Meeting of the FIFRA 

Scientific Advisory Panel

Draft Issue Paper:

Scientific Issues Concerning Health 

Effects of Chlorpyrifos 

April 10-13, 2012
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Table of Contents

1.0 INTRODUCTION 5

2.0 SOURCE-TO-OUTCOME CONSIDERATIONS 7

2.1 Exposure Pathways for Chlorpyrifos 7

2.2 Conceptual Framework 9

2.2.1. Adverse Outcomes 10

2.2.2. Potentially Sensitive Lifestages 11

2.2.3. Critical Duration(s) of Exposure 14

3.0 ADVERSE OUTCOME PATHWAYS: ACHE INHIBITION & PLAUSIBLE PATHWAYS 

LEADING TO NEURODEVELOPMENTAL OUTCOMES 15

3.1 Adverse Outcome Pathway: AChE Inhibition 15

3.1.1 Initiating Event & Health Outcomes 15

3.1.2 Dose Response Analysis for AChE Inhibition 16

3.1.3 Summary of the Dose Response Analysis 21

3.2 Adverse Outcome Pathway: Neurodevelopmental Outcomes 21

3.2.1 Biologically Plausible Toxicity Pathways Leading to Neurodevelopmental Outcomes 22

3.2.1.1 Acetylcholinesterase (AChE) as a morphogen 22

3.2.1.2 Cholinergic system 25

3.2.1.3. Endocannabinoid system 27

3.2.1.4. Reactive Oxygen Species 29

3.2.1.5. Serotonergic system 30

3.2.1.6. Tubulin, Microtubule Associated Proteins and Axonal Transport 33

3.2.1.7 Summary/Conclusions 35

3.2.2 Adverse Outcomes in Laboratory Animals: Effects on the Developing Brain 36

3.2.2.1 Summary of Existing Database 37

3.2.2.2 Developmental Impacts on Neurological Domains 39

3.2.2.2.1 Cognitive Function 39

3.2.2.2.2 Anxiety/Emotion 43

3.2.2.2.3 Social Behaviors/Interactions 46

3.2.2.2.4 Motor activity 48

3.2.2.3 Dose-Response Considerations 51

3.2.2.4 Conclusions 52

4.0 REVIEW OF CHLORPYRIFOS EPIDEMIOLOGY REGARDING CHILDREN’S HEALTH 54

4.1 Scope and Purpose 54

4.2 Summary of Epidemiology Findings 55
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4.2.1 Overview of Design and Methods of Children’s Health Studies 56

4.2.1.1 The Mother’s and Newborns Study of North Manhattan and South Bronx (Columbia University) 56

4.2.1.2 Inner-City Toxicants, Child Growth and Development Study (Mt. Sinai Hospital) 57

4.2.1.3 Center for Health Assessment of Mother’s and Children of Salinas Valley, CHAMCOS (University of 

California/Berkley) 57

4.2.2 Summary of Research Results 58

4.2.2.1 Measures of Fetal Growth 58

4.2.2.2 Brazelton Neonatal Neurological Functioning 61

4.2.2.3 Bayley Scale of Mental and Psychomotor Development 62

4.2.2.4 Attention Problems 63

4.2.2.5 Intelligence Measures 64

4.2.3 Strength and Limitations of Children’s Health Cohort Studies 65

4.2.3.1 Epidemiologic Inference of Combined Children’s Health Cohorts 70

4.3 Chlorpyrifos Epidemiology Synthesis and Evaluation 71

4.4 Conclusions 75

5.0 EXPOSURE PROFILE & BIOMONITORING RESEARCH 77

5.1. Overview of Biomarkers and Scientific Considerations 77

5.2 Chlorpyrifos Biomonitoring Study Comparison 79

5.3 Chlorpyrifos Biomonitoring and AChE Inhibition 87

5.4 Approaches to Interpreting Biomonitoring Data 90

5.4.1 Consideration of Approaches for Interpretation of Biomonitoring Data 90

5.4.2 Physiologically Based Pharmacokinetic-Pharmacodynamic Model for Chlorpyrifos: State of the Science 93

5.4.2.1 Lifestages considered in the current chlorpyrifos PBPK-PD model 93

5.4.2.2 Routes of Exposure simulated by the PBPK-PD model 95

5.5 Conclusions 95

6.0 SUMMARY & PRELIMINARY CONCLUSIONS 96

7.0 REFERENCES 104
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Table of Figures

Figure 1 Source to effects pathway (adapted from NRC, 2007). 6

Figure 2 Conceptual framework considering chlorpyrifos exposure, lifestages, critical duration, modes of 

action/adverse outcome pathways, and health outcomes of interest. 10

Figure 3 Generic Adverse Outcome Pathway (Adapted from Ankley et al., 2010) 15

Figure 4 Summary of Comparison of TCPy Biomonitoring Studies Involving Child Study Populations.a 83

Figure 5 Summary of Comparison of TCPy Biomonitoring Studies Involving Adult Study Populations.a 84

List of Tables

Table 1. Results of BMD Modeling of Rat Pup Brain and Blood ChE Inhibition following a Single Oral Dose of 

Chlorpyrifos 19

Table 2. Results of BMD Modeling of Adult Rat Brain and RBC AChE Inhibition following a Single Oral, Dose of 

Chlorpyrifos 19

Table 3. Results of BMD Modeling of Pup Rat Brain and RBC AChE Inhibition following 11-days Repeated Oral, 

Gavage Doses of Chlorpyrifos 20

Table 4. Results of BMD Modeling of Adult, Female Rat Brain, RBC and Heart Cholinesterase Inhibition following 

Repeat Oral Doses of Chlorpyrifos 20

Table 5 Summary of Developmental Chlorpyrifos Studies by Laboratory 38

Table 6 Summary of Outcomes on Cognitive Tests in Male (M) and Female (F) Rodents 42

Table 7 Summary of Anxiety/Emotion Outcomes in Male (M) and Female (F) Rodents 45

Table 8 Summary of Social/Interactive Behavior Outcomes in Male (M) and Female (F) Mice 47

Table 9 Summary of Motor Activity Outcomes in Male (M) and Female (F) Rodents 50

Table 10 Summary of Findings in the Columbia Mothers and Newborn Cohort, the Mount Sinai Hospital 

Children’s Environmental Health Cohort, and the UC Berkeley Center for the Health Assessment of Mothers 

and Children of Salinas Studies of Prenatal Organophosphate Pesticide Exposure and Child Development.59

Table 11 Common Chlorpyrifos Biomarkers of Exposure. 77

Table 12 Summary Results of TCPy Biomonitoring Studies (Modified from Eaton et al., 2008) 85

Table 13 : Comparison of Creatinine-Adjusted Urinary TCPy levels (μg /g creatinine) reported by Galabrant et al. 

(2009) and Farahat et al. (2011) with the General U.S. Population. 88

Table 14 Comparison of Approaches for Interpretation of Chlorpyrifos (CPF) Biomonitoring Data 92
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1.0 Introduction 

Chlorpyrifos (0,0-diethyl-0-3,5,6-trichloro -2-pyridyl phosphorothioate) is a broad-spectrum, chlorinated 

organophosphate (OP) insecticide. In 2000, nearly all residential uses were voluntarily cancelled by Dow 

AgroSciences but agricultural use remains. The 2000 human health risk assessment was largely based 

on adult laboratory animal data (rat or dog) for cholinesterase (ChE) inhibition and the application of 

default uncertainty factors (U.S. Environmental Protection Agency, 2000). In 2008, the Federal 

Insecticide, Fungicide and Rodenticide Act (FIFRA) Scientific Advisory Panel (SAP) reviewed a draft 

science issue paper on the human health effects of chlorpyrifos which provided a preliminary review of 

the scientific literature on experimental toxicology and epidemiology studies available since the 2000 

risk assessment. This draft issue paper considered a growing body of literature with laboratory animals 

(rats and mice) indicating that gestational and/or early postnatal exposure to chlorpyrifos may cause 

persistent effects into adulthood in addition to epidemiology studies evaluating prenatal chlorpyrifos 

exposure in mother-infant pairs that have reported associations with birth outcomes, childhood 

neurobehavioral and neurodevelopment outcomes in the offspring when evaluated in neonates, infants, 

and young children. In 2008, the SAP agreed with the Agency that although the epidemiology studies 

were high quality with respect to design, conduct and analyses and provided information for hazard 

characterization, ChE inhibition remained the most robust and sensitive data for deriving points of 

departure. In 2011, the Agency released a preliminary human health risk assessment for chlorpyrifos. 

The focus of the 2011 risk assessment was on the ChE inhibiting potential of chlorpyrifos including in 

young animals. Like other OPs, chlorpyrifos binds to and phosphorylates the enzyme 

acetylcholinesterase (AChE) in both the central (brain) and peripheral nervous systems. This can lead to 

accumulation of acetylcholine and, ultimately, at sufficiently high doses, to clinical signs of toxicity. 

Consistent with the focus on ChE inhibition, in the 2011 preliminary risk assessment, EPA evaluated the 

extensive database of ChE data for multiple lifestages and selected points of departure based on 

consideration of all quality and reliable data. 

In 2010, the Agency developed a draft “Framework for Incorporating Human Epidemiologic & Incident 

Data in Health Risk Assessment” which provides the foundation for evaluating multiple lines of scientific 

evidence in the context of the understanding of the adverse outcome pathway (or mode of action; (U.S. 

Environmental Protection Agency, 2010). The draft framework was reviewed favorably by the SAP in 

2010 (FIFRA Scientific Advisory Panel (SAP), 2010). Mode of action (Boobis et al., 2006; Boobis et al., 

2008; U.S. Environmental Protection Agency, 2005) and adverse outcome pathway (Ankley et al., 2010)

provide important concepts in this integrative analysis. Both a mode of action and an adverse outcome 

pathway are based on the premise that an adverse effect caused by exposure to a compound can be 

described by a series of causally linked biological key events that result in an adverse human health or 

ecological outcome. One of the key components of the Agency’s draft framework is the use the MOA 

framework /AOP concept as a tool for organizing and integrating information from different sources to 

inform the causal nature of links observed in both experimental and observational studies. Specifically, 

the modified Bradford Hill Criteria are used to evaluate the experimental support that establishes key 

events within a mode of action or an adverse outcome pathway, and explicitly considers such concepts 

as strength, consistency, dose response, temporal concordance and biological plausibility in a weight of 

evidence analysis.

The draft framework is oriented around the source to outcome pathway (Figure 1) as discussed in the 

National Research Council’s (NRC) report on toxicity testing in the 21st Century (National Research 

Council, 2007). Since the 2008 SAP on chlorpyrifos, the Agency has reviewed new experimental 
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toxicology studies evaluating adverse effects in laboratory animals and performed further analyses on 

the existing and new epidemiology results in mothers and children along with relevant biomonitoring 

data. In addition, the Agency has reviewed in vivo and in vitro studies evaluating mechanistic aspects of 

chlorpyrifos evaluating proposed adverse outcome pathways related to the effects on the developing 

brain. The Agency is in the process of developing the weight of evidence analysis integrating the 

epidemiology studies with the experimental toxicology studies for chlorpyrifos; the 2012 SAP review is a 

key milestone in the development of the weight of evidence analysis. This 2012 draft issue paper 

includes review of scientific information in the various areas provided in Figure 1: chlorpyrifos exposure, 

pharmacokinetics, toxicity pathways, and effects on individuals and/or populations. The state of the 

science on each of these areas is summarized in this document.

Figure 1 Source to effects pathway (adapted from NRC, 2007).

This draft science issue paper is organized in a manner consistent with the source-to-outcome paradigm 

(Figure 1). As such, the document begins with a summary of the exposure pathways which exist for 

chlorpyrifos to provide the Panel and the public with scientific context on the routes and duration of 

exposure for chlorpyrifos in addition to the potentially sensitive lifestages. These exposure pathways, 

potentially sensitive lifestages and relevant toxicity pathways and health outcomes are organized in a 

conceptual framework which provides the foundation for the Agency’s analysis (Chapter 2). The AChE 

inhibiting mode of action for chlorpyrifos (Section 3.1) along with several biologically plausible 

hypotheses on modes of action leading to effects on brain development are discussed (Section 3.2.1). 

