Science of the Total Environment 472 (2014) 289–295

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Review

Revisiting pesticide exposure and children's health: Focus on China Guodong Ding a,b, Yixiao Bao a,⁎ a b

Department of Pediatrics, Xinhua Hospital, Shanghai Jiao Tong University School of Medicine, Shanghai, China MOE and Shanghai Key Laboratory of Children's Environment Health, Xinhua Hospital, Shanghai Jiao Tong University School of Medicine, Shanghai, China

H I G H L I G H T S • • • • •

Pesticides are widely used in China and around the world. There is an increasing public awareness of pesticides and child health globally. China is still in the early stages and few epidemiological data are available. We have addressed some limitations and incompletenesses from current literature. These deficiencies should be interpreted with caution in future studies.

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Article history: Received 14 September 2013 Received in revised form 17 October 2013 Accepted 11 November 2013 Available online xxxx Keywords: Pesticide Organophosphate Pyrethroid Children Health China

a b s t r a c t China is now becoming the largest consumer of pesticides worldwide. In recent years, there has been a heightened public awareness of pesticides and children's health in North America and around the world. Human epidemiological studies have examined the relationship of pesticide exposures with children's health such as neurodevelopment and cancer, and they reported less consistent results. With regard to this topic, however, China is still in the early stages of cross-sectional or case–control design, and few data have been available. Furthermore, we have discussed several important limitations such as study design, exposure measurement, and developmental assessment from current literature, which should be interpreted with caution. We also presented the vulnerability and source of children's exposure to pesticides. © 2013 Elsevier B.V. All rights reserved.

Contents 1. Introduction . . . . . . . . . . . . . . 2. Pesticide use in China . . . . . . . . . 3. Children's vulnerability to pesticide . . . 4. Source of pesticide exposure . . . . . . 5. Profile of cohort studies in North America 6. Status of epidemiological studies in China 7. Current limitations and future prospects . 8. Conclusions . . . . . . . . . . . . . . Conflict of interest . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . .

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⁎ Corresponding author at: Department of Pediatrics, Xinhua Hospital, Shanghai Jiao Tong University School of Medicine, 1665 Kongjiang Road, 200092 Shanghai, China. Tel.: + 86 21 25078300. E-mail address: [email protected] (Y. Bao). 0048-9697/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.scitotenv.2013.11.067

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1. Introduction Pesticides are substances intended for preventing, destroying, repelling, or mitigating any pest. Pesticides are commonly referred to by their

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functional class, including insecticide, herbicide, fungicide, and various other substances used to control pests (US Environmental Protection Agency, 2012). China is one of the earliest countries to use pesticides. As early as in the Ming Dynasty (1378–1644 AD), the most wellknown Chinese herbal “Ben Cao Gang Mu”, compiled by Li Shizhen, had recorded a number of plants and minerals that used as pesticides such as veratridine, arsenolite, and lime (W. Zhang et al., 2011). Nowadays, the use of pesticides worldwide is increasing gradually as human population and food requirements increase. Meanwhile, the overuse and misuse of pesticides have become a major problem globally, especially in developing countries like China, which increase related environmental and health risks. A 2002 study for the Global Greengrants Fund revealed that as much as 40% of pesticides on the market in China are sold under false brand names (Pesticide Action Network North America, 2003). Infants and children have been identified by both the US National Research Council and International Life Sciences Institute as groups within the population who require special consideration in risk assessment because of their unique exposure patterns and special vulnerabilities to environmental hazards (Landrigan et al., 2004). Therefore, there is an urgent need for research to fill the gaps of information on exposures and health consequences of pesticide exposures to the fetus and children. The purpose of this article is to a) revisit the pattern and amount of pesticide use in China; b) present the vulnerability and source of children's exposure to pesticides; c) summarize the epidemiological studies from North America and China regarding exposure to pesticides and children's health consequences; and d) highlight some intrinsic limitations and incompletenesses of evidence from current literature which should be taken into consideration in future studies. 2. Pesticide use in China Worldwide, over 5.2 billion pounds of pesticides is applied each year, and roughly 85% of the pesticides currently used are devoted to the agricultural sector (Grube et al., 2011). Insecticides, herbicides, and fungicides are the most common types of conventional pesticide (Grube et al., 2011). Almost all countries make general use of the major pesticides, but they each use specific pattern of pesticides in correspondence with their typical crop. As shown in Fig. 1, 35% of the pesticide used is insecticide in China, as against 8% and 17% in the U.S. and the world, respectively. The use of herbicides and fungicides is correspondingly less heavy (Ministry of Environmental Protection of China, 2008; Grube et al., 2011). Although China has experienced a rapid urbanization transition over the past several decades, it is still a large developing agricultural country with agriculture accounting for more than 12% of the total gross domestic product (Economy Watch, 2010). In China, the majority of conventional pesticide use is dedicated to agriculture, while the remainder is used in residential, commercial, and industrial settings (W. Zhang et al., 2011). Pesticides

