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Mini-Monograph
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Lessons Learned for the Assessment of Children’s Pesticide Exposure: Critical Sampling and Analytical Issues for Future Studies Richard A. Fenske,1 Asa Bradman,2 Robin M. Whyatt,3 Mary
S. Wolff,4 and Dana B. Barr5 1Department of Environmental and Occupational Health Sciences, School
of Public Health and Community Medicine, University of Washington, Seattle,
Washington, USA; 2Center for Children’s Environmental Health
Research, School of Public Health, University of California, Berkeley, California,
USA; 3Columbia Center for Children’s Environmental Health,
Mailman School of Public Health, Columbia University, New York, USA; 4Department
of Community and Preventive Medicine, Mount Sinai School of Medicine, New York,
New York, USA; 5National Center for Environmental Health, Centers
for Disease Control and Prevention, Atlanta, Georgia, USA Abstract In this article we examine sampling strategies and analytical methods used in a series of recent studies of children’s exposure to pesticides that may prove useful in the design and implementation of the National Children’s Study. We focus primarily on the experiences of four of the National Institute of Environmental Health Sciences/U.S. Environmental Protection Agency/ Children’s Centers and include University of Washington studies that predated these centers. These studies have measured maternal exposures, perinatal exposures, infant and toddler exposures, and exposure among young children through biologic monitoring, personal sampling, and environmental monitoring. Biologic monitoring appears to be the best available method for assessment of children’s exposure to pesticides, with some limitations. It is likely that a combination of biomarkers, environmental measurements, and questionnaires will be needed after careful consideration of the specific hypotheses posed by investigators and the limitations of each exposure metric. The value of environmental measurements, such as surface and toy wipes and indoor air or house dust samples, deserves further investigation. Emphasis on personal rather than environmental sampling in conjunction with urine or blood sampling is likely to be most effective at classifying exposure. For infants and young children, ease of urine collection (possible for extended periods of time) may make these samples the best available approach to capturing exposure variability of nonpersistent pesticides ; additional validation studies are needed. Saliva measurements of pesticides, if feasible, would overcome the limitations of urinary metabolite-based exposure analysis. Global positioning system technology appears promising in the delineation of children’s time-location patterns. Key words: children, exposure, GPS, organophosphates, pesticides. Environ Health Perspect 113: 1455-1462 (2005) . doi:10.1289/ehp.7674 available via http://dx.doi.org/ [Online 24 June 2005] This article is part of the mini-monograph “Lessons Learned from the National Institute of Environmental Health Sciences/U.S. Environmental Protection Agency Centers for Children’s Environmental Health and Disease Prevention Research for the National Children’s Study.” Address correspondence to R. Fenske, Department of Environmental and Occupational Health Sciences, Box 357234, University of Washington, Seattle, WA 98195 USA. Telephone: (206) 543-0916. Fax: (206) 616-2687. E-mail: rfenske@u.washington.edu The National Children’s Study provided support for the preparation of the manuscript. The Children’s Center studies were supported by grants ES009605, ES009601, ES009584, and ES009600 from the National Institute of Environmental Health Science and R826709, R826886, R827039, and R827027 from the U.S. Environmental Protection Agency (EPA) . University of Washington Pacific Northwest Agricultural Safety and Health Center studies were supported by grants R819186, 916001537, and R82517101 from the U.S. EPA and S147-14/16 and U07/CCU012926 from the National Institute for Occupational Safety and Health (Association of Schools of Public Health and Agricultural Centers Program) . The authors declare they have no competing financial interests. Received 12 October 2004 ; accepted 24 May 2005. |
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Accurate characterization of children’s
exposure to pesticides has proven to be a particularly
challenging aspect of the field of exposure assessment.
First, the term “pesticides” encompasses
a diverse array of chemicals that can potentially
produce a wide variety of health effects. Second,
exposure of children to pesticides can occur through
multiple pathways and routes. For example, the U.S.
Environmental Protection Agency (EPA) considers food,
drinking water, and residential pesticide use all
to represent important sources of exposure, and these
exposures can occur simultaneously or sequentially
through the routes of ingestion, inhalation, and
dermal contact (Cohen Hubal et al. 2000). Certain
subpopulations, such as children living in agricultural
communities or children whose parents work with pesticides,
may be exposed through additional pathways. Third,
many pesticides have short residence times in the
body, making it difficult to characterize exposures
from biologic samples. Finally, chemical exposures
may have substantially different health consequences
for children depending on the developmental stage
during which the exposure occurs, requiring exposure
characterization at multiple time points.