The next sections of the document consider the adverse effects observed in laboratory animals (Section 

3.2.2) epidemiology studies with mothers and children (Chapter 4). Chapter 5 discusses biomonitoring 

data available for chlorpyrifos along with the state of the science with respect to physiologically based 

pharmacokinetic modeling (PBPK) and possible approaches for interpreting biomonitoring data. The 

Agency is soliciting comment from the SAP on scientific issues related to the interpretation of 

epidemiology, biomonitoring data, and experimental toxicology studies as the Agency prepares to 

integrate information and data from these scientific areas in a weight of the evidence analysis 

considering the extent to which chlorpyrifos may cause long term adverse effects from early life 

exposure (i.e., gestational, postnatal) and that such effects may (or may not) occur below doses 

established from ChE inhibition including in postnatal animals and used for regulatory purposes. 
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2.0 Source-to-Outcome Considerations

2.1 Exposure Pathways for Chlorpyrifos

Chlorpyrifos is an insecticide, acaricide and miticide that controls many insect pests. Chlorpyrifos was 

first registered in 1965 for use in a variety of food and feed crops. As chlorpyrifos use expanded in 

agriculture, use also grew in the residential marketplace. For several years thereafter, it maintained a 

large percentage of the residential marketshare. Residential product labels at that time allowed for use 

indoors as a pest control product. Examples of how it was used included: indoor space and surface 

sprays; as a lawn treatment; as a termiticide; for the control of fleas and ticks on pets; and for 

treatments of residential turf and gardens. The risks from these broad uses of chlorpyrifos were 

evaluated by the Agency in a 2000 risk assessment (Smegal and Leighton, 1999).1

 The outcome of this 

assessment was that most residential uses of chlorpyrifos were phased-out (sales in those markets 

ended by December 2001).2

 Currently, chlorpyrifos is still widely used in agriculture and it still has a 

variety of remaining non-agricultural uses. Agricultural uses can occur on a variety of food and feed 

crops including grapes, many vegetable and fruit row crops (e.g., cauliflower, strawberry), tree crops 

(e.g., citrus and many nut varieties), and field crops (e.g., sorghum, alfalfa). Non-agricultural uses 

include in industrial settings (e.g., industrial plants, railcars, warehouses), on ornamental plants and in

their production, on some types of turf (e.g., sod and golf course), and on wood products (e.g., logs, 

pallets, utility poles). Some non-agricultural products are highly specialized for use in niche markets 

including cattle ear tags, as a mosquito adulticide, and in residential ant bait stations that are sold in 

child resistant packages. 

Chlorpyrifos is being reevaluated under the Registration Review process,3 and a preliminary risk 

assessment has recently been completed that reflects the current use patterns (U.S. Environmental 

Protection Agency, 2011). Risk assessments for pesticides address all manner of exposure pathways as 

well as all pertinent routes and durations of exposure. Metabolites and/or degradate compounds found 

to be of significance are also considered as appropriate. The oxygen analog of chlorpyrifos, referred to 

as its oxon, is known to form in vivo in humans after exposure and in the environment after application 

which can lead to exposures. The oxon of chlorpyrifos is also known to be more toxic than chlorpyrifos, 

per se. The pertinent sources of the oxon and the levels at which direct human exposures can occur are 

described in detail in the Agency’s 2011 risk assessment.

The Food Quality Protection Act (FQPA) of 1996 requires that the Agency complete risk assessments that 

consider all possible exposure routes and pathways and to aggregate the resulting exposures as 

appropriate. This ensures that all sources of exposure for a particular pesticide are reasonably 

considered in regulatory decision making. In this approach, exposures through the diet, from drinking 

water, and residential sources are calculated. Risk assessment for chlorpyrifos considers exposures 

through the diet, from drinking water, and from all other sources that can lead to exposures in the 

general public (referred to as residential exposure). Residential exposures occur because people buy 

products to treat a pest inside their homes or outside (e.g., on lawns). They can also occur because 

people live in or near treated environments where they can be exposed because residues move away 

 1 http://www.epa.gov/pesticides/chem_search/cleared_reviews/csr_PC-059101_5-Oct-99_426.pdf 2 http://www.epa.gov/oppsrrd1/REDs/factsheets/chlorpyrifos_fs.htm

3 http://www.epa.gov/oppsrrd1/registration_review/chlorpyrifos/index.htm 
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from their intended application area (e.g., through spray drift during application or by volatization from 

treated areas). They can also occur because they frequent areas (e.g., golf course). Dietary and drinking 

water exposures are calculated using information which couples how much of a commodity (or drinking 

water) is consumed based on survey data with possible residues in the food or water which is 

consumed. Consumption is defined using a national food intake survey. Commodity residue values are 

based on monitoring data but can be refined using many techniques (e.g., at what point is commodity 

sampled in distribution chain, how much of a given crop is treated with a particular pesticide). 

Residential exposures are calculated using a scenario-driven approach which reflects anticipated 

behaviors for different aged children and adults in various residential settings. Exposures are calculated 

using methodologies that couple what is known about the use of a pesticide and available information 

on its residues in the environment with the scenarios that correctly match how the product can be used. 

Risk assessments also consider occupational exposures individuals can receive through their jobs. For 

chlorpyrifos, as is typical for most agricultural use pesticides, such occupational exposures can occur for 

those involved in the application of a pesticide and those who may be exposed because they are work in 

areas that have been previously treated (e.g., while harvesting fruit or vegetables). Occupational 

exposures are evaluated based on the requirements of the FIFRA. In a manner analogous to residential 

exposures described above, occupational exposures are also evaluated using a scenario based process. 

Scenarios are defined by the types of crops (or non-agricultural areas) and pests that a pesticide is used 

on coupled with any restrictions regarding its use (e.g., certain requirements for protective clothing). 

For those involved in application, a series of standardized exposure rates (i.e., referred to as unit 

exposure values4

) and standard production throughput values (e.g., acres treated per day are commonly 

used for agriculture) are coupled with specific information about each crop where a pesticide can be 

used (e.g., application rates and levels of personal protection required). Exposures for those who work 

in previously treated areas are evaluated using a similar approach. Exposure rates associated with the 

type of crop and activity individual applicators may perform (i.e., known as transfer coefficients5

) are 

coupled with specific information related to how persistent residues are for a pesticide in the 

environment (i.e., typically referred to as dislodgeable foliar residues) to calculate risks. The results of 

these types of assessments assist in defining risk management approaches for occupational tasks. Such 

approaches might entail requiring additional protective clothing or equipment or extending periods 

after application to allow residues to dissipate before allowing certain tasks because exposures are too 

great.

The exposure and risk assessment methods used by the Agency have been extensively vetted using 

mandated public peer review processes. They are also based on guidance used across the Agency and 

throughout the Federal government.6,7

 Additionally, testing guidelines have been established for 

developing the data required to evaluate pesticides for regulatory purposes which have also undergone 

extensive review and input.8

 Science policy papers have also been developed describing how input 

values should be developed and results should be interpreted.9

 Specifically, dietary and drinking water 

exposure and risk assessment methods have been peer reviewed by the SAP on several occasions.10,11

 4

 

http://epa.gov/pesticides/science/handler-exposure-data.html 5 http://www.epa.gov/pesticides/science/post-app-exposure-data.html

6 http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=15263

7 http://www.epa.gov/risk_assessment/history.htm

8 http://www.epa.gov/pesticides/science/guidelines.htm

9 http://www.epa.gov/oppfead1/trac/science/

10 http://www.epa.gov/scipoly/sap/tools/atozindex/dietexp.htm
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Similarly, the methods used to evaluate both residential and occupational exposures have also been 

extensively peer reviewed.12

For detailed information documenting the kinds of exposures identified for chlorpyrifos and the 

calculated exposure values, please refer to Appendix 1. Residential risk estimates calculated prior to the 

removal of those uses from the label in 2001 are presented. These are included because some cohorts 

of interest in this analysis were initiated prior to such uses being removed from the label (e.g., Columbia 

University cohort of lower income women in New York City was initiated in 1997). The remaining values 

included in Appendix 1 are based the 2011 risk assessment. The values in Appendix 1 represent a subset 

of exposure scenarios considered in the 2011 risk assessment; these have been selected as they are a 

representative cross section of current uses of chlorpyrifos. The Agency is not soliciting comment from 

the Panel on the methods or data used to estimate the values in Appendix 1; these are provided for 

informational purposes only in order to provide context for the discussions below regarding exposure. 

[Please refer to the 2011 risk assessment for more information concerning the exposure potential of 

chlorpyrifos associated with current use practices].

2.2 Conceptual Framework

As an aid in organizing the available information and to identify the complex scientific issues being 

considered, the Agency developed a conceptual framework (Figure 2). This conceptual framework is 

consistent with the source-to-outcome pathway (Figure 1) and provides the foundation for the analysis 

being undertaken by the Agency to delineate outcomes derived from AChE inhibition in different 

lifestages and from alternative modes/mechanisms. This conceptual framework provides key 

information on which lifestages are thought be the most susceptible to chlorpyrifos, the exposure 

pathway by which these individuals are exposed, possible adverse outcome pathways and related 

critical duration(s) of exposure leading from exposure to adverse health outcome.

 11 http://www.epa.gov/scipoly/sap/tools/atozindex/drinkrisk.htm 12 http://www.epa.gov/scipoly/sap/tools/atozindex/workerexposure.htm, 

http://www.epa.gov/scipoly/sap/tools/atozindex/residentexp.htm, 

http://www.epa.gov/scipoly/sap/meetings/2009/100609meeting.html, and 

http://www.epa.gov/scipoly/sap/meetings/2008/120208_mtg.htm
Page 10 of 124

Figure 2 Conceptual framework considering chlorpyrifos exposure, lifestages, critical duration, modes of 

action/adverse outcome pathways, and health outcomes of interest. 

2.2.1. Adverse Outcomes

AChE inhibition is the well-established mode of action for OPs and is typically used as the critical effect 

in hazard characterization for members of this class of pesticides. The 2008 SAP concurred with the 

Agency that AChE/ChE data when taking into account sensitive lifestages remains the most robust dose 

response data for use in derivation of points of departure. However, over the last 10-15 years, 

experimental toxicology studies on neurotoxicological as well as epidemiology studies have been 

published which suggest that the developing brain of the fetus and young children may also be affected. 

There are a significant number of literature studies evaluating neurobehavioral outcomes in 

experimental animals (rats, mice). These studies vary in their study design but many involve gestational 

and/or early postnatal dosing with behavioral evaluation in adulthood. Epidemiological studies from 

three cohorts of mothers and children, funded in part by EPA and National Institutes of Environmental 

Health, have reported birth and neurodevelopmental outcomes associated with prenatal exposures to 

OPs. Two of these (Mt. Sinai and Berkeley) have focused primarily upon non-specific urinary biomarkers 

of OP exposure, the dialkyl phosphate metabolites (i.e., DAPs), and thus present some uncertainty as to 

the extent these results relate to chlorpyrifos per se. The Columbia University investigators are studying 

a cohort of mothers and children in inner-city New York and have published multiple papers on 

associations among level of prenatal chlorpyrifos exposure and birth and neurodevelopmental 

outcomes at multiple time points from birth through age seven years. With respect to 

neurodevelopmental outcomes, multiple investigators have reported hypothesized modes/mechanisms 
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of action. In the conceptual framework, both adverse health outcomes of interest for chlorpyrifos are 

being tracked—AChE inhibition and potential for neurodevelopmental effects.

2.2.2. Potentially Sensitive Lifestages

The Agency believes that pregnant women and their fetuses, newborns, and young children represent 

potentially the most sensitive lifestages for exposure to chlorpyrifos. This conclusion is based on 

multiple lines of evidence.

• Pregnant Women: 

Chlorpyrifos RBC AChE data from pregnant rats provides the most sensitive data for use in 

deriving a point of departure from repeated dosing studies. Although it is unclear whether 

metabolic changes during pregnancy are sufficient to affect metabolism at environmental 

exposure, some have suggested that there may be reduced ability to detoxify chlorpyrifos 

and/or the oxon may affect sensitivity during pregnancy. 