can also be grouped by their chemical class. Among these, organochlorine (OC) pesticides such as DDT were widely used in agriculture and pest control between the 1940s and 1970s. However, concern over their persistence in the environment and tendency to bioaccumulate led the Chinese government to restrict or ban their use during the 1980s (W. Zhang et al., 2011). Consequently, insecticides such as organophosphates (OPs) and pyrethroids (PYRs) have become attractive alternatives to OC pesticides because they do not persist in the environment. In the last two decades, the use of OP pesticides has been responsible for nearly 30% of total pesticide applications nationwide (He, 2008), it indicates that OP pesticides play important roles in controlling insects, weeds, and diseases on farms and in urban landscapes. The amount of pesticide use in China has been dramatically increased in recent years (Fig. 2), and China is now becoming the largest consumer of pesticides in the world. More than 1.6 million tons of pesticides (roughly 3.5 billion pounds) are used annually in China (National Bureau of Statistics of China, 2011). However, approximately 1.1 billion pounds of pesticides were applied to the U.S. each year (Grube et al., 2011). Indeed, increasing food production has long been the priority of the Chinese government to meet food requirements for its large population. Under the adage “if little is good, more is better”, many farmers overuse or improper use pesticides to get greater yields, which in turn, contribute to the excess residues of agricultural crops and the loss of agricultural land. In China, organophosphates are currently the most heavily used pesticide in agriculture whereas pyrethroids are the most common class of pesticide used in homes. 3. Children's vulnerability to pesticide The primary benefits are the consequences of the pesticide's effects — the direct gains expected from their use. Without pesticide uses in China, the production of fruits, vegetables, and cereals would lose 78%, 54%, and 32%, respectively (Cai, 2008). Although the benefits of pesticides are well recognized, their potential adverse effects on human health remain unclear. Historically, health risk assessment focused on adult exposure and toxicity and gave little consideration to vulnerable life stages such as fetal development and early childhood (Landrigan et al., 2004). The hazard of pesticides to children's health has been the subject of great concern globally since the publication of the report “Pesticides in the Diets of Infants and Children” by the National Academy of Sciences (U.S. NAS) in 1993 (National Research Council, 1993). Evidence has shown that pesticide from the mother can readily pass through the placenta and transfer to the fetus, making the fetus susceptible to pesticide poisoning (Bradman et al., 2003). Even low-level pesticide exposure early in life, at levels that may not harm the pregnant mother, could impair infant growth and development. Fetus and young children may be more susceptible to the potentially toxic effects of pesticides, not only because their organ systems, specifically the brain and central nervous system, are developing rapidly but also because they have lower levels of

Fig. 1. Consumption pattern of pesticides (Grube et al., 2011; Ministry of Environmental Protection of China, 2008).

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Fig. 2. Amount of pesticides used in China between 2007 and 2011 (National Bureau of Statistics of China, 2011).