Our purpose in this article is to examine sampling
strategies and analytical methods associated with
a series of recent population studies that have sought
to characterize children’s pesticide exposure,
and to distill from these experiences a number of
lessons learned. In this article, we focus primarily
on the experiences of the National Institute of Environmental
Health Sciences/U.S. EPA Children’s Centers
located at Columbia University, the University of
California at Berkeley, Mount Sinai Medical Center,
and the University of Washington. We have also included
a review of several University of Washington studies
that predated establishment of the children’s
centers and that were conducted under the auspices
of the Pacific Northwest Agricultural Safety and
Health (PNASH) Center, sponsored by the National
Institute for Occupational Safety and Health, and
the U.S. EPA Science To Achieve Results (STAR) Grant
Program. This article is not meant to be an exhaustive
review of exposure assessment methods, but rather
a first-hand commentary on the use of particular
methods in our studies. We therefore have not been
able to include an analysis of a number of important
studies conducted at other institutions, such as
the Minnesota Children’s Pesticide Exposure
Study (Adgate et al. 2001; Quackenboss et al. 2000)
and studies of children’s exposure along the
U.S.-Mexican border (U.S. EPA 2004).
In this article we first examine the rationale
and methods of exposure data collection in the population
studies and then review the substantial challenges
associated with the analysis of pesticides in novel
and complex matrices, and the interpretation of these
analytical findings. It is our hope that experience
gained from this work will prove useful to researchers
embarking on longitudinal cohort studies, such as
the proposed National Children’s Study.
Sampling Strategies in Population Studies
Table
1
![Table 1](tab1sm.gif) |
Table 2
![Table 2](tab2sm.gif) |
Data used to construct exposure estimates or classifications
can be drawn from a variety of sources, ranging from
general information regarding pesticide use to personal
measurements. Table 1 presents the approaches taken
in the studies under review. The first two columns
provide source information and environmental measurement
methods; the remaining columns categorize various
types of exposure samples collected according to
age, because different sampling strategies are more
or less practical and valuable within these time
frames. Table 2 indicates the analytes measured in
five biologic sample matrices collected in these
studies.
Pesticide source information. Virtually
all children’s exposure studies collect historical
and contemporaneous information regarding pesticide
use. In most cases, these data are collected through
parental questionnaires or interviews and pertain
to pesticides in and around the residence. In general,
we have found that parents are best able to provide
general information regarding the use of products
(e.g., control of particular insects, control of
weeds) but may not be able to provide detailed information
on specific chemicals (Lu et al. 2004; Whyatt et
al. 2002). In preliminary analyses of questionnaires
administered by the Columbia center, women provided
a pesticide product name for only 39% of the pest
control methods reported to be used in the home during
pregnancy and, in particular, were rarely able to
identify the pesticide products used by an exterminator.
Further, pesticide products can have the same brand
name but contain different active ingredients, further
complicating use of questionnaire data in pesticide
exposure assessment.
Investigators for most of the reviewed studies
have thus gone a step further to visually inspect
the pesticide products in the home, sometimes referred
to as a pesticide inventory. For example, study staff
from the Berkeley center recorded the U.S. EPA registration
number and the active ingredients on the label of
each home pesticide. The registration number was
later entered into a pesticide product database maintained
by the California Department of Pesticide Regulation
to confirm all active ingredients. Records of commercial
pesticide applications can also be accessed during
home visits (Berkowitz et al. 2003; Whyatt et al.
2003).
Identification of specific products can be very
helpful in determining whether or not a particular
class of chemicals has been used in the residence
and may inform subsequent sampling plans, but the
presence or absence of specific products does not
generally enter into the development of an exposure
metric for the residents. Frequency of residential
pesticide use could be used potentially to sort children
into exposure categories, but such an approach has
not been fully validated. One study has shown that
personal air levels of organophosphate (OP) pesticides
were significantly higher among women who reported
using exterminator sprays, can sprays, and/or pest
bombs during pregnancy compared with those reporting
no OP pesticide use (Whyatt et al. 2002, 2003). Another
study demonstrated that children whose parents reported
garden use of insecticides had higher levels of OP
pesticide metabolites than did children whose parents
did not use garden insecticides (Lu et al. 2001).
Food can be an important source of pesticide exposure
for children, but most of the studies reviewed here
have not devoted substantial resources to an evaluation
of the dietary pathway. The Mount Sinai center obtained
maternal prenatal dietary food frequency data during
pregnancy only, with specific information about fish
consumption. The Berkeley center also obtained a
detailed prenatal food frequency questionnaire. Additional
information was also obtained on fruit and vegetable
consumption for the pregnant women and, later on,
for their children. The Berkeley center and the PNASH
center have collected duplicate diets from a relatively
small number of children (Fenske et al. 2002a). Such
an approach provides very useful quantitative information
on exposure but is extremely time-consuming and expensive.