Metabolic activities can be altered during pregnancy (Anderson, 2006; Anger & Piquette-Miller, 

2008; Bologa, Tang, Klein, Tesoro, & Koren, 1991; Carpintero, Sanchez-Martin, CabezasDelamare, & Cabezas, 1996; Czekaj, Wiaderkiewicz, Florek, & Wiaderkiewicz, 2000, 2005; 

Dickmann et al., 2008; Ejiri, Katayama, Kiyosawa, Baba, & Doi, 2005; Ferre et al., 2006; Hines, 

2007; Homma et al., 2000; R. M. Howard & Sugden, 1993; Tsutsumi et al., 2001). For example, 

Chanda et al. (2002) showed that pregnant female rats had lower plasma, brain, and liver 

carboxylesterase activity compared to non pregnant females. Regarding A-esterase activity, 

Ferre et al. (2006) showed that the paraoxonase (i.e., A-esterase in serum decreased from a 

nonpregnant background of 146 U/L to 111 U/L in late gestation, indicating 76% of normal 

activity in late gestation pregnant women. Carpintero et al. (1996), however, found that phenyl 

acetate metabolism increased from 23.6 to 33.5 μkat/g in the third trimester, but it must be 

kept in mind that phenyl acetate metabolism is not necessarily a specific measure of A-esterase 

activity. Data in mice support the findings of Ferre et al. (2006) in humans suggesting a 

reduction in A-esterase activity during pregnancy. In mice, Weitman et al. (1983) found that 

PON1 activity after exposure to parathion was 50 nmol/min/ml in non-pregnant females, but it 

decreased as low as 14 nmol/min/ml during gestation (Weitman, et al., 1983). With regard to 

plasma butylcholinesterase (BuChE) activity, Howard et al. (1978) have shown that in six healthy 

pregnant women levels of plasma BuChE activity dropped by approximately 30% during the first 

trimester, but returned to close to pre-pregnancy levels in the third trimester. Similarly, other 

investigators have also reported decreases in plasma BuChE activity in pregnant women (de 

Peyster, Willis, & Liebhaber, 1994; Venkataraman, Iyer, Narayanan, & Joseph, 1990; Whittaker, 

Crawford, & Lewis, 1988). Evans et al. (1988) showed that serum ChE levels in 39 of 44 pregnant 

women dropped after conception; in 20 of those women, the decline in ChE activity continued 

throughout pregnancy. 

Pregnancy is a remarkably dynamic biological process in which rapid changes occurring in both 

the developing system (embryo/fetus) and the mother can significantly impact the 

pharmacokinetics of chemicals. In addition to the potential for decreased clearance of chemicals 

due to immature metabolic systems in the developing embryo/fetus (mentioned above), other 

pregnancy-related changes can also have a pronounced effect on pharmacokinetics. For a 

typical human pregnancy, total body weight gain is in the order of 10-30% while cardiac output 

can increase as much as 50% (Corley, Mast, Carney, Rogers, & Daston, 2003; Young et al., 1997). 
Page 12 of 124

A significant fraction of the body weight gain is due to increases in total body water and fat that 

can lead to a greater volume of distribution by simple dilution. For instance, increases in total 

body water dilute plasma proteins which can increase the free fraction of highly plasma proteinbound chemicals (e.g., chlorpyrifos) available for distribution. Increases in fat, on the other 

hand, can provide a larger storage volume for lipophilic chemicals (Corley, et al., 2003). 

Additionally, significant changes take place in the maternal circulatory system to accommodate 

the development of placental blood flow which undergoes considerable increases throughout 

pregnancy along with embryo/fetus which increases in volume over a billion-fold from 

conception to birth (Young, et al., 1997).

• Fetuses, Infants, Toddlers & Young Children: 

With respect to AChE inhibition, there are multiple studies on chlorpyrifos (Appendix 2) and 

other OPs (U.S. Environmental Protection Agency, 2006) which show that the pregnant dam 

exhibits similar or more AChE inhibition than the fetus at a given dose to the dam. As such, for 

AChE inhibition, by protecting against AChE inhibition in the pregnant female is expected to be 

protective for AChE inhibition in the fetus. With respect to neurodevelopmental outcomes, as

discussed in Section 4.0 below, numerous epidemiological investigations have observed a link 

between prenatal, and notably not post-natal, exposure to chlorpyrifos (measured as parent, 

3,5,6-trichloro-2-pyridinol [TCPy], or DAPs) and adverse effects on fetal growth and 

neurodevelopment through age seven years. Within this database, notable consistencies are 

observed and areas of inconsistencies may be attributed to differences in study methods (i.e., 

measurement of surrogate for prenatal chlorpyrifos exposure), among other sources of 

variablity. Overall, these studies are strong, well-conducted studies in which likely sources of 

error or bias would more likely tend to underestimate, rather than overestimate an effect size 

(e.g., relative risk measure). However, additional analyses are needed to determine with 

certainty whether chlorpyrifos is the etiological causative agent in the adverse 

neurodevelopmental outcomes observed. In addition, there a growing body of literature with 

laboratory animals (rats and mice) indicating that gestational and/or early postnatal exposure to 

chlorpyrifos may cause persistent effects into adulthood. The results of both in vivo and in vitro

studies on chlorpyrifos have led some research groups to propose that changes in brain 

connectivity and/or neurochemistry may underlie these changes into adulthood. 

Potential sensitivity, particularly from acute, single dose exposures, is largely derived from 

immature metabolic systems that have less capacity to detoxify the parent OP and its 

toxicologically active metabolite chlorpyrifos oxon. Differential inhibition of the AChE enzyme 

itself does not appear to account for the observed age-related sensitivity found in young animals 

as suggested by in vitro studies (Atterberry, Burnett, & Chambers, 1997; Benke & Murphy, 1975; 

Chanda, Harp, Liu, & Pope, 1995; Mortensen, Chanda, Hooper, & Padilla, 1996). Rat fetuses and 

juveniles and newborn humans, however, have lower capacity to detoxify than adults. This 

decreased capacity to detoxify has been associated with increased sensitivity in rats. 

Specifically, in rats, A-esterase activity is virtually nonexistent in the fetus (Lassiter et al., 1998)

and increases from birth to reach adult levels around PND21 (W. F. Li, Matthews, Disteche, 

Costa, & Furlong, 1997; Mortensen, et al., 1996). Mortenson et al. (1996) showed that in the 

plasma level of CPOase13

 13 CPOase is A-esterase (PON1) activity specific to chlorpyrifos oxon

was 1/11 that of adult animals. The animal data regarding the role of 

carboxylesterase in mediating OP toxicity are also quite extensive (e.g., Clement, 1984; Fonnum, 

Sterri, Aas, & Johnsen, 1985; Maxwell, 1992a, 1992b). Fetal rats possess very little 
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carboxylesterase activity (Lassiter, et al., 1998) with increasing activity as the postnatal rat 

matures, reaching adult values after puberty (50 days-of-age; Karanth & Pope, 2000; Morgan, 

Yan, Greenway, & Parkinson, 1994; Moser, Chanda, Mortensen, & Padilla, 1998). The temporal 

pattern of A-esterase and carboxylesterase activities correlates reasonably well with studies on 

OP sensitivity. Several studies have shown an increased sensitivity of newborn rats to OP 

compounds, and specifically chlorpyrifos, which are detoxified via the A-esterase and/or 

carboxylesterase pathways (Benke & Murphy, 1975; Chambers & Carr, 1993; Chanda, et al., 

2002; Gagne & Brodeur, 1972; Karanth & Pope, 2000; Morgan, et al., 1994; Mortensen, et al., 

1996; Moser, et al., 1998; C. N. Pope, Chakraborti, Chapman, Farrar, & Arthun, 1991). Although 

maturation of the cytochrome P450s may also play a role in age-related differences in the ability 

of young and adults to detoxify or activate chlorpyrifos to the oxon, the degree to which this 

maturation impacts juvenile sensitivity to chlorpyrifos has not been extensively studied.

There are, however, relatively little data in human tissues which could evaluate age-related 

maturation of expression of these detoxification esterases. For example, Pope et al. (2005)

evaluated maturational expression of liver carboxylesterases in human liver tissues from infants 

(2–24 months) and adults (20–36 years). The authors report relatively small (and not 

statistically significant) differences in activities between children ages 2–24 months and adults 

(20–36 years). More recently, others have measured expression and activity of human 

carboxylesterases 1 and 2 in human fetuses, newborns, and children, and while there are 

considerable inter-individual differences, the lowest activities are recorded at the youngest ages 

(Yang et al., 2009; Zhu, Appel, Jiang, & Markowitz, 2009).

There are also studies in the literature that have assessed A-esterase (also known as PON1) 

activity in children. Based on these studies, it appears that serum A-esterase levels are very low 

in human infants compared to adults (Augustinsson & Barr, 1963; Augustinsson & Brody, 1962; 

J. Chen, Kumar, Chan, Berkowitz, & Wetmur, 2003; Cole et al., 2003; Holland et al., 2006; 

Mueller et al., 1983). After birth, there is a steady increase of this activity (Augustinsson & Barr, 

1963) with approximately a 2x increase in A-esterase activity between birth and adulthood 

(Ecobichon & Stephens, 1973). In a related study of the age-dependence of total serum 

arylesterase (ARase) activity (of which a large component is A-esterase activity), adult levels 

were achieved by two years-of-age (Burlina, Michielin, & Galzigna, 1977). Recent studies have 

shown age-dependant increases in PON1 activity from birth to aged 7, and that by age 7 the 

activity were similar to or slightly less than mothers in the study (Huen et al., 2011; Huen, 

Harley, Bradman, Eskenazi, & Holland, 2010; Huen et al., 2009). In addition, other studies (J. 

Chen, et al., 2003; Holland, et al., 2006) have provided A-esterase and/or carboxylesterase

activities in newborns and their mothers which show that newborns have lower activities than 

their mothers. Thus, the lower detoxification potential of chlorpyrifos and other OPs in young 

rats is also suggested in humans, leading to the potential of greater sensitivity of infants and 

children due to kinetic factors.

There are critical windows of vulnerability (Rice & Barone, 2000; Rodier, 2004) with regard to 

toxicant effects on brain development. This vulnerable period in humans spans early pregnancy 

to adolescence (Rice & Barone, 2000). In fact, recent evidence shows that synapse formation 

peaks quite late in human brain development at 4-8 years of age (Glantz, Gilmore, Hamer, 

Lieberman, & Jarskog, 2007). Within these vulnerable periods there are key 

neurodevelopmental processes (e.g. cell division, migration, differentiation, synaptogenesis, and 

myleination) and each of these is region and stage specific. Consequently, the time of toxicant 
Page 14 of 124

exposure will be a major determinate in the spectrum of neurotoxic effects. In addition, young 

children may be more susceptible with respect to exposure, as they age because they engage in 

more active behaviors for more sustained periods that can lead to higher overall exposures. 

These types of considerations are accounted for in the processes used by the Agency to evaluate 

the residential exposures of pesticides.

2.2.3. Critical Duration(s) of Exposure

It is important to appropriately match the exposure potential with the toxicology profile in risk 

assessment. As such, it is important to consider a variety of toxicokinetic and toxicodynamic factors 

such as the half life of the chemical in the body and the duration of the toxic effect (i.e., the half life of 

recovery). AChE inhibition can occur from a single exposure or from repeated exposures. Chlorpyrifos 

exposure can occur from a single exposure (e.g., eating a meal) or from repeated days of exposure (e.g., 

worker). With respect to OPs, repeated exposures generally cause more inhibition at a given 

administered dose compared to acute studies. Moreover, AChE inhibition in repeated dosing guideline 

toxicology studies with OPs show a consistent pattern of steady state inhibition at or around 2-3 weeks 

of exposure in adult laboratory animals (U.S. Environmental Protection Agency, 2002). Specifically, with 

repeated dosing at a given dose, the amount of inhibition comes at equilibrium with production of new 

enzyme. As such, AChE studies of 2-3 weeks generally show the same degree of inhibition with those of 

longer duration (i.e., up to 2 years of exposure). Thus, for the health outcomes relevant to the AChE 

inhibition mode of action in adults, the critical durations range from a single day up to 2-3 weeks (i.e., 

the time to reach steady state). In newborns and young children, there are uncertainties associated as 

the concept of steady state may not hold as the metabolic capacity of children is changing over time 

(i.e., increasing towards adult levels with age). 

