverse pregnancy or child health outcomes were supported by limited epidemiologic evidence (see Tables 1–9).
For many of the relationships supported by limited or inadequate epidemiologic evidence,
there is a scarcity of published studies, as opposed to inconsistencies among several large, highquality studies. For instance, several environmental exposures are known to exacerbate existing
asthma but there has been less research on the role of environmental factors in asthma onset. About
20 epidemiologic studies have linked childhood leukemia to pesticide exposure (Bérubé, 2006),
but most have lacked the statistical power and exposure quantification and specificity needed to
produce strong evidence for or against causal relationships.
The burden of adult disease attributable to preconceptual, prenatal, or childhood environmental exposures is largely nonquantifiable at present. Known links include chronic disability from
environment-related childhood diseases (birth defects, asthma, cancer). There is limited evidence
for associations between childhood exposure to ETS and breast cancer.
The dose and timing of exposure to toxicants are important determinants of adverse effects on
developing tissues (Faustman et al., 2000). For instance, prenatal or early childhood exposures to
lead, MeHg and PCBs produce more severe neurotoxicity compared to similar exposures of adults.
However, there are many knowledge gaps concerning the role of early-life low-level exposure to
environmental toxicants in child health and development. Major challenges include the need to
identify the potential health effects of both specific exposures (e.g., a specific type of pesticide) and
multiple exposures. To date, relatively few epidemiologic studies of child health and development
have assessed potential interactions between two or more toxicants.
A child-centered agenda for research and risk assessment should include (Landrigan, 1999): (1)
exploration and quantification of unique exposure patterns among children, (2) adoption of more
sensitive methods to test chemicals for developmental toxicity, (3) clinical and epidemiologic studies
to identify child health outcomes of environmental toxicants, and (4) cellular and molecular
research on the pathogenesis of pediatric environmental illness. Although the U.S. federal government
has moved in the direction recommended by Landrigan, Canada and most other countries have
not. A reviewer recently recommended that Canada conduct an analysis of the economic and
social costs of adverse child health outcomes related to environmental hazards to support investments in research in this field (Bérubé, 2006). This review proposed that Canada invest initially in
biomonitoring of children’s and pregnant women’s environmental exposures and initiate a longitudinal cohort study of environmental influences on child health and development. Developed
countries in particular should show leadership and strengthen national and international population
and laboratory research on the role of environmental contaminants in fetal, child and adult health
and development. One of the most pressing needs is large-scale, dedicated funding of new research
infrastructure (scientists, laboratories, childhood disease registries) and large-scale epidemiologic
research programs and projects. Only adequate funding will permit the conduct of high quality
epidemiologic studies that incorporate strong statistical power, robust exposure assessment and
sufficient size to explore dose-response relationships, biomarkers of exposure and susceptibility,
and the potential roles of confounding variables.
Longitudinal studies of relatively large populations (e.g., Framingham Heart Study, U.S. Nurses
Health Study) have provided a wealth of knowledge about risk factors for heart disease, cancer and
other adult health conditions. The equivalent investigations of child health and the environment
would be large longitudinal studies with intensive environmental exposure assessments beginning
before conception or during early pregnancy and prolonged follow-up to identify health outcomes
during pregnancy, infancy, childhood, adolescence and adulthood. Such studies, initiated in
Europe and at the planning stage in the United States, are extremely valuable and will provide a
wealth of information on a wide variety of potential health outcome and environmental exposure
relationships (Golding et al., 2001; National Institute of Child Health and Human Development
2002). Given the unique prenatal and childhood environments in diverse geographic regions, other
national and international bodies with mandates for environmental health should fund similar longitudinal studies. With advances in understanding of the human genome and molecular markers of
environmental contaminant exposure and improved epidemiologic study design, it seems likely that
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future studies will be able to better define the range and risks of environmental health outcomes in
children.
National and smaller scale biomonitoring systems have documented successes and emerging
issues, including:
• Declining levels of blood lead (Annest JL et al., 1983), serum or urinary cotinine (indicators of ETS
exposure) (Pirkle et al., 2006) and breast milk p,p¢-DDT/p,p¢-DDE (the latter is the major metabolite
of DDT) (Atuma et al., 1998).
• Increasing levels of PBDEs (flame retardants) in breast milk (Meironyte et al., 1999).
• Unexpectedly high urinary levels among children and adults of metabolites of several phthalates
used in many consumer products including cosmetics, pharmaceutical coatings, food containers/
packaging and toys (Centers for Disease Control and Prevention, 2005b).
• Low levels of TCDD: This potent toxicant was detectable in only 0.7% of U.S. serum samples
from persons age 6 yr or older, and observed levels were far below those associated with health
effects in occupationally or accidentally exposed persons (Centers for Disease Control and
Prevention, 2005b).