A diet diary has also been used to distinguish children
whose intake of fresh produce and juices was primarily
organic and proved effective in classifying children’s
OP pesticide exposure (Curl et al. 2003a).
Studies of children of agricultural workers have
focused on potential paraoccupational exposure, collecting
data on the transmission of pesticides from the workplace
to the home by parents or other adult household members,
as well as data on residential proximity to pesticide
applications (Bradman et al. 1997; Curl et al. 2002;
Eskenazi et al. 2003; Koch et al. 2002; Lu et al.
2000; Simcox et al. 1995). Results to date indicate
that both of these pathways can contribute to children’s
exposures in agricultural communities and would need
to be considered in the design of a study that included
rural populations. Studies at the Berkeley center
have taken advantage of California’s unique
Pesticide Use Reporting system, and researchers there
are investigating the use of these data as predictors
of pesticide exposure in their cohort (Castorina
et al. 2003). The Washington center completed a 2-year
intervention to reduce take-home exposure in 2002;
the Berkeley center is currently conducting a similar
intervention.
Environmental monitoring. House dust
samples have been collected in most of the reviewed
studies and have served as a reliable indicator of
residential pesticide contamination (studies conducted
at the PNASH Center), although not necessarily as
a surrogate for children’s exposures (Curl
et al. 2002; Fenske et al. 2002b; Lu et al. 2000;
Simcox et al. 1995). A practical problem can arise
when insufficient dust is available for analysis,
as was the case for the Mount Sinai studies. In the
Berkeley center studies, the average mass of 509
dust samples was 9 g/m2. The average of
the fraction < 150 µm in diameter used for
chemical analyses was 7 g/m2. About 20%
of the samples had a fine fraction of < 0.5 g
total. Most laboratory methods for pesticides require
0.5-2 g dust. It is likely that only a single chemical
analysis will be possible for a significant fraction
of homes, thus limiting future tests for other chemicals.
The Berkeley, Mount Sinai, and PNASH centers have
investigated alternate methods of measuring pesticide
concentrations in child environments, such as indoor
air and surface wipe sampling (Lu et al. 2004). A
protocol that is currently being validated involves
mailing study participants an alcohol wipe with instruction
for wiping dust on the top of a specified doorframe.
The sample is then placed in a resealable plastic
bag and mailed back to the study team. Advantages
include low cost of sample collection and low participant
burden. However, research is currently ongoing to
determine detection limits and detection frequencies
using this method.
The Columbia center has conducted extensive indoor
air sampling. For chlorpyrifos and diazinon, the
correlation between 48-hr personal air samples collected
from the mother during the third trimester and average
2-month indoor air levels over the final 2 months
of pregnancy were strong (r > 0.7, p < 0.001)
(Whyatt et al. 2003). Air and dust levels were not
significantly correlated in a pilot study conducted
by the Mount Sinai group; this may have been due
to the very small amount of dust collectable in these
homes (Markowitz S, personal communication). In addition
to the OP pesticides several carbamates and pyrethroids
have been measured in personal air samples collected
from the mother over 48-hr during pregnancy (Whyatt
et al. 2002, 2003).
Evidence of chemicals in a child’s environment
does not necessarily provide the basis for a sound
exposure metric. Dust, wipe, and indoor air measurements
(including personal air samples) have not shown strong
associations with biologic measurements (Curl et
al. 2002; Whyatt et al. 2003). It is not clear whether
the lack of strong associations is due to confounding
factors (e.g., dietary exposure), to variability
in the biologic measurements (including toxicokinetic
considerations (discussed below), or to a relatively
weak link between residential contamination and child
exposures.
Environmental monitoring in these studies has focused
almost exclusively on the home or residential setting
and has not yet been extended to child care centers
and schools. The Washington studies have included
wipe sampling and dust sampling of commuter vehicles
of workers to document the movement of agricultural
pesticides from the workplace to the home (Curl et
al. 2002; Lu et al. 2000).
Hand wipe sampling. Initial attempts
to look at direct child exposures have included the
use of hand wipes to collect pesticides from children’s
hands. These methods include wiping the child’s
hand with sterile gauze dressing pads that have been
moistened with isopropanol, or asking the child to
place a hand in a bag containing isopropanol (Bradman
et al. 1997). Gordon et al. (1999) found excellent
correlations between chlorpyrifos in indoor air and
corresponding dermal wipes but poor correlations
between chlorpyrifos in dust and dermal wipes. Another
study reported weak associations between OP pesticide
concentrations in hand wipes, house dust, and urinary
levels of OP metabolites (Shalat et al. 2003). The
Columbia center conducted hand wipes but found all
samples to be less than the limit of detection.