Very little is known about the duration of chlorpyrifos exposure needed to precipitate adverse effects in 

the developing brain. Because of the dynamic processes in the developing brain (i.e., vulnerable 

windows) it is difficult to determine if the effect or differences in effects is due to duration of exposure 

or if different vulnerable windows were affected. As such, it is impossible at this time to rule out even a 

single day of exposure having a potential effect in humans. 
Page 15 of 124

3.0 Adverse Outcome Pathways: AChE Inhibition & Plausible Pathways 

Leading to Neurodevelopmental Outcomes

Mode of action and adverse outcome pathways provide important concepts in the draft “Framework for 

Incorporating Human Epidemiologic & Incident Data in Health Risk Assessment” (U.S. Environmental 

Protection Agency, 2010). As mentioned above, both a mode of action and an adverse outcome 

pathway (AOP) include a set of measureable key events that describe the biological processes leading to 

an apical effect. However, an adverse outcome pathway explicitly considers the steps linking a 

molecular initiating event and an adverse outcome at a biological level of organization that is relevant 

for risk assessment. Figure 3 is a graphical presentation of a generic adverse outcome pathway (Ankley, 

et al., 2010). This figure is an extension of the source-to-outcome pathway (Figure 1) and provides 

additional detail on the types of scientific information from various levels of biological organization used 

in establishing an adverse outcome pathway. As shown in the conceptual framework (Figure 2), two 

different adverse outcomes in laboratory animals are being evaluated: AChE/ChE inhibition and 

neurodevelopmental outcomes. Each of these is discussed below.

Figure 3 Generic Adverse Outcome Pathway (Adapted from Ankley et al., 2010) 

3.1 Adverse Outcome Pathway: AChE Inhibition

3.1.1 Initiating Event & Health Outcomes

AChE inhibition is the well-established mode of action for OPs and is typically used as the critical effect 

in hazard characterization for members of this class of pesticides. Specifically, the initiating event for OP 

mediated neurotoxicity is interaction at the serine residue at the active site of the AChE. This 

interaction leads to accumulation of acetylcholine and ultimately to clinical signs of neurotoxicity at high 

doses. In experimental toxicity studies in laboratory animals, high acute doses produce signs of 
Page 16 of 124

autonomic stimulation, ataxia, fasciculations, tremors, respiratory difficulties, and convulsions. The 

Agency’s recent review of human incident reporting databases (Recore & Oo, 2011) indicates that 

humans exposed to high levels of chlorpyrifos may complain of gastrointestinal and neurological effects 

consistent with acute AChE poisoning. Specifically, individuals have reported: nausea, vomiting, 

diarrhea, tremors, headaches, dizziness, numbness and tingling sensations, muscle spasm, shortness of 

breath, and seizures (Recore & Oo, 2011).

Although inhibition of AChE as the initiating event in this adverse outcome pathway has long been 

established, the quantitative dose response linkage between AChE inhibition and toxicity has been 

shown to be more variable. In other words, the amount of AChE inhibition required to elicit such 

responses from OP exposure is not consistent but instead varies among different OPs, different study 

designs, and different neurotoxic outcomes. Specific to chlorpyrifos, there are limited human data 

which inform the quantitative dose-response linkage between AChE inhibition and apical neurotoxic 

effects. There is one human deliberate dosing study in human subjects (Nolan, Rick, Feshour, & 

Saunders, 1982)14 which provides plasma ChE and RBC AChE data from a single oral dose (0.5 mg/kg). 

AChE/ChE activity (plasma and RBC) was monitored at 2, 6, 12, 24 hours, and 2, 3, 4, 8, 14, 22, 27, and 

30 days post dose. Despite plasma ChE inhibition of 83-89% at the peak time of effect, no clinical signs 

of toxicity or inhibition of RBC ChE activity were observed. In addition to this deliberate dosing study, 

several observational worker studies have monitored urinary TCPy and ChE activity. The Agency has 

previously reviewed these monitoring studies in its 2008 SAP meeting on chlorpyrifos and concluded 

that urinary TCPy was not predictive of AChE inhibition. Further description of the Agency evaluation of 

the relationship between TCPy and AChE inhibition is described in Section 5.4 of this report. 

3.1.2 Dose Response Analysis for AChE Inhibition

Points of departures can be no-observed-adverse-effect-levels (NOAELs), low-observed-adverse-effectlevels (LOAELs), or derived from benchmark (BMD) modeling. Numerous scientific peer review panels 

over the last decade have supported the Agency’s application of the BMD approach as an improvement 

over the historically applied approach of using NOAELs or LOAELs and as a scientifically supportable 

method for deriving points of departure in human health risk assessment. The NOAEL/LOAEL approach 

does not account for the variability and uncertainty in the experimental results, which are due to 

characteristics of the study design, such as dose selection, dose spacing, and sample size (U.S. 

Environmental Protection Agency, 2000). With the BMD approach, all the dose response data are used 

to derive a point of departure. 

For the chlorpyrifos risk assessment, the Agency has used a decreasing exponential dose-response 

model similar to that used for the OP and N-methyl carbamate cumulative risk assessments and multiple 

risk assessments for individual AChE inhibiting pesticides (U.S. Environmental Protection Agency, 2006, 

2007). The use of this empirical dose-response procedure has been previously reviewed and supported 

by the FIFRA SAP on several occasions (FIFRA Scientific Advisory Panel (SAP), 2001, 2002, 2005a, 2005b, 

2008b). In lieu of a defined response level (i.e., percent inhibition) that quantitatively links AChE 

inhibition to neurotoxic effects, the Agency typically uses a 10% decrease in brain and peripheral ChE

inhibition as the response level in deriving points of departure. This 10% response level is called the 

benchmark response level (BMR). This level has been shown to be a level which is protective for 

neurotoxic outcomes for OPs and can also be reliably measured in most guideline experimental 

 14 Nolan (1982) has been reviewed by the Human Studies Review Board (HSRB) and found to be ethically 

and scientifically conducted (HSRB, 2009). 
Page 17 of 124

toxicology studies (U.S. Environmental Protection Agency, 2002, 2006). The Agency has calculated 

BMD10s and BMDL10s where the BMD10 is the estimated dose where AChE is inhibited by 10% compared 

to background and the BMDL10 is the lower confidence bound on the BMD10. As a matter of science 

policy, the Agency uses the BMDL, not the BMD, for use as the point of departure since the BMDL 

accounts for variability of the data (U.S. Environmental Protection Agency, 2000). 

As part of the 2008 SAP, the Agency performed a comprehensive review of the literature on available 

AChE data across multiple lifestages, durations, and routes (FIFRA Scientific Advisory Panel (SAP), 

2008a). This literature review was considered by the SAP at that time and provided part of the scientific 

rationale for the Panel’s recommendation that the Agency continue to use AChE data in the most 

sensitive lifestages for dose-response analysis and deriving points of departure. This extensive 

literature is not summarized here but can be found in Draft Appendix B of the 2008 draft science issue 

paper presented to the SAP (FIFRA Scientific Advisory Panel (SAP), 2008a). Instead, the following text 

focuses on the most robust dose-response studies used in the deriving points of departure for the 2011 

risk assessment. 

For endpoint selection, the Agency considered the quality of all the available studies, both previously 

available at the time of the 2008 SAP and newest ones. Route-specific dermal and inhalation studies 

are available for chlorpyrifos and are being used to assess dermal and inhalation exposure, respectively. 

Information on these studies can be found in the 2011 risk assessment (U.S. Environmental Protection 

Agency, 2011). There are also high quality dose-response data in adults and juvenile rats for chlorpyrifos 

oxon. Dose response analysis for the oxon data has been conducted and can also be found in the 2011 

risk assessment. 

For the preliminary risk assessment, the Agency has conducted BMD modeling on selected oral AChE 

studies. These studies were selected based on the availability of at least two treatment groups in 

addition to a control group and those that showed good dose response. The Agency has also focused on 

those studies where AChE was measured at or near the time of peak effect, typically within one to 

several hours after dosing. AChE data measured 24 hours or longer after dosing will underestimate the 

amount of AChE inhibition and are thus not appropriate for deriving points of departure. In addition, 

the studies used in the dose response assessment were selected as they represented a variety of ages, 

lifestages, and durations. The Agency has focused on those studies representing single and repeating 

dosing in post-natal exposure to rat ages (PND 10 and older); PND10 and older is considered to be 

approximately concordant with human postnatal exposure (Benjamins & McKhann, 1981; Dobbing & 

Smart, 1974). AChE data from adult laboratory animals (both pregnant and non-pregnant) from single 

dose and repeating dosing studies has also been collected. The focus has been on AChE data from rats 

as these are generally the most robust and sensitive studies. Data in fetuses have not been evaluating 

using BMD techniques as these studies tend to be less sensitive than those in post-natal juveniles and 

adults. The Agency has conducted BMD analysis on available data for blood (RBC), brain (whole or 

sections as appropriate), and peripheral (i.e., heart) tissues. 

The Agency typically performs risk assessments for oral exposure to food and water, in addition to nondietary ingestion to young children. For assessment of food, water, and aggregate dietary risk, the 

Agency typically conducts single day (i.e., acute) and chronic assessments. For assessment of nondietary ingestion to young children, the Agency conducts short- and intermediate-term assessments 

which involve durations of 1-30 days and 1 month to 6 months, respectively. 
Page 18 of 124

There are several studies available which inform the acute oral point of departure. For the assessment 

of infants and children, the most robust studies are 1) the new comparative cholinesterase15 (CCA) study 

(MRID 48139301) which provides brain and RBC data in PND11 pups and 2) (Moser, Simmons, & 

Gennings, 2006) which provides data from brain and whole blood ChE data in PND17 pups. In both of 

these studies, the rat pups were dosed directly and ChE was measured at the time of peak effect. In 

addition, two studies (Timchalk, Poet, & Kousba, 2006; Q. Zheng, Olivier, Won, & Pope, 2000) provide 

supporting data. These two studies show AChE data in post-natal rats which are consistent with the 

comparative cholinesterase study and Moser et al (2006). However, Timchalk et al. (2006) and Zheng et 

al. (2000) are not robust datasets for dose-response modeling (U.S. Environmental Protection Agency, 

2011). 

For the assessment of adults, single dosing AChE data from the comparative cholinesterase study (MRID 

48139301) provide the most robust data for BMD analysis16. Tables 1 and 2 provide a summary of the 

BMD results for the single dosing studies in post-natal rat pups and adult rats, respectively. Generally 

for the both the pups and the adults, the BMD and BMDL estimates for the blood measures are lower 

than those for the brain. In addition, the results from the pups are approximately 2-fold lower than 

those for adults (i.e., when comparing the same compartment, administration vehicle). The BMDL10 for 

RBC AChE inhibition of 0.36 mg/kg from the PND11 pups exposed to chlorpyrifos in milk provides the 

most sensitive data for deriving a point of departure. These results are supported by the Moser et al 

(2006)BMDL10 of 0.43 mg/kg from whole blood ChE, from the RBC AChE measures from the gavage 

administered PND11 pups in the comparative cholinesterase study (BMDL10s of 0.65 and 0.75 mg/kg), 

and the adult female results from exposure in the diet for 12 hours (0.60 mg/kg). 

 15 Comparative cholinesterase (CCA) studies are specially designed studies in juvenile and adult rats 

which provide direct comparative ChE/AChE data for evaluating lifestage sensitivity. For OPs, data are 

typically collected from acute dosing (typically PND11, young adult) and from 11-days repeated dosing 

(PND11-21 and young adults).