detoxifying enzymes that deactivate pesticides than adults (Eskenazi et al., 2007). Furthermore, pound for pound of body weight, young children drink more water, eat more food, and breathe more air than adults. In addition, they also spend more time playing and crawling on the floor where pesticides may settle and have increased nondietary ingestion through frequent hand-to-mouth and object-to-mouth contacts (Eskenazi et al., 2007). All these factors indicate that fetus and young children are not “little adults”, their unique physiological characteristics and age-related behaviors leave them particularly susceptible and vulnerable to the health effects of pesticide exposure. 4. Source of pesticide exposure As stated above, children's exposure to pesticides may be linked to multiple sources (e.g., air, food, and dust) and multiple routes (e.g., inhalation, ingestion, and dermal absorption) that differ from adults (Garry, 2004; Council On Environmental Health, 2012). A biologic monitoring survey performed among preschool children in the Seattle metropolitan area showed that no significant differences were found in urinary levels of OP metabolites related to season, community, gender, age, family income, or housing type (Lu et al., 2001). The authors speculated that diet might be a significant source of OP exposure for these children. Curl et al. (2003) assessed OP pesticide exposure from diet by biological monitoring among urban and suburban preschool children. They collected 24-h urine samples from 18 children with organic diets (produced without pesticide) and 21 children with conventional diets and examined for five OP metabolites. Children with organic diets had significantly lower urinary levels of dimethyl metabolites than children with conventional diets (Curl et al., 2003). Similarly, an intervention study that substituted children's conventional diets with organic food items observed drastic and immediate decrease in urinary excretion of pesticide metabolites (Lu et al., 2006). Thus, diet is likely to be a predominate source in the OP exposure of urban and suburban children. In agricultural setting, pesticide spray drift from pesticide-treated fields to nearby residences or children's playground seems to be an important source of pesticide exposures to rural children. Koch et al. (2002) measured OP pesticide exposures of young children living in an agricultural community over an entire year and evaluated the impact of agricultural spraying on exposure. Urinary dimethyl and diethyl dialkylphosphate (DAP) levels were significantly higher during spray months compared to nonspray months. They observed no differences for age, parental occupation, or residential proximity to fields. This study indicated that the ultimate source of exposure of children was the result of airborne pesticide dispersal from multiple sites within the agricultural region (Koch et al., 2002). The take-home exposure on clothing and footwear of farm workers could be another significant

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pathway in rural children's exposures to pesticides. Curl et al. (2002) conducted a study of OP exposure of 213 farm workers and 211 children that house dust contains on average 0.53 μg/g azinphosmethyl (an OP) and 0.75 μg/g in vehicle dust. Dimethyl DAP, a metabolite of the OP, in urine from parents (0.09 μmol/g) and children (0.14 μmol/g) was nearly identical. Dimethyl DAP levels in child and adult urine from the same household were also positively associated (Curl et al., 2002; Garry, 2004). However, in regard to other classes of pesticides such as PYR pesticides (the most commonly used household pesticides), little information was available on the source and amount of children's exposure. Whether this pattern of children's exposure to OP pesticides in North America can extend to Chinese children is uncertain. A recent study in China found that among the 2520 milled rice samples examined, 235 (9.3%) contained detectable residues of at least one of the seven target OP pesticides (chlorpyrifos, dichlorvos, omethoate, methamidophos, parathion-methyl, parathion, and triazophos), with levels ranging from 0.011 to 1.756 mg/kg, and 165 samples (6.5%) exceeded the national maximum residue limits (0.05 mg/kg for triazophos and omethoate, 0.1 mg/kg for the other five OPs listed above) (Chen et al., 2009). Given that the majority of pesticides are dedicated to agriculture in China, we speculate that diet may be the most influential source for Chinese children. 5. Profile of cohort studies in North America In recent years, there has been a heightened public awareness of pesticides and children's health in North America and around the world. A number of animal studies have demonstrated that exposure to pesticides during gestation or the early postnatal period could adversely affect offspring's developmental outcomes (Eskenazi et al., 1999). However, a few epidemiological studies have investigated the possible relationship of exposure to pesticides with children's health (e.g., neurobehavioral development, fetal growth, and childhood cancer), and these studies reported either similar or inconsistent results. The literature from Columbia University (the New York City cohort), University of California at Berkeley (the CHAMACOS cohort), and Mount Sinai School of Medicine (the Mt. Sinai cohort) was the most interesting and impressive. For example, both the CHAMACOS and Mt. Sinai birth cohorts suggested that prenatal DAP metabolite levels in urine were positively associated with numbers of abnormal reflexes as measured in the Brazelton Neonatal Behavioral Assessment Scale (BNBAS). No associations were found between urinary DAP metabolite levels and any other clusters of the BNBAS (Young et al., 2005; Engel et al., 2007). Although similar results across studies have emerged in neonatus, the results are less consistent in older infants or children. Researchers from Columbia University revealed that cord blood levels of chlorpyrifos, which metabolize to diethyl DAP levels, were responsible for the observed adverse associations with child neurodevelopment according to the Bayley Scales of Infant Development (BSID) (Rauh et al., 2006). In contrast, researchers from University of California at Berkeley found that child neurodevelopment was adversely associated with prenatal dimethyl DAP instead of diethyl DAP levels (Eskenazi et al., 2007). These inconsistent results across studies may arise from differences in exposure scenarios and in the method of exposure assessment (Rosas and Eskenazi, 2008). Nonetheless, growing evidence supports the adverse association between exposure to pesticides and children's health consequences. 6. Status of epidemiological studies in China Pesticides are widely and heavily used in agriculture throughout China, concerns regarding the adverse health effects of exposure to pesticides among children are increasing. However, most of the previous studies performed in developing countries like China focused on workers with occupational exposure. Little information on non-occupational