Clothing dosimeters. Other techniques
for assessing children’s dermal exposures include
use of clothing dosimeters such as cotton gloves,
union suits, and socks (Fenske 1993; Lewis 2005).
The Berkeley center has experimented with clothing
dosimeters in recent studies. Infants (children 6
and 12 months of age) wore precleaned cotton socks
and union suits for several hours in their residential
environments.
Maternal exposure. Personal air sampling
has been used effectively to monitor maternal exposures
during pregnancy by Columbia researchers (Whyatt
et al. 2002, 2003). Investigators used motion detectors
to determine whether or not the women complied with
the request to carry the personal air monitors; motion
detectors were installed in the backpacks of randomly
selected women. Results were obtained from monitors
worn by 113 women for approximately 48 hr each. For
the average woman, nearly 95% of the total number
of motion detections occurred during waking hours.
In addition, 98% of the women self-reported that
the air monitor was near them for least 40 of the
48 hr of the personal air monitoring.
This study (Whyatt et al. 2003) also found that
levels of several OP and carbamate pesticides measured
in the 48-hr personal air samples were significantly
correlated with levels in 2-week indoor air samples,
indicating that, at least for these pesticides, the
48-hr air samples provided a reasonable estimate
of exposure over a longer period during pregnancy.
In addition, there was little variability in indoor
air levels of the insecticides, and the correlations
between each of the insecticides in each of the 2-week
air samples were highly significant. In cases where
sampling bracketed an application event, it is likely
that high levels would be observed initially, increasing
temporal variability.
Blood samples have been collected throughout pregnancy
to assess body burden of pesticides in the Berkeley,
Columbia, and Mount Sinai center studies. No association
was seen between insecticide levels in maternal blood
collected at delivery and maternal self-reported
pesticide use during pregnancy in one study (Whyatt
et al. 2003). Weak correlations were seen between
pesticide levels in the maternal personal air samples
collected during pregnancy and in blood samples collected
at delivery (r = 0.10-0.19). However, the
correlations were generally stronger when analyses
were restricted to women for whom the personal air
sample was collected within a month of collection
of the blood samples at delivery (r = 0.13-0.45).
Maternal and umbilical blood insecticide levels (chlorpyrifos,
diazinon, the propoxur metabolite 2-isopropoxyphenol,
and bendiocarb) at delivery were highly correlated,
indicating that the pesticides are readily transferred
to the fetus during pregnancy. Significant inverse
associations were seen between chlorpyrifos in umbilical
cord blood and both birth weight and length, whereas
no association was seen between chlorpyrifos in maternal
personal air samples and the same measures of fetal
growth (Whyatt et al. 2004). These results suggest
that the biomarkers may better reflect exposure from
all routes, not only the amount of insecticides absorbed
by the mother but also the amount of the absorbed
dose that has been transferred to the developing
fetus (Whyatt et al. 2004).
Urine samples have also been collected from women
during pregnancy in several studies. Investigators
at the Berkeley center found that pesticide metabolites
in samples collected in the first and third trimester
were not correlated. Within-person variability was
approximately two times higher than between-person
variability, suggesting that more urine samples collected
during pregnancy would improve exposure classification
(Eskenazi et al. 2004). A moving estimate of the
coefficient relating dimethyl OP metabolite levels
to shorter gestation was used to show that exposures
in later pregnancy may be associated with shorter
pregnancies. Blood cholinesterase levels were inversely
correlated with gestational duration, consistent
with findings for dimethyl OP pesticide metabolites,
although no significant correlation between blood
cholinesterase and urinary metabolite levels was
observed.
The Mount Sinai center collected urine samples
in the third trimester of pregnancy and found that
approximately 70% of the women in the cohort had
been exposed to pesticides, but no associations were
found between these biologic levels and pesticide
questionnaire data (Berkowitz et al. 2003). In a
preliminary analysis of data from the Columbia center,
weak but significant correlations were seen between
average chlorpyrifos and diazinon levels in indoor
air samples collected over the final 2 months of
pregnancy and their respective metabolites in urine
samples collected biweekly from the mothers over
the same time frame.
In summary, it is unlikely that questionnaire data
alone can prove adequate for exposure classification
of women during pregnancy. However, it appears that
systematic monitoring through personal air sampling
and biologic monitoring in combination with questionnaire
data would yield useful exposure data for epidemiologic
investigations.