16 Only adult female rats were tested in MRID 48139301. 
Page 19 of 124

Table 1. Results of BMD Modeling of Rat Pup Brain and Blood ChE Inhibition following a Single Oral Dose of 

Chlorpyrifos

Table 2. Results of BMD Modeling of Adult Rat Brain and RBC AChE Inhibition following a Single Oral, Dose of 

Chlorpyrifos

Reference

Sex & 

Administration 

Method

Endpoint BMD10 (mg/kg) BMDL10 (mg/kg)

CCA 

(MRID 48139301)

Female 

Gavage Brain AChE 4.11 2.26

CCA 

(MRID 48139301)

Female

Gavage RBC AChE 1.9 1.2

CCA 

(MRID 48139301)

Female

12 hr diet Brain AChE 4.47 3.30

CCA 

(MRID 48139301)

Female

12 hr diet RBC AChE 1.03 0.6

Several high quality repeated dosing oral studies are available determination of oral points of departure

for chronic dietary scenarios (Tables 3 and 4). For BMD analysis in the preliminary risk assessment, the 

Agency selected the repeated dosing portions of the comparative cholinesterase study (MRID 48139301) 

in juvenile and non-pregnant adult rats;, the data from pregnant dams in the developmental 

neurotoxicity rat study (Maurissen, Hoberman, Garman, & Hanley, 2000) (MRID 44556901) and the data 

Reference Sex & Age at 

Dosing Endpoint BMD10 (mg/kg) BMDL10 (mg/kg)

Moser et al., 2006 Male PND 17

Gavage

Brain AChE 1.89 1.54

Moser et al., 2006 Male PND 17

Gavage

Whole Blood 

AChE 0.62 0.43

CCA Study 

(MRID 48139301)

Male PND 11 

Gavage

Brain AChE 2.13 1.53

CCA Study 

(MRID 48139301)

Male PND 11

Gavage 

RBC AChE 0.82 0.65

CCA Study 

(MRID 48139301)

Female PND 11 

Gavage

Brain AChE 2.18 1.56

CCA Study 

(MRID 48139301)

Female PND 11 

Gavage

RBC AChE 0.96 0.75

CCA Study 

(MRID 48139301)

Male PND 11 

Milk Brain AChE 4.4 2.4

CCA Study 

(MRID 48139301)

Male PND 11 

Milk RBC AChE 0.47 0.36

CCA Study 

(MRID 48139301)

Female PND 11 

Milk Brain AChE 1.42 0.91

CCA Study 

(MRID 48139301)

Female PND 11 

Milk RBC AChE 0.5 0.36
Page 20 of 124

from pregnant dams in the special cholinesterase study in which dams where administered chlorpyrifos 

from gestation day 6 through lactation day 10 (GD6 – LD10) (Mattsson, Maurissen, Nolan, & Brzak, 

2000); MRID 44648101). Although the Zheng et al (2000) data are not robust data for BMD modeling, 

the repeated dosing portion of this study provides additional supportive data for the findings from the 

comparative cholinesterase study in both pups and adults. 

Table 3. Results of BMD Modeling of Pup Rat Brain and RBC AChE Inhibition following 11-days Repeated Oral, 

Gavage Doses of Chlorpyrifos

Reference

Sex & 

Age/Duration of 

Dosing

Endpoint BMD10 

(mg/kg/day)

BMDL10 

(mg/kg/day)

CCA 

(MRID 48139301) 

Female PND 11-21

Gavage Brain AChE 0.80 0.69

CCA 

(MRID 48139301)

Female PND 11-21

Gavage RBC AChE 0.17 0.15

CCA 

(MRID 48139301)

Male PND 11-21

Gavage Brain AChE 0.63 0.52

CCA 

(MRID 48139301)

Male PND 11-21

Gavage RBC AChE 0.11 0.09

Table 4. Results of BMD Modeling of Adult, Female Rat Brain, RBC and Heart Cholinesterase Inhibition following 

Repeat Oral Doses of Chlorpyrifos

Reference Sex, Age/Duration & 

Administration method Endpoint BMD10 (mg/kg/day)

BMDL10 

(mg/kg/day)

Dow (MRID 44556901; 

Maurissen et al., 2000)

Dams, GD6-20, 

Gavage Brain AChE 0.65 0.54

Dow (MRID 44556901; 

Maurissen et al., 2000)

Dams, GD6-20, 

Gavage RBC AChE 0.06 0.03

Dow (MRID 44648101; 

Mattsson et al., 2000) Dams, GD6-20, Gavage Brain AChE 1.1 0.8

Dow (MRID 44648101; 

Mattsson et al., 2000) Dams, GD6-20, Gavage RBC AChE 0.14 0.08

Dow (MRID 44648101; 

Mattsson et al., 2000) Dams, GD6-20, Gavage Heart ChE 0.85 0.22

Dow (MRID 44648101; 

Mattsson et al., 2000)

Dams, GD6-LD10

Gavage Brain AChE 1.13 0.89

Dow (MRID 44648101; 

Mattsson et al., 2000)

Dams, GD6-LD10

Gavage RBC AChE 0.050 0.044

Dow (MRID 44648101; 

Mattsson et al., 2000)

Dams, GD6-LD10

Gavage Heart ChE 0.21 0.18

CCA (MRID 48139301) Adult, non-pregnant

11 daysGavage

Brain AChE 1.03 0.95

CCA (MRID 48139301) Adult, non-pregnant

11 days Gavage

RBC AChE 0.45 0.35
Page 21 of 124

As shown in Tables 3 and 4, there is no meaningful difference between BMD10s and BMDL10s from pups 

and adults in repeated dosing studies. This is not unexpected given that over the duration of the 

repeated dosing study, the metabolic system in pups is maturing closer to adult levels. Thus, it is not 

unexpected that as the metabolic system matures that the degree of sensitivity decreases. Typically, 

studies submitted for pesticide registration and most studies from the public literature only measure 

brain and/or blood ChEs. It is rare for data from peripheral tissues to be available for consideration. 

Chlorpyrifos is unique in that multiple studies are available which provide such peripheral data. It is 

notable that the BMDs and BMDLs for heart ChE are more sensitive than brain and closer in magnitude 

than those for RBC AChE. This finding supports the use of the blood measures. 

There is good concordance in the estimates across the BMD10 and BMDL10 estimates across the adult 

studies evaluated in this analysis. Notably, brain AChE estimates range from approximately 0.5 to 1 

mg/kg/day and the RBC AChE estimates range from 0.03 to 0.35 mg/kg/day. In addition, for the OP 

cumulative risk assessment (U.S. Environmental Protection Agency, 2006), the Agency performed BMD 

modeling using a sophisticated meta-analysis approach on brain AChE data from adult, non-pregnant 

female and male rats from one subchronic oral toxicity study and two chronic oral toxicity studies (MRID 

nos. 40952801, 40952802, and 42172802). The BMD results reported in the OP cumulative risk 

assessment for chlorpyrifos are BMD10s of 1.48 and 1.50 mg/kg/day and BMDL10s of1.26 and 1.27, 

mg/kg/day respectively in females and males. The values for brain AChE inhibition reported in the 

cumulative assessment from subchronic and chronic dosing are quite similar to those reported in Table 

4 above for both pregnant and non-pregnant females and males in the selected studies and add further 

support to the analysis conducted for the preliminary risk assessment for chlorpyrifos. For purposes of 

deriving a point of departure for risk assessment, the Agency is using the BMDL10 of 0.03 mg/kg/day for 

RBC AChE inhibition from the DNT study (MRID 44556901, also called (Maurissen, et al., 2000).

3.1.3 Summary of the Dose Response Analysis

Acute oral point of departure is BMDL10 = 0.36 mg/kg based on inhibition of RBC AChE in male and 

female rat pups from the CCA study and is supported by RBC BMDs from data on PND17 males (Moser 

et al.,(2006), and adult female in the comparative cholinesterase study exposed in the diet. These 

studies are further supported by Timchalk et al. (2006) and Zheng et al. (2000) studies. The chronic oral 

point of departure is BMDL10 = 0.03 mg/kg/day based on inhibition of RBC AChE in rat dams (GD6-20) 

from the DNT study (MRID 44556901, also called Maurisse et al., 2000) with support from pregnant 

(GD6-LD10) rats (MRID 44648101). 

3.2 Adverse Outcome Pathway: Neurodevelopmental Outcomes

With respect to modes of action/adverse outcome pathways leading to neurodevelopmental effects, at 

the present time, there is no established series of causal key events at a biological level of organization 

relevant to the risk assessment (i.e., adverse neurodevelopmental effects from gestational and/or 

postnatal exposure). The Agency does not believe that any of the current lines of research support a 

coherent set of key events and that much work remains to elucidate the MOAs and AOP of chlorpyrifos 

toxicity. Even though a rigorous demonstration of an adverse outcome pathway for chlorpyrifos has not 

been constructed at this time, there are multiple studies from numerous laboratories that show that 

exposure to a variety of OPs (including chlorpyrifos, other pesticides, nerve gases) during the 

developmental period affects key events in nervous system development (see below). These studies
Page 22 of 124

conducted to date generally report a correlation between OP exposure and a tested effect without 

consideration of temporal concordance (i.e., the sequence of biological events leading to effects) and 

quantitative linkages (i.e., degree of change that results in neurodevelopmental consequences). Some 

of this experimental work specific to chlorpyrifos is summarized below, and may be shown in the future 

actually to be steps in the same AOP and not discreet pathways. However, at this time, the experimental 

studies below do not provide all the information needed to demonstrate linkages among initiating 

event(s), subsequent events at the molecular and cellular level, and the adverse outcome of interest and 

are thus described separately.

3.2.1 Biologically Plausible Toxicity Pathways Leading to Neurodevelopmental 

Outcomes

3.2.1.1 Acetylcholinesterase (AChE) as a morphogen

The classically understood role of AChE is the rapid hydrolysis of acetylcholine at synapses in the brain 

and at neuromuscular junctions, thereby regulating cholinergic neurotransmission. Consistent with this 

role, AChE is predominant at cholinergic synapses at neurons and in muscle, and inhibition of its 

catalytic activity results in the signs and symptoms of cholinergic overstimulation. Several lines of 

evidence, however, suggest that AChE can also serve as a morphogen, influencing the growth of cells 

during neurodevelopment. Alterations in the expression or structure of the AChE protein can disrupt 

various aspects of neuronal differentiation and growth. Because chlorpyrifos can interact with AChE, 

perturbation of this morphogenic activity represents a plausible adverse outcome pathway leading to 

developmental neurotoxicity.

One of the first indications that AChE had roles other than the termination of cholinergic 

neurotransmission was the timing and distribution of its expression in the developing nervous system. In 

certain brain regions containing non-cholinergic neurons and few cholinergic synapses, AChE levels are 

still high. In addition, AChE is highly expressed throughout the brain during periods of active axonal 

outgrowth in the absence of other cholinergic markers and before synaptic connections are made 

(Bigbee, Sharma, Gupta, & Dupree, 1999; Brimijoin & Koenigsberger, 1999). For example, AChE levels 

are high during early cerebellar development while there is very little evidence for the presence of 

cholinergic cells or cholinergic neurotransmission (Appleyard & Jahnsen, 1992; Parvari, 1983). A second 

line of evidence for a morphogenic role arose after the cloning and sequencing of rodent and human 

AChE. Surprising sequence homologies were found with a family of proteins lacking a catalytically active 

esterase site but pocessing a similar extacellular domain. Several of the homologs of AChE are 

extracellular matrix components essential to neuronal adhesion, axon guidance, and synapse formation 

(Brimijoin & Koenigsberger, 1999; Grisaru, Sternfeld, Eldor, Glick, & Soreq, 1999). Thus, the sequence of 

AChE is similar to other proteins which have a morphogenic role during the development of the nervous 

system. Finally, experimental evidence indicates that manipulation of AChE levels or activity can 

influence neuronal growth in vitro. Studies in cultured neuroblastoma cells (Koenigsberger, Chiappa, & 

Brimijoin, 1997) or PC12 cells (Grifman, Galyam, Seidman, & Soreq, 1998) have shown that antisense 

suppression of AChE decreased neurite outgrowth, and that neurite outgrowth could be rescued or 

increased by transfection with sense AChE. In cultures of rat dorsal root ganglion cultures, treatment 

with an antibody against AChE (which did not inhibit the catalytic activity) decreased neurite outgrowth, 

and neurite outgrowth was restored upon removal of the antibody (Bigbee, et al., 1999). Expression in 

frog embyros of a recombinant human AChE lacking the ability to hydrolyze acetylcholine resulted in 

increased neurite outgrowth from cultured spinal neurons (Sternfeld et al., 1998). Thus, the catalytic 

ability of AChE does not seem to be required for its morphogenic properties. 
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Studies using pharmacologic inhibition of AChE activity to alter neurite outgrowth also suggest a 

dissociation of the catalytic and morphogenic activities. Several laboratories using different in vitro

systems have demonstrated that some potent AChE inhibitors including BW284c51 supress neurite 

outgrowth while other, equally potent AChE inhibitors like ecothiophate do not (Bigbee, et al., 1999; 

Koenigsberger, et al., 1997; Layer, Weikert, & Alber, 1993). These differential effects may be related to 

the degree of chemical binding to separate sites on the AChE enzyme. There are at least two sites that 

can bind inhibitors: the catalytic (active) site and the peripheral anionic site (Grisaru, et al., 1999). It has 

been suggested that inhibitors that bind directly to the peripheral anionic site, or that bind to the 

catalytic site in such a way that changes the conformation of the peripheral anionic site, will alter the 

morphogenic function of AChE (Bigbee, et al., 1999; Yang et al., 2008). Thus inhibitors that bind only to 

the catalytic site (e.g. edrophonium, tacrine) did not affect neurite outgrowth, while ligands that bind to 

the peripheral anionic site (propidium, gallamine) or both sites (BW284c51) inhibited neurite outgrowth 

(Koenigsberger, et al., 1997; Munoz, Aldunate, & Inestrosa, 1999). At least one study indicates that 

chlorpyrifos oxon can bind at both sites (Kousba, Sultatos, Poet, & Timchalk, 2004).