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exposures has been available in China, and several studies regarding pesticide exposures and children's health are still in the early stages of crosssectional or case–control design (Table 1) (Y. Zhang et al., 2011; Qi et al., 2011; P. Wang et al., 2012; Ding et al., 2012b; L. Wang et al., 2012; Ding et al., 2012a). Our study group has conducted two cross-sectional studies to investigate OP urinary DAP levels from pregnant women and young children in Shanghai, and to examine their relationships with birth outcomes and child neurodevelopment, respectively. Both pregnancy

women and children displayed substantially higher levels of OP urinary metabolites compared with corresponding population reported in developed countries (P. Wang et al., 2012; Ding et al., 2012b). Increased prenatal individual diethylphosphate (DEP) levels were related to decreased gestational duration in female infants, but no associations were observed between individual or total DAP levels and birth weight or length in all infants or with gestational duration in male infants (P. Wang et al., 2012). Similarly, children's concurrent DAP levels were not found to be

Table 1 Pesticide exposures and children's health consequences from epidemiological studies in China. Reference/ Study design definition of cases/outcomes authors P. Wang et al. (2012)

Ding et al. (2012b)

L. Wang et al. (2012)

Ding et al. (2012a)

Y. Zhang et al. (2011)

Qi et al. (2011)

Cross-sectional: n = 187 mother–newborn pairs. Pregnant women (18–45 years of age): recruited from the departments of obstetrics & gynecology of two municipal hospitals in Shanghai, between September 2006 and January 2007. Birth outcomes (birth weight, length, length of gestation): collected from the mothers' and infants' medical records after delivery. Cross-sectional: n = 301 children. Young children (2 years of age): recruited from the departments of child and adolescent healthcare of two community hospitals in Shanghai between February and October 2008. Children's developmental quotients (DQs): measured by the Gesell Developmental Schedules (including motor, adaptive, language, and social scores). Cross-sectional: n = 503 mother–newborn pairs. Pregnant women (15–46 years of age): recruited from the department of obstetrics & gynecology of the Ling County People's Hospital of Dezhou City in the Shandong Province from November 2009 to February 2010. Birth weight (g): obtained from medical charts. Low birth weight: defined as b2500 g. Case–control: n = 176/180. Case children (0–14 years of age) with acute lymphocytic leukemia (ALL): recruited from the departments of hematology and oncology of four children's hospitals located in Shanghai. Control subjects: selected from the departments of child and adolescent healthcare and matched to the ALL patients based on gender, age, and hospital at the time of diagnosis. Case–control: n = 80/96. Case children (b15 years of age) with acute leukemia (AL) were recruited from the department of hematology and oncology of Shanghai Children's Medical Center. Control subjects: healthy children who attended the outpatient clinic for health examination and matched to the AL patients based on gender and age. Longitudinal: n = 301 mother–newborn pairs. Pregnant women (18–45 years of age): recruited from the Sheyang County Maternity Hospital in the Jiangsu province from June 2009 to January 2010. One-year-old infants' DQ and mental index (MI) scores: measured by the Denver Developmental Screen Test II.

Exposure assessment (pesticides identified)

Main outcomes

Other covariates considered

DAP metabolite levels (DMP, DMTP, DEP, DETP, and DEDTP) in maternal urine during hospital admission for delivery (within 3 d before delivery).

Individual DEP level: duration of gestation Gestational age, maternal height, pregnancy weight (weeks): ß = −1.79; p = 0.001 in gain, and family income. female infants. No associations were found between any other or total DAP levels with birth outcomes.

Dimethyl DAP (DMP and DMTP), diethyl DAP No evidence of decreased DQ scores with increased concurrent urinary (dimethyl, (DEP, DETP, and DEDTP), and total DAP diethyl, or total) DAP metabolite levels. metabolite levels in urine.