Perinatal exposure. Several novel
sampling methods are under development to determine
perinatal exposure levels, including sampling of
amniotic fluid, meconium, and cord blood. A pilot
study from the Berkeley center of 100 amniotic fluid
samples, slated for disposal after amniocentesis,
were analyzed for a number of pesticides and their
metabolites, including the OP pesticides (Bradman
et al. 2003). Target analytes were detected with
frequencies ranging from 5 to 70%. Levels were low
compared with levels reported in urine, blood, and
meconium. Because of risks to the fetus, amniotic
fluid typically can be collected only when medically
indicated amniocenteses are conducted, usually around
18-20 weeks of gestation, or during scheduled cesarean
sections. Therefore, the population sampled will
not necessarily be representative of a larger population
of pregnant women. For women already undergoing this
procedure, the collection of amniotic fluid for research
purposes is noninvasive and causes no additional
risk.
At the Columbia center, meconium samples were collected
from 20 newborns and analyzed for OP pesticide metabolites
(Whyatt and Barr 2001). Detection frequencies were
very high for some of these analytes, but others
were not detected. Metabolite levels were similar
to those seen in adult urine in population-based
research. Metabolites were stable at room temperature
over 12 hr. These initial results indicate that the
measurement of pesticide levels in meconium has promise
as a biomarker of prenatal exposure.
Cord blood has been sampled in three studies. Mount
Sinai center investigators collected cord blood for
enzyme, lead, and gene analyses. The Mount Sinai
group relied on hospital staff for cord blood retrieval,
with prenotification of impending delivery and a
note on the chart, with the result that 59% of the
participants’ cord blood was obtained. Columbia
center investigators reported that successful collection
of these samples required that a member of the research
staff team follow the progress of the labor, go to
the labor room before delivery to remind the delivery
room staff that the woman is in the study, and assist
with the sample collection. Umbilical cord blood
was obtained by syringing the blood into heparinized
syringes at the point the cord enters the placenta.
To date, a cord blood sample has been obtained from
81% of the infants in the study. An average of 29
mL (range, 2-58 mL) was collected per delivery, with > 22
mL collected in 75% of deliveries and ≥ 30
mL collected in 50% of the deliveries (Whyatt et
al. 2003). The Berkeley center investigators reported
a similar proportion of cord blood samples collected
and found that successful collection of cord blood
required close cooperation with hospital staff to
develop procedures that eliminated risks of inadvertent
sticks (Eskenazi et al. 2003).
In summary, the perinatal sampling procedures described
here are in the early stages of development and will
need additional study and validation. However, they
hold promise for collecting quantitative exposure
data at a critical stage of child development.
Infant and toddler exposure. Traditional
urine bags have been used in clinical settings and
have proven useful for pesticide-related studies
in children (Royster et al. 2002). The Berkeley center
has been successful collecting urine from children
6-24 months of age who were not toilet trained. Urine
was collected by applying pediatric urine bags to
the children during office or home visits (Eskenazi
et al. 2003). When children were not able to produce
a void during scheduled contacts, study staff trained
parents to apply the urine bag at home and to then
place the urine in a clean cup provided to them.
The parent was instructed to call the field office
as soon as the void was produced, and study staff
then retrieved the sample.
Cotton inserts have also been used to recover urine
from diapers (Hu et al. 2000). However, the most
promising development for sampling infants and toddlers
who are not yet toilet trained appears to be extracting
the metabolites from the diaper gel matrix, although
this method still needs to be evaluated for multiple
groups of pesticides (Hu et al. 2004).
Preschool children’s exposure. Urine
samples have been collected in nearly all studies
of pesticide exposure among preschool children. Urine
samples have been analyzed for common metabolites,
such as the dialkylphosphate (DAP) compounds or for
compound-specific metabolites [e.g., 3,5,6-trichloro-2-pyridinol
(TCPy) for chlorpyrifos]. Major exposure assessment
issues of concern are duration of collection (spot
samples vs. 24-hr samples) and frequency of sampling.
Collection of single urine voids, often referred
to as spot urine samples, has been selected as a
primary sampling strategy for several practical reasons.
The burden it places on study participants is relatively
low, and sample processing and analysis are manageable
and affordable. However, several studies have now
determined that pesticide metabolite concentrations
in children’s spot urine samples can exhibit
high intraindividual (within-child) variability (Adgate
et al. 2001; Koch et al. 2002). In studies in which
it is possible to collect only a single urine sample
per day, the first morning void is preferred, because
the urine is more concentrated, the collection period
is longer (usually > 8 hr), and it appears this
sample is most representative of the daily average
(Kissel et al. 2005). Collection of repeated spot
urine samples during a single day or over several
days is one means of addressing the issue of intraindividual
variability. These repeated measures can be averaged
to produce a more stable estimate of exposure and
would allow evaluation of exposures during specific
windows of vulnerability.