Data showing a potential action of chlorpyrifos on the morphogenic function of AChE are derived 

primarily from in vitro studies. Chlorpyrifos (as well as the activated oxon form and the inactive 

metabolite TCPy) has been examined for effects on neurite outgrowth using a number of cell lines and in 

primary neuronal cultures. In some cases concurrent assessments of cholinesterase inhibition or 

cytotoxicity were performed. Li and Cassida (W. Li & Casida, 1998) examined the effect of chlorpyrifos 

oxon on neurite outgrowth in the PC12 cell line. In this study cells were exposed prior to, but not during 

the active phase of neurite growth, which was assessed after 5 days. Chlorpyrifos oxon inhibited neurite 

outgrowth by 50%, but only at a relatively high concentration (200 uM) that was both cytotoxic and 

decreased cholinesterase activity. Sachana et al. (Sachana, Flaskos, & Hargreaves, 2005; Sachana, 

Flaskos, Sidiropoulou, Yavari, & Hargreaves, 2008) examined the effect of 3 uM chlorpyrifos on the 

initiation of neurite growth in mouse N2a neuroblastoma cells. The number of cells exhibiting neurites 

after 4 or 8 hours of exposure (considered as a measure of cell differentiation) was decreased by 

approximately 50% in the absense of cytotoxicity. Cholinesterase activity was not measured. Axelrad et 

al. (2002) exposed mouse NB2a cells to chlorpyrifos for 24 hr and directly measured neurite length. 

Chlorpyrifos inhibited neurite outgrowth by 50% at 25 uM. Again, cholinesterase activity was not 

measured. While the studies cited above suggested that chlorpyrifos and chlorpyrifos oxon could affect 

measures related to neuronal differentiation and neurite growth, they were not specifically designed to 

examine the morphogenic role of AChE. 

Using PC12 cells, Das and Barone (1999) examined the concentration-related effects of chlorpyrifos, 

chlorpyrifos oxon, and TCPy on both neurite outgrowth and cholinesterase inhibition. Exposure to 

chlorpyrifos for 24 hr inhibited neurite outgrowth at a concentration (3 uM) 10-fold below that which 

inhibited cholinesterase activity, while chlorpyrifos oxon inhibited both measures at equivalent 

concentrations (1 nM). TCPy, which is inactive against AChE, inhibited neurite growth at 5 uM. Similar 

studies of chlorpyrifos and its metabolites were performed in a series of experiments in the Lein 

laboratory using primary neuronal cultures derived directly from the mammalian nervous system. Using 

rat sympathetic neurons, Howard et al. (2005) showed that 24 hr exposure to chlorpyrifos and 

chlorpyrifos oxon decreased axonal outgrowth at concentrations (0.001 uM and 0.001 nM, respectively) 

well below the concentrations that inhibited AChE activity (1 uM and 1 nM). In the same study 

chlorpyrifos, chlorpyrifos oxon and TCPy enhanced dendrite outgrowth. A follow-up study from the 

same laboratory used sensory neurons from the dorsal root ganglion, which extend only axons (Yang, et 

al., 2008). Similar results were observed, with both chlorpyrifos and chlorpyrifos oxon decreasing axonal 
Page 24 of 124

outgrowth at concentrations below those that inhibited AChE activity. To establish whether the target of 

chlorpyrifos and chlorpyrifos oxon was AChE or some other molecule, Yang et al. (Yang, et al., 2008)

repeated the experiments in cultures from AChE-null animals. Chlorpyrifos and chlorpyrifos oxon had no 

effect in the AChE-null cultures, suggesting that inhibition of neurite outgrowth required AChE. These 

later three studies provide the most convincing evidence for selective inhibition of the morphogenic 

activity of AChE. It is not yet clear whether the effects are a direct action of chlorpyrifos on AChE, or the 

result of in vitro conversion to the active oxon form. It should be mentioned that studies from the 

Slotkin laboratory suggest an effect of chlorpyrifos on neurite outgrowth in vitro (Song, Violin, Seidler, & 

Slotkin, 1998). In those experiments, however, neurite outgrowth is inferred from biochemical 

measurements and cholinesterase activity is not assessed concurrently.

Testing the hypothesis that exposure to chlorpyrifos (or any chemical) can inhibit the morphogenic 

function of AChE in vivo and alter brain development is difficult. Neuronal differentiation and the 

subsequent development of axonal and dendritic networks occurs in a temporal- and region-specific 

manner. In the absence of prior information extensive studies would be required to survey brain 

morphology during both the prenatal and postnatal period. In addition, the methods needed to assess 

the morphology of axonal and dendritic growth in vivo and detect potentially subtle chemical-induced 

changes are tedious and time consuming. Nevertheless, some attempts have been made to detect 

chlorpyrifos-induced changes in neurite growth in vivo. In particular, several laboratories have examined 

the effect of chlorpyrifos and chlorpyrifos oxon on neuronal morphogenesis in developing zebrafish. 

Jacobson et al. (2010) exposed zebrafish embryos at 3 hr post fertilization to 300 nM chlorpyrifos oxon. 

When examined 24 hr later (1 day post fertilization), AChE was inhibited by 50%. Gross morphology was 

only slightly affected and muscular development (including the neuromuscular junction) was normal; 

however, there was a decrease in the number of Rohon-Beard sensory neurons accompanied by 

abnormal extension of their axons. As a follow up to this study and to their in vitro work described 

above, Yang et al. (2011) exposed zebrafish embryos from 24 to 72 hr post fertilization to chlorpyrifos 

and chlorpyrifos oxon. Chlorpyrifos had no effect, but chlorpyrifos oxon significantly decreased AChE 

activity (30 nM), decreased touch-induced swimming (100 nM) and inhibited axon growth from both 

sensory neurons and motor neurons (100 – 1000 nM). These two studies show that exposure to 

chlorpyrifos oxon in an intact, developing organism can alter neurite outgrowth. However, because 

catalytic activity was also inhibited, it is not clear that this effect was solely due to disruption of the 

morphogenic role of AChE. 

As noted above, neurite outgrowth has been assessed indirectly in vitro. This method has also been 

applied in vivo in mammals. Qiao et al. (2003) exposed rats to 1 or 5 mg/kg chlorpyrifos on gestational 

days 17-20. In a previous study from the same laboratory this exposure paradigm resulted in a nonsignificant 5% decrease in AChE activity after 1 mg/kg and a significant 50% decrease in AChE activity 

after 5 mg/kg when assessed 24 hr after the last dose (Qiao, Seidler, Padilla, & Slotkin, 2002). The ratio 

of membrane protein to total protein in several brain regions was assessed as a surrogate index of 

neurite outgrowth. The results indicated that there was no effect of treatment during early postnatal 

development (PND 4-21), but that both doses decreased this ratio on PND 30 and 60. While the authors 

interpret this biochemical measurement as a decrease in neurite growth, there has not been data 

presented to correlate the changes in the membrane protein to total protein ratio with actual 

quantification of neurite length. In addition, no effects were observed early in development when axons 

and dendrite growth is high.

In summary, while perturbation of the morphogenic activity of AChE is a plausible adverse outcome 

pathway for chlorpyrifos, a number of questions remain. There is substantial evidence for a 
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morphogenic role of AChE in nervous system development distinct from its role as an esterase to 

hydrolyze acetylcholine. In vitro evidence suggests that AChE can regulate neurite outgrowth, and that 

cholinesterase inhibitors (including chlorpyrifos and its metabolites) can interfere with this process at 

concentrations that do not inhibit the esterase activity. There is, however, no direct evidence showing 

that disruption of the morphogenic function of AChE can alter axon or dendritic growth in vivo. While 

limited in vivo studies using zebrafish indicate that chlorpyrifos or its metabolite chlorpyrifos oxon can 

disrupt axonal growth, it has not been demonstrated that this effect is due to alteration of the 

morphogenic function of AChE versus other potential mechanisms.

3.2.1.2 Cholinergic system 

There are several lines of evidence showing that signaling through cholinergic receptors is involved in 

neurodevelopment. Activation of muscarinic and/or nicotinic cholinergic receptors can regulate neural 

progenitor cell proliferation and differentiation (Resende & Adhikari, 2009), and in vivo studies 

demonstrate that cholinergic signaling is likely involved in brain morphogenesis (Hohmann & BergerSweeney, 1998). While cholinesterase inhibitors can affect cholinergic signaling by inhibition of the 

catalytic activity of AChE and subsequent increase in acetylcholine, some, including chlorpyrifos and 

chlorpyrifos oxon, can also directly interact with cholinergic receptors. Thus, direct interaction with 

cholinergic receptors by chlorpyrifos represents a potential adverse outcome pathway for disruption of 

neurodevelopment distinct from AChE/ChE inhibition.

Some OPs have been shown to directly interact with cholinergic muscarinic receptors at relatively low 

concentrations. The muscarinic receptors are members of the G-protein receptor family and five 

subtypes (m1-m5) have been identified. Volpe et al. (1985) reported that nanomolar concentrations of 

several OPs decreased binding of quinuclidinyl benzilate (QNB), which binds equally to all five subtypes. 

This effect was noncompetitive and occurred at only a small fraction of the total QNB binding sites. A 

further study from the same laboratory proposed these receptor subtypes to be m2 and/or m3 (L. S. 

Katz & Marquis, 1989). Bakry et al. (1988) found that a number of OP nerve agents and an OP

therapeutic agent inhibited QNB binding to rat brain membranes with low potency, but competitively 

inhibited binding of cis-methyldioxolane (CD) with high potency. In rat cardiac tissue, OPs also 

competitively inhibited CD binding with very high potency (Silveira, Eldefrawi, & Eldefrawi, 1990). CD is a 

muscarinic agonist that binds to the high affinity state of the m2 receptor in mammalian brain and heart 

(Closse, Bittiger, Langenegger, & Wanner, 1987; Vickroy, Roeske, & Yamamura, 1984; Watson, 1986). 

Subsequent research using Chinese hamster ovary (CHO) cells transfected with cDNA for the five distinct 

genes for muscarinic receptors (ml-m5) has identified the CD binding site as the high affinity state (GDP 

bound G-protein) of the m2 receptor (Huff & Abou-Donia, 1994). In light of these findings, Ward et al. 