Child sex, maternal education level, and household income.

Maternal interview via questionnaire shortly Exposure to pesticides during pregnancy was associated with a non-significant after delivery regarding pesticide-related increase in low birth weight. exposures during pregnancy. Exposure to pesticide use was classified into three categories: never exposed to pesticides; exposed to pesticides when not pregnant; and exposed to pesticides during pregnancy.

Maternal age, race, education, menarche (first menstruation) age, gestational age, child sex, and place of residence.

Urine samples were collected at the time of the interview in the hospitals. Individual and total pyrethroid metabolite levels (cis-DCCA, trans-DCCA, and 3-PBA) levels in urine.

Compared with those in the lowest quartiles of total and individual metabolites, the highest quartiles: total metabolites: OR = 2.75 (1.43–5.29); cis-DCCA: OR = 2.21 (1.16–4.19); trans-DCCA: OR = 2.33 (1.23–4.41); 3-PBA: OR = 1.84 (1.00–3.38).

Child age, sex, household income, parent education, place of residence, and breast-feeding duration.

Urine samples were collected at the time of the interview in the hospital. Individual DAP metabolite levels (DMP, DMTP, DEP, DETP, and DEDTP) in urine.

The four individual DAP metabolite levels in cases were significantly higher than those in controls. DMP: Z = −5.75, p b 0.001; DMTP: Z = −2.80, p = 0.005; DETP: Z = −3.32, p = 0.001; DEDTP: Z = −8.12, p b 0.001.

–

All pregnancy women were asked to provide urine collections prior to delivery. Total pyrethroid metabolite levels (sum of the cis-DCCA, trans-DCCA, and 3-PBA) levels in urine. Using the 25th and 75th of total metabolite levels as cutpoints to form the low, median, and high exposure groups.

– DQ scores were highest (90.8 ± 11.4) in the low exposure group, followed by the median group (86.4 ± 14.6), and DQ scores (83.7 ± 10.2) were lowest in the high group. There were no significant differences in MI scores among the three groups.

DAP, dialkyl phosphate; DMP, dimethylphosphate; DMTP, dimethylthiophosphate; DEP, diethylphosphate; DETP, diethylthiophosphate; DEDTP, diethyldithiophosphate; cis-DCCA, cis-3(2,2-dichlorovinyl)-2,2-dimethylcyclopropane carboxylic acid; trans-DCCA, trans-3-(2,2-dichlorovinyl)-2,2-dimethylcyclopropane carboxylic acid; 3-PBA, 3-phenoxybenzoic acid.

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associated with neurodevelopment based on the Gesell Developmental Schedules (GDS) (Ding et al., 2012b). We also have conducted a case– control study to examine the association between the urinary levels of PYR pesticide metabolites and the risk of childhood acute lymphocytic leukemia (ALL). Compared with those in the lowest quartiles of total and individual PYR metabolites, the highest quartiles were associated with an approximate 2–3 fold increased risk of ALL (Ding et al., 2012a). However, our results were limited by the study design and small sample size, and more studies of pregnant women and children living in China are warranted given the relatively high levels of exposure to pesticides. 7. Current limitations and future prospects In view of the intrinsic limitations and incompletenesses of scientific evidence from current human epidemiological studies, determining a link between in utero or early exposure to pesticides and children's health risks poses a monumental challenge. The use of improved methodology for both study design and exposure or outcome assessment will benefit the value of future epidemiological studies to public health. First and most importantly, many epidemiological studies fail to cover probable weak but significant associations between developmental impairments as a result of exposure to two or more classes of pesticides. Improvements in analytical equipment and testing procedure, such as the development of non-invasive sampling methods, testing for pesticides and their metabolites in human urine, have made it easy to monitor pesticide exposure even at very low levels in infants and children (Colborn, 2006). Thus, it is not safe to say that each child conceived today in developing countries like China is not exposed to OC pesticides from conception throughout gestation and lactation, although OC pesticides have been banned for many years throughout China. In general previous experimental models regarding the toxicological risks are performed using single pesticide compounds; however, in practice, human exposures have become too complex because of the hundreds of formulations and mixtures of pesticides on the market. This subject area is particularly important in China because pesticide mixtures of OP and PYR for agriculture have become far more prevalent, and multiple pesticide exposure problems are associated with rapid development and a large population. The prevalence of acute poisoning in Chinese farmers with exposure to combined pesticides and single OP pesticides was 10.1% and 2.3%, respectively, and the number of cases of acute poisoning caused by exposure to OP and PYR mixtures in the late 1990s was about 4 times greater than the number reported in the late 1980s (He and Chen, 1999; Yáñ ez et al., 2002). Furthermore, recent epidemiological studies also suggest that both OP and PYR pesticide exposures are widespread in Chinese pregnant women and children (P. Wang et al., 2012; Ding et al., 2012a,b; Qi et al., 2012; Wu et al., 2013). Therefore, when we embark on determining the health risks of contemporary-use pesticides (e.g., OPs), the impacts of transgenerational exposure on human health should be meticulously inventoried, especially the new pesticides coming on the market (e.g., PYRs). Second, the measurement of nonspecific metabolites of OP and PYR pesticides in urine is the most current method to characterize and integrate exposure to multiple pesticides that originate from different sources (Heudorf and Angerer, 2001; Eskenazi et al., 2004; Barr et al., 2010). Despite this exposure assessment method of identifying special chemicals has progressed from crude surrogates such as living in a rural area or self-reported pesticide use, the two classes of pesticides are readily metabolized, and exposure can vary considerably and most often is transient and unpredictable (Heudorf and Angerer, 2001; Eskenazi et al., 2004; Barr et al., 2010). Also, nonspecific metabolite measurements do not allow differentiation between exposures that result from more or from less toxic pesticides (Eskenazi et al., 2004). In addition, nonspecific metabolites may reflect exposure not only to