Collection of complete 24-hr urine samples has
become a standard part of many occupational exposure
studies but has generally been viewed as impractical
for small children. Several studies reviewed here
have attempted to collect 24-hr samples but have
been only partially successful. A recent study (Kissel
et al. 2005) of 25 children in a low-income, low-literacy
population by the Berkeley center provided intensive
training of participants, detailed record keeping
by participants, use of small refrigerators, and
daily contact by research staff to improve compliance;
it was estimated that 28% of participants provided
complete samples, an additional 12% were likely complete,
52% missed one or two voids, and 8% likely missed
more than two voids.
Several of the centers have also collected blood
samples from children postnatally. The Columbia center
has employed a pediatric phlebotomist to draw blood
when children came to the center for the developmental
assessment. Samples were collected from 98% of the
children that were seen. However, volumes were generally
low (an average of 6.8 mL collected at 24 months
and 6.2 at 36 months). The Berkeley center hired
a pediatric phlebotomist to collect blood for both
state-required lead screening and the CHAMACOS (Center
for Health Analysis of Mothers and Children of Salinas)
study, increasing the rate of blood collection. Repeat
blood samples can be collected from young children
but are more difficult to obtain than are urine samples.
Children’s activities are an important variable
in assessing pesticide exposure. The Berkeley center
has used a visually based, low-literacy child activity
time line for parents to record child activity and
location. The University of Washington center and
the PNASH center have employed miniaturized global
positioning system (GPS) units to produce detailed
documentation of children’s time-location patterns
(Elgethun et al. 2003). Recent studies have found
that time-location diaries kept by parents produce
relatively poor agreement with the GPS measurements,
suggesting that such diary data would result in substantial
misclassification. The GPS analysis has also shown
that transient peak exposures can occur both temporally
and spatially and that such exposures are not adequately
captured within the resolution of most microenvironmental
analysis studies.
School-age children exposure. Sampling
procedures for school-age children are similar to
those described above for preschool children. However,
as children reach school age, they are more likely
to be able to participate more actively in studies.
They may be able to assent to study procedures, wear
personal sampling devices, collect more complete
urine samples, and provide helpful information regarding
pesticide sources and their own activities. Here
we would stress greater emphasis on personal sampling
devices to improve the quality of exposure data for
this age group.
Saliva monitoring. The PNASH center
has explored the feasibility of saliva sampling for
pesticides in both workers and children (Denovan
et al. 2000; Lu et al. 2003). Current saliva sample
collection methods require that children chew on
a cotton or synthetic plug for approximately 2 min.
The plug, containing up to 2 mL of saliva, is then
placed in a vial for storage. The plug is similar
in size to a dental sponge and could pose a choking
hazard to children < 3 years of age. The Berkeley
center has experimented with pipettes to directly
transfer saliva from a child’s mouth into a
collection container. Sample volumes, however, have
been < 1 mL. In rare cases, children have spit
directly into a beaker. It is not clear that these
techniques provide an adequate or appropriate saliva
sample for pesticide analysis.
Table
3
![Table 3](tab3sm.gif) |
Table 4
![Table 4](tab4sm.gif) |
Participation of cohort members in environmental
and biologic sampling. Collection of
an array of biologic and environmental samples
from women during pregnancy and soon after
birth places a burden on study participants
and may
lead to attrition regarding participation in
the exposure assessment component of these
studies. Tables 3 and 4 provide data from the
birth cohort
studies reviewed here to indicate what might
be anticipated in the National Children’s
Study. Sample sizes are presented for each
study and for each relevant time category;
the percentage
of enrolled study members is then provided
for each of the biologic or environmental samples.
It is important to recognize that not all of
the rates in Tables 3 and 4 are directly comparable.
For example, the Berkeley study accepted all
eligible enrollees with no condition that they
participate in every exposure assessment event;
in contrast, enrollment criteria for the Columbia
study included collection of a cord blood sample
from each participant at delivery. Participation
in environmental and biologic sampling tends
to drop over time and can be relatively low
for
certain types of samples. Factors contributing
to low participation include reliance on delivery
staff, emergency deliveries, inability to schedule
appointments that include both parents, mobile
populations that are hard to track, and the
absence of children from the home at the time
of visits
by study staff. Participation can also be enhanced;
for example, the Berkeley center saw an increase
from 64 to 81% between 12 and 24 months for
child blood samples because of the hiring of
a child
phlebotomist who went to each home.