(1993) examined the relationship between cholinesterase inhibition and direct binding to muscarinic 

receptors for a series of OPs and their active “oxon” metabolites. The results indicated a strong 

correlation between anticholinesterase activity of OPs, including chlorpyrifos and chlorpyrifos oxon, and 

the ability to compete for CD binding sites (m2 receptors) in rat brain homogenates. Binding affinities of 

the oxons were in the nanomolar range, at or below concentrations that inhibited AChE (Huff, Corcoran, 

Anderson, & Abou-Donia, 1994); specifically, chlorpyrifos oxon had a binding affinity of 22 nM in rat 

striatum and 2 nM in rat cortex (Huff, et al., 1994; Ward & Mundy, 1996). In total, these studies suggest 

that direct interactions with muscarinic receptors, and especially the m2 subtype, represent an 

alternative site of action for OPs including chlorpyrifos and chlorpyrifos oxon, with the oxon forms 

having high affinity.
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Further studies determined whether binding of OPs to the m2 receptor had functional consequences on 

downstream second messenger signaling. The m2 and m4 receptors are coupled to an inhibitory Gprotein (Gi

) and activation inhibits adenylate cyclase and decreases cAMP formation. In contrast, m1, 

m3, and m5 receptors increase PI hydrolysis via the stimulatory G-protein Gp (Lameh et al., 1990). Jett et 

al. (1991) reported that paraoxon could inhibit both CD binding and cAMP formation in rat striatum. The 

effects of paraoxon were similar to the classical muscarinic agonist carbachol, and were blocked by the 

muscarinic antagonist atropine. Similar results were observed in rat frontal cortex for paraoxon and 

malaoxon (Ward & Mundy, 1996). In addition, Ward and Mundy (1996) observed that the OPs had no 

effect on PI hydrolysis. These data suggest that OPs can act as agonists at the m2 receptor (but not the 

m1, m3, or m5 receptor) and decrease cAMP formation. Huff et al. (1994) extended these findings to 

chlorpyrifos oxon, which had an IC50 for inhibition of cAMP of 155 nM. Unlike the other OPs, however,

the effects of chlorpyrifos oxon were not completely blocked by atropine; an observation confirmed by 

Ward and Mundy (1996). Thus, like other OPs chlorpyrifos oxon can act directly as an agonist at the m2 

receptor, but may also act at another site downstream of the receptor to inhibit adenylate cyclase.

Cholinergic receptor signaling has been shown to be involved in apoptosis, cell proliferation and 

neuronal differentiation (Resende & Adhikari, 2009). While the m2 receptor subtype is widely expressed 

in proliferating neuroepithelial cells in the ventricular zone of embryonic brain (Ma, Li, Zhang, & Pant, 

2004), evidence for a specific role of in the modulation of neurodevelopment is derived primarily from in 

vitro studies. Neural precursor cells derived from embryonic rat cortex expressed the m2 receptor, and 

exposure to muscarinic agonists increased cell proliferation and enhanced differentiation into neurons 

(Ma et al., 2000). Using a P19 embryonic cell line as a model for neurogenesis, Resende et al. (2008)

used pharmacologic agonists and antagonists to demonstrate the ability of m2 receptors to induce cell 

differentiation. The m2 receptor is highly expressed in developing cells of the dorsal root ganglia, and is 

thought to regulate both neuronal and non neuronal cell differentiation (Biagioni, Tata, De Jaco, & 

Augusti-Tocco, 2000; Tata, Cursi, Biagioni, & Augusti-Tocco, 2003). Another line of evidence for 

involvement of m2 receptor signaling in neurodevelopment is the role of cAMP in neurite outgrowth. 

Studies in both cell lines and primary neuronal cultures show that activation of adenylate cyclase and 

subsequent formation of cAMP stimulate neurite outgrowth (Kamei & Tsang, 2003; Mattson, Guthrie, & 

Kater, 1988), while inhibition of adenylate cyclase decreases neurite outgrowth (Tam, Rosenberg, & 

Maysinger, 2006; Wong, Bruch, & Farbman, 1991). Thus, in vitro evidence supports a role of m2 

receptor signaling in neurodevelopment at the cellular level. However, evidence showing that 

manipulation of m2 receptors can alter brain development in vivo is lacking.

Because they can also influence neurodevelopment, it should be noted that there are several studies 

demonstrating that OPs including chlorpyrifos and chlorpyrifos oxon can bind directly to and desensitize 

nicotinic receptors (E. J. Katz, Cortes, Eldefrawi, & Eldefrawi, 1997; Smulders, Bueters, Vailati, van Kleef, 

& Vijverberg, 2004). This binding, however, occurs at relatively high concentrations (5 – 30 uM) and has 

not been demonstrated in neuronal tissue.

In summary, in vitro studies have shown that chlorpyrifos oxon can bind to and activate m2 receptors at 

levels similar to those that inhibit AChE activity. Other work has shown that m2 receptor signaling can 

regulate various aspects of neurodevelopment. Together, the studies cited above outline a plausible 

adverse outcome pathway for chlorpyrifos and chlorpyrifos oxon to affect brain development via actions 

at the m2 subtype of muscarinic receptors. However, while there are studies showing that chlorpyrifos 

oxon can affect neurite outgrowth in vitro and decrease cell proliferation and differentiation both in 

vitro (Jameson, Seidler, Qiao, & Slotkin, 2006; Qiao, Seidler, & Slotkin, 2001; Song, et al., 1998) and in 
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vivo (Dam, Seidler, & Slotkin, 1998; Qiao, et al., 2003), there is no experimental evidence that these 

effects are a result of direct actions on the m2 receptor.

3.2.1.3. Endocannabinoid system

Several lines of research have suggested that disruption of the endocannabinoid (EC) system due to 

chlorpyrifos exposure could play a role in its acute and/or long-term toxicity, and could also be extended 

to potential developmental toxicity. The EC system modulates neurotransmission as well as playing a 

morphogenic role during development of the nervous system. Chemicals e.g., drugs of abuse, which act 

on this system, produce long-term neurodevelopmental disorders in animal models and human studies. 

Chlorpyrifos also interacts with this system, both in vitro and in vivo. By this reasoning, the EC system 

represents a possible adverse outcome pathway for developmental effects of chlorpyrifos.

The EC system includes inhibitory G-protein-coupled receptors (CB1 and CB2), endogenous ligands (2-

arachidonoylglycerol, or 2-AG, and anadamide, or AEA), and hydrolases (monoacylglycerol, or MAG, and 

fatty acid amide hydrolase, or FAAH) that end the receptor actions of these ligands. CB1 receptors are 

predominant in certain brain regions, and the EC system modulates neuronal transmission and is 

involved in several physiological processes including appetite, pain sensation, mood, and cognition 

(Wilson & Nicoll, 2002). Emerging lines of research indicate that during development, this system plays 

a major role in controlling the morphological and functional specification of neurons (e.g., Fride, 2008; 

Harkany et al., 2007; Harkany, Keimpema, Barabas, & Mulder, 2008; Harkany, Mackie, & Doherty, 2008); 

reviewed in Campolongo, Trezza, Palmery, Trabace, & Cuomo, 2009). Its involvement in neurogenesis, 

proliferation, and axon guidance suggests a role in overall neuronal connectivity. 

There are epidemiological data on neurodevelopmental outcomes following prenatal exposure to 

cannabinoids such as cannabis and THC through maternal drug abuse, but interpretations are often 

somewhat difficult given co-exposures to other neurotoxicants (e.g., ethanol). Despite limitations, an 

overall picture has emerged of long-lasting impaired cognitive processes, including attention and 

problem-solving deficits, as well as anxiety and depressive symptoms in humans (e.g., Campolongo, et 

al., 2009; Fried & Smith, 2001; Jutras-Aswad, DiNieri, Harkany, & Hurd, 2009; Trezza, Cuomo, & 

Vanderschuren, 2008). In addition, animal models have indicated various changes in neurotransmission 

of catecholaminergic and indolaminergic systems, ontogeny of motor function, and cognitive function in 

adults developmentally exposed to cannabis as well as other CB1 agonists (e.g., WIN55,212-2) (reviewed 

in Campolongo, Trezza, Ratano, Palmery, & Cuomo, 2011; Marco et al., 2009). This supports the 

possibility that other chemicals, unrelated to cannabis, that also act on the EC system early in 

development could lead to long-term neurological dysfunction. Despite the number of studies in this 

area, however, the specific responsible cellular responses have not been delineated, and most studies 

have not evaluated basic dose-response such as correlations of the degree of receptor binding with any 

structural or functional outcomes.

In addition to drugs of abuse (e.g., cannabis), there are several classes of chemicals that act on the EC 

system, including organosulfonyl fluorides and some OPs (e.g., Casida & Quistad, 2004; Segall, Quistad, 

Sparks, Nomura, & Casida, 2003). OPs phosphorylate and thereby block the action of the hydrolases 

FAAH and MAG. This inhibition of the hydrolysis of endogenous cannabinoids could prolong their 

actions on the receptor, which inhibits several neurotransmitter systems, including ACh. OPs also bind 

the CB1 receptor and block the binding of specific receptor ligands. The OPs do not bind at the agonist 

site, however, and may have no inherent agonist or antagonists consequences; however, this has not 

been adequately studied. Thus, prolongation of the ligand residence time, plus blockade of ligand 
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binding at the receptor, could lead to downstream modulation of cholinergic activity and signaling, 

separate from AChE inhibition. It is unclear which actions could predominate in vivo.

Casida and colleagues have produced several papers relating FAAH inhibition to that of AChE and 

neurotoxic esterase (NTE), and considered its involvement in acute cholinergic, intermediate, and 

delayed toxicity syndromes (e.g., Casida & Quistad, 2004; Quistad, Klintenberg, Caboni, Liang, & Casida, 

2006; Quistad, Nomura, Sparks, Segall, & Casida, 2002; Quistad, Sparks, & Casida, 2001; Quistad, Sparks, 

Segall, Nomura, & Casida, 2002; Segall, et al., 2003). Several studies have included chlorpyrifos. They 

have reported that in vitro chlorpyrifos oxon inhibits FAAH and MAG with IC50s of 40 nM and 34 nM, 

respectively, compared to an AChE IC50 of 19 nM. Other OPs (paraoxon, dichlorvos, diazoxon) have 

higher hydrolase inhibitory actions in vitro (IC50s 540-14,000 nM). In vitro receptor binding assays of 

CB1 agonists reveal an IC50 of 14 nM for chlorpyrifos oxon, and higher IC50s (>64 nM) for other OPs 

(chlorpyrifos methyl oxon, diazoxon, dichlorvos). 

Casida’s group has also dosed adult mice with OPs and evaluated motor behavior and cholinergic signs, 

followed by ex vivo biochemical assays. These studies report inhibition of FAAH activity and CB1 binding 

only at highly toxic doses of chlorpyrifos and/or chlorpyrifos oxon. Chlorpyrifos dose-response differs 

somewhat across studies, but in general, FAAH is inhibited more than MAG and AChE, and CB1 binding is 

moderately inhibited, but only at doses producing cholinergic signs. Chlorpyrifos inhibition of FAAH 

does not correlate with the hypomotility (akinesia, rigidity) produced by the EC agonist amandamide. 

Furthermore, chlorpyrifos does not potentiate the effects of amandamide, suggesting that it does not 

appear to have functional impacts in the EC system. In contrast, treatment with other chemicals that do 

produce marked (>76%) FAAH inhibition does indeed potentiate anandamide effects in mice; however, 

chlorpyrifos doses that would be required to produce this level of FAAH inhibition are likely to be lethal. 

This line of research has led Casida and colleagues to conclude that chlorpyrifos actions on the EC 

system in vivo are less important to its acute toxicity than AChE inhibition. 

A series of studies from the laboratories of Pope and colleagues has evaluated the influence of CB1 

agonists administered in vivo on the signs of toxicity produced by several OPs (Baireddy, Liu, Hinsdale, & 

Pope, 2011; Nallapaneni, Liu, Karanth, & Pope, 2006, 2008; C. Pope, Mechoulam, & Parsons, 2010; 

Wright, Liu, Nallapaneni, & Pope, 2010). They reported that CB1 agonists and other cannabinomimetics 

block the acute signs of exposure to paraoxon or DFP (chlorpyrifos was not tested). Studies with 

chlorpyrifos show increased extracellular levels of 2-AG but not amandamide in hippocampus. 

Somewhat different conclusions were drawn, however, in a paper showing that the toxicity and AChE 

inhibition produced by an acute dose of chlorpyrifos is not altered in mice without the CB1 receptor 

(knockout) compared to their wild type littermates. Pope et al. also measured chlorpyrifos oxoninduced ACh release from these tissues and saw only a small difference. These somewhat contradictory 

results suggest that potential EC involvement in the acute toxicity of certain OPs is still uncertain.

Only one study has examined chlorpyrifos effects on the EC system in developing animals. Carr et al. 

(2011) dosed preweanling rats for 5 days (1-5 mg/kg/day, p.o.) and took brain tissue 4 hr after the last 

dose. The lowest dose produced 40% inhibition of FAAH, 14% inhibition of MAG, and only 18% 

inhibition of AChE; in addition, the highest dose eliminated FAAH activity and produced about 52-55% 

inhibition of AChE. This suggests a greater sensitivity of the EC system, at least in terms of the hydrolase 

compared to AChE activity, in the pups. However, there was no time-course, dose-response, or other 

ages tested, and no downstream or correlative measure of changes in EC system function. Additional 

studies along these lines are needed.
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In summary, most of the research in this area has evaluated acute responses in the EC system, even 

though such changes during a critical period of development could have long-term consequences. 