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pesticide parent compounds, but also to the potentially nontoxic preformed metabolites in the environment (Ding et al., 2012b). It is important to keep in mind that we are limited in our ability to understand the average cumulative dose from different sources and to what extent these measurements accurately reflected the exposure throughout the entire critical period of child development (Ding et al., 2012b). Although existing biomarkers such as urine and blood measurements can be useful in understanding the role of environmental pollutants during pregnancy, infancy, and childhood, they provide short-term dosimeters only (Whyatt and Bar, 2001). Research on the effects of prenatal pesticide exposure is limited and has been hampered partly because of the lack of biomarkers reflecting cumulative exposures. Measurements of pesticide nonspecific metabolites in amniotic fluid and meconium have been proposed frequently as better biomarkers reflecting cumulative exposures (Whyatt and Bar, 2001; Bradman et al., 2003). In human, the fetus is continuously swallowing and inhaling amniotic fluid which is present soon after implantation, this cycling indicates that toxicants excreted into the amniotic fluid may continuously reexpose the fetus (Bradman et al., 2003). Meconium begins to accumulate in the bowels at approximately 16 weeks gestation and is generally not excreted until after delivery, and thus probably represents a longer-term dosimeter of prenatal exposure (Whyatt and Bar, 2001). Measurements of pesticide nonspecific metabolites in amniotic fluid and postpartum meconium may not fully reflect average cumulative dose, but exploration of this hypothesis in human epidemiological studies is necessary. Third, concerns with regard to developmental neurotoxicity due to pesticides have been fuelled by recent epidemiological observations that children exposed during pregnancy or early postnatal life suffer from various neurological deficits. Unlike obvious birth defects, most neurodevelopmental effects could not be observed directly at birth or even later in life. Instead, adverse effects on the brain and nervous system are expressed in terms of how an individual behaves or functions (Colborn, 2006). Indeed, there is often an overlap in behavioral characteristics among different neurodevelopmental disorders such as hyperactivity in attention deficit hyperactivity disorder and autism (Connors et al., 2008). A good functional impairment measure tool should be standardized, and reproducible, and have high sensitivity and specificity. Functional impairment in laboratory animal models is measured at the gene, cell, biochemical, and physiologic levels, and further requires high-tech instrumentation to specify and quantify. Neurological deficits are not “on” and “off” conditions but instead range from mild, barely noticeable effects, to severe, in which the individual has extremely poor motor skills and mental retardation (Colborn, 2006). Therefore, it is difficult to quantify and compare neurodevelopmental impairment. With the introduction of toxicogenomics and the development of psychological and mental assessment tools with increasing specificity and sensitivity at the human level, we believe that we will better characterize and quantify functional impairment. Fourth, a number of epidemiological investigations have also examined the relationship between exposure to pesticides and risk of childhood cancer, but they reported less consistent results. These differences were probably due to several inherent limitations such as retrospective case–control design and relatively small number of case children, as cancer is a rare occurrence and prospective studies are difficult to conduct (Belson et al., 2007). Many existing studies within a few hundred case children could not have adequate statistical power to identify cancer risks to specific pesticides, and null results from these studies should not be interpreted as evidence that pesticides pose no risk to children's health. Case–control studies are critical for exploring the association of pesticide exposures with childhood cancer, and could provide the opportunity to collect confounder information and biological samples in a costefficient manner (Alavanja et al., 2004). However, recall bias and general recall issues remain an important consideration of all case–control studies. The design that case–control study nested in cohort study with prospectively collected exposure information may be the best approach,