Challenges in the Analysis of Pesticide Exposure
Samples
Increased interest in children’s exposure
to pesticides has resulted in the generation of large
numbers of samples for analysis. In this section
we discuss several key issues and lessons learned
regarding analysis.
Laboratory capacity. As studies of
the type described here grow larger and a series
of longitudinal samples are collected from each participant,
the sample size may become too large for the capacity
of one or two laboratories. Multiple laboratories
should be enlisted for large studies to avoid sample
backlogs. As laboratory capacity is improved, it
is imperative to produce comparable data across studies,
as the U.S. EPA did in its interlaboratory comparison
study among the North American laboratories performing
DAP analyses (James et al. 2003).
Intra- and interpersonal variability in urine
samples. Several methods have been evaluated
to “correct” for the variability
in urine dilution across spot samples, the most
popular being creatinine (Boeniger et al. 1993).
Creatinine excretion varies because of many factors,
including the size of the participant, so interindividual
variation, especially among diverse populations,
is large. Thus, creatinine-adjusted pesticide
concentrations should never be compared among
individuals of vastly different age groups (i.e.,
children vs. adults). Changes in creatinine excretion
during pregnancy should be thoroughly evaluated
before comparing with other women in similar
age groups. The validity of creatinine adjustment
may also be analyte dependent. Further studies
to assess the variability of commonly measured
analytes in urine should be conducted to identify
the most effective sampling strategies for cohort
studies. In all likelihood, sampling for nonpersistent
chemicals will require multiple samples taken
over the course of the study at regular intervals
(e.g., weekly, monthly, semiannually).
Selectivity of analysis. Selectivity
can refer to either the ability of a measurement
technique to differentiate a single analyte that
is measured from other components of the matrix (i.e.,
reducing false positives) or the ability of the analyte
measured to accurately, and unequivocally, identify
exposure to the target chemical of interest. However,
high selectivity techniques are costly and require
specialized training for operation (Barr et al. 1999).
Methods such as immunoassays and less specialized
technologies may be employed, but harmonization should
be performed to ensure that data generated using
different methods are comparable.
The selectivity of the analyte measured to accurately
reflect the exposure of interest may depend on the
biomarker being measured rather than the measurement
technique. Many OP pesticides, for example, can be
metabolized to common DAP compounds, so it is not
possible to derive chemical-specific exposure estimates
from such data. Further complicating the issue, the
DAPs, as well as compound-specific metabolites, may
be present in environmental media as the environmental
degradates of the pesticides (Curl et al. 2003b;
Wilson et al. 2004). No studies to date have shown
that these environmental degradates can be absorbed
and excreted unchanged; but if this does occur, then
DAPs and other pesticide metabolites detected in
urine would represent exposure to both the pesticide
and its degradate. Some metabolites are very selective
for the chemical measured. For example, 2-isopropoxy-4-methyl-6-hydroxypyrimidine,
a metabolite of diazinon, is selective for diazinon
exposure, although potentially the environmental
degradates could contribute to the urinary levels
as well. In some cases, the parent pesticide can
be excreted in urine, such as for the herbicide 2,4-D
(2,4-dichlorophenoxyacetic acid).
One way to unequivocally identify exposure to a
particular pesticide is by measuring the intact pesticide,
presumably in blood or similar samples, because the
intact pesticide is not appreciable in urine. However,
blood measurement levels are typically about 1,000
times lower than urinary metabolite measurements;
this requires highly sensitive analytical techniques,
driving up the cost of analysis. In addition, target
chemicals in blood may exhibit some degree of instability.
Finally, there are no laboratory methods available
for many common use agricultural or home pesticides
in blood. Saliva sampling is an attractive alternative
to blood sampling, as discussed above.
Sensitivity of analysis. The sensitivity
of an analytical method--the ability of the method
to measure the chemical at the desired level--should
be considered before a study begins (Barr et al.
1999). The biologic half-lives of nonpersistent chemicals
are relatively short, usually on the order of hours
or days (Needham and Sexton 2000). Samples collected
several days after an exposure event may require
ultrasensitive methods for analyte detection. These
measurements must provide adequate sensitivity to
allow detection of the chemicals of interest in a
sufficient proportion of the population to provide
a realistic representation of the populations’ exposure.
The current method for analysis of OP pesticide metabolites
developed by the Centers for Disease Control and
Prevention was used for many but not all of the studies
described in this article and has proven to be quite
sensitive (Bravo et al. 2002).