• Widespread maternal and fetal plasma bisphenol A concentrations in the range showing reproductive toxicity in experimental animal offspring (Calafat et al., 2005; Schonfelder et al., 2002).
• Among women undergoing routine 2nd trimester amniocentesis, about a third of amniotic fluid
samples contained detectable levels of hormonally active environmental contaminants including
α-HCH, ρ,ρ?-DDE and specific PCB congeners (Foster et al., 2000). Although contaminant
concentrations were quite low, these findings show that the fetus is directly exposed to potentially
harmful hormonally active toxicants.
Only the United States and Germany have implemented national population biomonitoring to
measure and track exposures to environmental contaminant levels in human blood, urine, hair,
breast milk and other samples. The Commission for Environmental Cooperation, under the North
American Regional Action Plan on Environmental Monitoring and Assessment, plans to implement
biomonitoring of persistent toxicants in neonates and infants in Canada, the United States, and
Mexico (Commission for Environmental Cooperation, 2002).
For the environmental contaminant and adverse health outcome relationships supported by
limited epidemiologic evidence, there are several major health policy and program needs: (1) In the
case of proven causal relationships, ongoing interventions are essential to minimize exposure levels
in the general population and in high-risk subgroups. (2) For relationships supported by limited evidence, precautionary interventions are warranted to minimize population exposures. (3) Biomonitoring
of toxicant exposure is essential to demonstrate the effectiveness of existing interventions and to
identify subgroups with exposures exceeding risk-based standards or guidelines. (4) For the many
knowledge gaps, strengthened epidemiologic research is needed to build the evidence base.
Many examples of knowledge gaps are documented in this review for relationships supported by limited or inadequate evidence. Moreover, we have adequate toxicologic assessments for only a fraction of the thousands of chemical toxicants in the environment and
something approaching adequate epidemiologic evidence for very few of them. The potential
anthropogenic sources of environmental chemicals include fossil fuel combustion, manufacturing processes, various uses of commercial products (pesticides, building materials, solvents),
human activities (e.g., smoking indoors), waste disposal (hazardous waste disposal sites, incineration), and accidents. One of the major drivers is the vast and rapidly growing number of
commercial chemicals—over 70,000 commercial chemicals are registered for use in the United
States, and the U,.S. EPA annually receives about 1500 petitions to approve new chemicals or
new uses of existing chemicals (U.S. Environmental Protection Agency, 2001). Intervention and
tracking programs should target not only the general population but also subgroups, especially
those at high risk (e.g., because of poverty, occupation, geographic region or other factors).
Considerable time and effort are generally needed to develop sufficient evidence of causal
relationships in observational human research studies. Protection of public health requires
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489
precautionary interventions when faced with limited scientific evidence and related uncertainties. Failure to take precautionary measures can have disastrous consequences, e.g., the introduction of tetraethyl lead as a fuel additive during the early 20th century over the objections of
public health authorities (Needleman, 1997).
Although this review does not address specific intervention needs, others have identified a broad
range of needs, including: (1) explicit recognition of children’s vulnerability in environmental legislation; (2) improved developmental and reproductive toxicity testing of commercial chemicals;
(3) coordination of chemical toxicity testing; (4) creation of publically accessible databases on toxicity
of commercial chemicals; (5) increased capacity for epidemiologic and toxicologic research and
biomonitoring of environmental contaminants in children—including infrastructure (personnel,
equipment) and major research program funding; (6) increased operational capacity through establishment of a national comprehensive Children’s Environmental Health Program to oversee resource
allocation, research and biomonitoring initiatives and to ensure that new policies and regulations for
hazards address children’s health; (7) increased legislative capacity to address children’s environmental health through amendments to the Canadian Environmental Protection Act (CEPA), reviews
of other relevant federal legislation to identify where an additional margin of safety is needed to protect child health, examination of legislation across government jurisdictions to assess the net impact
on total environmental exposures of children and reproductive-age adults (e.g., legislation relating to
toxicants in foods, drinking water, air, commercial products) and development of mechanisms to
ensure that new research findings are incorporated in a timely fashion into future risk assessments
and risk management decisions; (8) strengthened communication and education strategies to assure
that clear and effective public health messages reach critical target groups; and (9) assured accountability through annual reports to Parliament including measures taken to protect and improve children’s environmental health (Krewski et al., 2007; Tyshenko et al., 2006). Increased investment in
environmental interventions promises to be offset not only be reduced health care costs, but also by
improved health status and productivity of tomorrow’s labor force (Landrigan et al., 2002).
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