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but in practice cohort studies are expensive and time consuming for rare diseases (Alavanja et al., 2004). We can minimize recall bias by using detailed questionnaires that ask for specific time periods of exposure and doses of exposure, and by giving appropriate attention to confounding factors and effect modifiers while designing studies. Given the multifactorial development of childhood cancer, we need to conduct a large-scale multicenter study that is capable of providing sufficient power to detect even modest associations with precision, and need to build a diagram of known and proposed causal relationships in order to take all potential confounders and effect modifiers into consideration (Buffler et al., 2005). Finally, it should be noted that genetic variation may influence children's susceptibility to pesticide exposures. Individuals with specific genetic polymorphisms leading to decreased activity or expression of paraoxonase 1 (PON1) or cytochrome P450s might be expected to be at greater risk of ill-health following exposure (Dai et al., 2001; Costa et al., 2003; Povey, 2010). Recently, polymorphisms in PON1 were associated poorer developmental quotient scores among young Mexican-American children from the CHAMACOS birth cohort (Eskenazi et al., 2010). How pesticide exposure and these genes interact and then contribute to adverse health outcomes is yet poorly understood, and the role of genetic variation in pesticide toxicity is a new and promising area that needs further attention and better understanding. 8. Conclusions To sum up, there is a great deal of uncertainty about the adverse effects of exposure to pesticides on children's health. Despite the weight of evidence suggests that exposure to pesticides is probably associated with adverse children's health, the relationship is not considered causal at least in part because of the intrinsic limitations and incompletenesses of available epidemiological evidence. Given the importance of pesticides in agriculture, public health, and homes, they will continue to be used and will therefore be present in the human environment. Improving epidemiological study design and exposure assessment technique, and integrating this information with existing toxicological data, will allow the children's health risks of pesticide exposures to be more accurately examined. We are gratified that the Ministry of Agriculture of China announces an overall 20% reduction in pesticide use within the next five years. In the meantime, it is important to take precautions to minimize children's exposure to pesticides wherever possible. Conflict of interest None declared. References Alavanja MC, Hoppin JA, Kamel F. Health effects of chronic pesticide exposure: cancer and neurotoxicity. Annu Rev Public Health 2004;25:155–97. Barr DB, Olsson AO, Wong LY, Udunka S, Baker SE, Whitehead RD, et al. Urinary concentrations of metabolites of pyrethroid insecticides in the general U.S. population: National Health and Nutrition Examination Survey 1999–2002. Environ Health Perspect 2010;118:742–8. Belson M, Kingsley B, Holmes A. Risk factors for acute leukemia in children: a review. Environ Health Perspect 2007;115:138–45. Bradman A, Barr DB, Claus Henn BG, Drumheller T, Curry C, Eskenazi B. Measurement of pesticides and other toxicants in amniotic fluid as a potential biomarker of prenatal exposure: a validation study. Environ Health Perspect 2003;111: 1779–82. Buffler PA, Kwan ML, Reynolds P, Urayama KY. Environmental and genetic risk factors for childhood leukemia: appraising the evidence. Cancer Invest 2005;23:60–75. Cai D. Understand the role of chemical pesticides and prevent misuses of pesticides [in Chinese]. Bull Agric Sci Technol 2008;1:36–8. Chen C, Li Y, Chen M, Chen Z, Qian Y. Organophosphorus pesticide residues in milled rice (Oryza sativa) on the Chinese market and dietary risk assessment. Food Addit Contam Part A Chem Anal Control Expo Risk Assess 2009;26: 340–7.

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