Alternative matrices and/or biomarkers. Pesticides
have been measured successfully in saliva (Lu et
al. 2003), meconium (Whyatt and Barr 2001), and amniotic
fluid (Bradman et al. 2003). Matrices such as meconium
may provide longer term dosimeters for exposure to
nonpersistent chemicals; saliva may provide a measure
of internal dose without the invasiveness of blood
sampling. Preliminary studies evaluating the partitioning
of chemicals in the various matrices should be conducted
that will allow for comparison of data among matrices
and validate the usefulness of alternative matrices
for biologic monitoring. An alternative matrix that
may prove useful is the gel matrix in disposable
diapers. Extraction techniques for solid materials
may prove practical for the gel matrix and might
improve sample collection procedures for infants
and children who are not toilet trained.
Quality assurance and control. A
vital component of all biomonitoring methodology
is a sound quality assurance/quality control (QA/QC)
program. QA/QC procedures supporting these studies
have included proficiency testing, repeat measurements
of known biologic materials, and round-robin studies
to confirm reproducible measurement values among
laboratories, as well as field spikes and field blanks
to confirm sample integrity.
Sample storage issues. The time
frame for sample testing and long-term storage becomes
an issue for large studies. The long-term stability
of analytes has been demonstrated for some matrices
but not for others, for example, blood. One final
logistical complexity is physical freezer space for
storage, and the substantial cost of maintaining
that storage. Archiving samples in the smallest containers
possible would enhance the ability to keep the samples
long term under proper storage conditions.
Epidemiologic investigations have often relied
on questionnaire information for exposure classification,
but this approach alone is unlikely to capture the
complexity of children’s pesticide exposure.
In contrast to the Agricultural Health Study, for
example, which draws on the records of pesticide
applicators and has derived a complex exposure algorithm
from 40 years of occupational exposure studies (Dosemeci
et al. 2002), the everyday use of pesticides in homes,
schools, and other child environments is not easily
codified, and dietary pesticide exposures can only
be inferred from questionnaire data. It seems, therefore,
that some level of environmental and/or biologic
monitoring will be required for all study participants.
The type of sampling needed will depend primarily
on the purpose of the study, be it exposure characterization,
long-term health outcomes, or short-term toxic response
in children. Lessons learned regarding pesticide
exposure can be summarized as follows:
-
Biologic monitoring appears to be the
best available method for assessment of children’s
exposure to pesticides. However, all pesticide
biomarkers have limitations. It is likely that
a combination
of biomarkers, environmental measurements, and
questionnaires will be needed after careful consideration
of the
specific hypotheses posed by investigators and
the limitations of each exposure metric.
-
Environmental measurements, such as
surface wipes and indoor air or house dust samples,
can characterize
residential pesticide contamination, but their
validity for exposure classification has not been
established.
Their value in epidemiologic studies deserves
further investigation.
-
Emphasis on personal rather than environmental
sampling in conjunction with urine or blood
sampling is likely to be most effective at classifying
exposure.
-
A focus on maternal exposures during
pregnancy is particularly important for making
associations
with infant health, given the critical developmental
stages during this period.
-
Questionnaires will need to be validated
with carefully designed studies that involve
personal sampling or biologic monitoring.
-
Interpretation of urinary metabolites
is not straightforward, but because of ease of
collection,
these samples may provide the best available
approach to capturing exposure variability of nonpersistent
pesticides in young children; additional validation
studies are needed.
-
Repeated exposure measures will be needed
to overcome high intraindividual variability
of biologic samples for most pesticides in use
today.
-
Postnatal exposure can also contribute
to health effects in early childhood. For infants
and
young children, it appears possible to collect
urine samples for extended periods of time.
-
Expansion of laboratory capacity will
require careful attention to QA/QC and will need
to include
formal procedures for ensuring interlaboratory
comparability in sample analysis.
-
Saliva measurements of pesticides, if
feasible, would overcome the limitations of urinary
metabolite-based
exposure analysis.
-
GPS technology appears promising in the
delineation of children’s time-location
patterns.
It is clear from this review that the critical
tools needed for accurate characterization of children’s
pesticide exposure are not yet in place. Most of
the work discussed here has been conducted in the
past 6-8 years, and many of the exposure methods
have been exploratory in nature. Substantial resources
will be needed for validation of existing methods,
support of novel methods, and enhancement of analytical
capabilities. It may be possible to initiate epidemiologic
investigations and validation studies simultaneously,
if biomarker samples can be properly archived. Whatever
sampling strategies are employed for epidemiologic
investigations, they will need to be selected to
support specific hypotheses and focus on specific
pesticides. Studies with substantial exposure assessment
activities will be costly but should ultimately pay
benefits in terms of the quality of scientific information
produced. |
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Last Updated: September 16, 2005 |
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