While some steps along a chlorpyrifos adverse outcome pathway for developmental neurotoxicity are 

plausible, there are remaining questions as to whether the EC system is sufficiently sensitive to low 

doses of chlorpyrifos to alter its function during development. Despite the increasing understanding of 

the EC system, there are no data during development on 1) inhibition of the CB1 receptor at any time 

during development (animal studies used only agonists); 2) inhibition of FAAH during development; 3) 

time-course and dose-response of OPs in vivo compared to AChE inhibition; and 4) whether these 

actions on the receptors and/or hydrolases actually change EC signaling to a point that would impact 

downstream nervous system development.

3.2.1.4. Reactive Oxygen Species

The production of reactive oxygen species (ROS) and resulting cellular damage has been proposed as a 

mechanism for a wide variety of neurotoxicants. Due to lower levels of protective enzymes and 

antioxidants, and relatively low numbers of glia relative to the adult, the developing brain may be 

particularly sensitive to neural cell damage caused by oxidative stress. In addition, recent work suggests 

that ROS can act as second messengers. Relatively small changes the oxidative status of the cell (redox 

potential) can lead to changes in redox sensitive signaling pathways that regulate cell physiology. In the 

nervous system, redox signaling is involved in the regulation of neurodevelopmental processes including 

neural stem cell proliferation and differentiation (Le Belle et al., 2011; Vieira, Alves, & Vercelli, 2011). A 

number of studies suggest that chlorpyrifos and chlorpyrifos oxon can induce oxidative stress in various 

neural cell types. Thus, generation of reactive oxygen species and/or alteration of cellular redox 

potential by chlorpyrifos represent a possible initiating event leading to developmental neurotoxicity.

Both chlorpyrifos and chlorpyrifos oxon have been shown to induce oxidative stress in neuronal cells in 

vitro. A series of studies using PC12 cells were performed by the Slotkin laboratory. Crumpton et al. 

(2000) reported that acute exposure of undifferentiated PC12 cells to 1.5 – 150 uM chlorpyrifos for 10 

min resulted in a rapid increase in ROS production. Exposure to 30 uM chlorpyrifos oxon for 10 min had 

no effect. Consistent with the generation of ROS, further studies demonstrated an increase in oxidative 

damage (lipid peroxidation) in both undifferentiated and differentiated PC12 cells after exposure to 

chlorpyrifos (3 – 100 uM) for 24 hr (Qiao, Seidler, & Slotkin, 2005). Finally, Slotkin and Seidler (2009)

used microarrays to examine the expression of genes involved in the response to oxidative stress after 

exposure of PC12 cells to 30 uM chlorpyrifos for 24 or 72 hr. They observed a significant effect on 

transcription of genes including glutathione peroxidase, glutathione-S-transferase, and superoxide 

dismutase, which was greater in differentiated cells compared to undifferentiated cells. PC12 cells were 

also used by a separate group of researchers who reported that very high concentrations of chlorpyrifos 

(1.4 – 14 mM) increased ROS production during a 5 hr exposure period (Geter et al., 2008). ROS 

production has also been reported in other neuronal cell types. Giodano et al. (2007) examined the 

effects of chlorpyrifos and chlorpyrifos oxon on ROS levels and lipid peroxidation in primary cultures of 

cerebellar granule cells from mice. Exposure to 1 uM chlorpyrifos for 60 min increased both ROS and 

lipid peroxidation, and a similar response was observed with 1 uM chlorpyrifos oxon. Saulsbury et al. 

(2009) examined oxidative stress in oligodendrocyte progenitor cells, which are the precursors to glial 

cell types in the brain. Exposure to 15 – 120 uM chlorpyrifos for 3 hr resulted in a significant increase in 

ROS production. These in vitro studies suggest that exposure of cells to the micromolar concentrations 

of the parent compound chlorpyrifos (which is not a potent AChE inhibitor) can result in oxidative stress. 

The active metabolite chlorpyrifos oxon was found to be active in cerebellar granule cells but not in 

PC12 cells. Inhibition of AChE was not determined in these studies.
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A limited number of studies have examined the oxidative stress response in rat brain after in vivo

exposure to chlorpyrifos. Bagchi et al. (1995) dosed adult female rats p.o. twice with 41 mg/kg 

chlorpyrifos, once at 0 hr and once at 21 hr. Animals were sacrificed at 24 hr. Both ROS production and 

lipid peroxidation were increased in homogenates from whole brain. In a follow-up study, the same 

dosing paradigm resulted in an up-regulation of heat shock protein, which in some cases can be a 

response to oxidative stress (Bagchi, Bhattacharya, & Stohs, 1996). There are two studies of oxidative 

stress in response to chlorpyrifos exposure in developing animals. Slotkin et al., (2005) used three dosing 

paradigms: s.c. exposure of pregnant rats on GD 17-20 followed by collection of the fetal brain on GD 

21, s.c. exposure of rat pups on PND 1-4 followed by brain collection on PND 5, or s.c. exposure of rat 

pups on PND 11-14 followed by brain collection on PND 15. Lipid peroxidation was measured in brain 

regions including the brainstem, forebrain and cerebellum. There was a 20% increase in lipid 

peroxidation in the brain only after exposure to 5 mg/kg chlorpyrifos during the later postnatal period 

(PND 11-14) and only in males. AChE inhibition was not assessed. Ray et al. (Ray, Liu, Ayoubi, & Pope, 

2010) dosed 7 day old rat pups p.o. with 0.1, 0.5, 1, or 2 mg/kg chlorpyrifos and examined gene 

expression in the forebrain 24 hr later using microarrays. Only the highest dose inhibited AChE (25%). 

Gene expression changes were observed at all doses, and ‘oxidative stress’ was one of the six canonical 

pathways that was significantly affected.

Data from both in vitro studies with neuronal cells (including neural precursors) and in vivo studies in 

developing brain demonstrate chlorpyrifos can induce oxidative stress. The in vitro data suggests that 

this effect may not be due to AChE inhibition, since the parent compound chlorpyrifos is either 

equipotent or more potent than the oxon. There was, however, no concurrent analysis of AChE 

inhibition in most of these studies. Several known developmental neurotoxicants have been shown to 

disrupt neural precursor cell proliferation in vitro through a common pathway that is initiated by 

increasing the oxidative state of the cell (Z. Li, Dong, Proschel, & Noble, 2007), and the antioxidant 

vitamin E protected PC12 cells from the anti-proliferative effect of chlorpyrifos (Slotkin, MacKillop, Ryde, 

& Seidler, 2007). Thus, the in vitro data suggest that chlorpyrifos can affect a critical 

neurodevelopmental process, at least in part, via generation of ROS. Though limited, in vivo studies 

show both direct evidence (lipid peroxidation) and indirect evidence (alteration in the expression of 

oxidative stress response genes) of oxidative stress in the developing brain after exposure to 

chlorpyrifos. Recent evidence suggests that oxidative stress can alter neurodevelopment in vitro and in 

vivo by the dysregulation of signaling pathways controlling neuroprogenitor cell function (Le Belle, et al., 

2011; Vieira, et al., 2011). Thus, while there is the potential for initiation of an adverse outcome 

pathway via induction of oxidative stress, it has yet to be demonstrated in vivo that treatments 

preventing the induction of oxidative stress can ameliorate the developmental neurotoxicity of 

chlorpyrifos.

3.2.1.5. Serotonergic system

The serotonergic system (reviewed in Rho & Storey, 2001; Serotonin Receptors in Neurobiology, 2007)

includes 7 distinct families of serotonin receptors regulating diverse functions such as neuronal 

excitability, appetite control, memory, learning, circadian rhythm, sexual behavior, anxiety, aggression, 

and respiration. While the majority of the receptors are either negatively or positively coupled to 

adenylate cyclase, the 5-HT3 serotonergic receptor is a ligand-gated ion channel. Serotonin receptors are 

distributed throughout the brain and body. Synthesized from L-tryptophan, the rate limiting step in 

serotonin biosynthesis is catalyzed by tryptophan hydroxylase. After release, the action of serotonin is 

primarily terminated by reuptake into the presynaptic terminal via monoamine transporters. Various 
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drugs inhibit the reuptake of serotonin, effectively increasing the available synaptic concentrations and 

actions: Ecstasy, amphetamine, cocaine, tricyclic antidepressants and the newer antidepressants: SSRIs 

(selective serotonin reuptake inhibitors).

Beyond its classical neurotransmitter actions (described above), serotonin has other roles during 

development. In their review, Thompson and Stanwood (Thompson & Stanwood, 2009) described 

serotonin as a pleiotropic molecule, meaning that it can produce multiple, diverse effects, regulating 

different functions at different times during development. The serotonergic system is integral in many 

developmental processes including, but not limited to, neurogenesis, migration, and differentiation, 

synaptogenesis, and cardiac development before assuming its more well-known function as a 

neurotransmitter in the adult nervous system (reviewed in (Frederick & Stanwood, 2009). Serotonin also 

plays crucial roles in thalamocortical patterning (reviewed in (Frederick & Stanwood, 2009). As serotonin 

is present extremely early in development, it is thought that it modulates cellular function even before 

neurogenesis. Later in development, serotonin is temporarily taken up by so-called transient 

serotonergic neurons mainly involved in sensory processing, and is involved in activity-dependent 

patterning of the brain. Later in development, serotonin has also been shown to modulate 

differentiation in the brain. 

Perturbation of the development of the serotonergic system in laboratory animals causes permanent 

changes in the neurochemical, behavioral and structural aspects of the adult nervous system (reviewed 

in Borue, Chen, & Condron, 2007; Daubert & Condron, 2010; Frederick & Stanwood, 2009; Oberlander, 

Gingrich, & Ansorge, 2009). Moreover, it is generally recognized that the development stage of 

serotonergic perturbation determines the developmental outcome (reviewed in Oberlander, et al., 

2009; Olivier, Blom, Arentsen, & Homberg, 2011), meaning that the time at which the developing animal 

experienced the insult to the serotonergic nervous system would determine the degree and the result. 

The review by Daubert and Condron (2010) includes a table summarizing many laboratory studies where 

different aspects of the serotonergic system were manipulated (either through knock-out, conditional 

knock-out, or chemical treatment) during development accompanied by the behavioral and 

morphological results of that manipulation. For example, SERT (serotonin reuptake receptor) knock-out 

mice had alterations in sensory patterning, sleep pattern abnormalities, and long-term behavioral 

deficits (Q. Li, 2006), and animals treated with SSRIs during development show many similarities to the 

SERT knock-out mice. Developmental exposure to SSRIs in mice produced permanent (adulthood) 

effects on behavior (Lisboa, Oliveira, Costa, Venâncio, & Moreira, 2007; Noorlander et al., 2008) and 

brain chemistry (Noorlander, et al., 2008) and changes in morphology in the hippocampus (J. Zheng et 

al., 2011) and somatosensory cortex (L. J. Lee, 2009). 

Because of the extensive usage of SSRI antidepressants in the human population, there is a rich 

literature SSRI effects during pregnancy and lactation (reviewed in Alwan & Friedman, 2009; Ellfolk & 

Malm, 2010; Gentile, 2005; Oberlander, et al., 2009; Olivier, et al., 2011; Tuccori et al., 2009). The main 

effects of SSRI usage are on cardiac development, and while there is some support for concluding that 

actions on the serotonergic system early in development can lead to long-term neurological dysfunction, 

there are only few studies that have studied this relationship in humans in detail. A recent study has 

noted that children born to mothers who used SSRIs have increased likelihood for social-behavioral 

abnormalities (Klinger et al., 2011). 

There are numerous studies of the effects of perinatal chlorpyrifos administration on the patency of the 

serotonergic system. Endpoints in various brain regions include serotonin levels (Aldridge, Meyer, 

Seidler, & Slotkin, 2005a; Slotkin & Seidler, 2007b, 2007c), serotonin turnover (Aldridge, Meyer, et al., 
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2005a; Slotkin & Seidler, 2007b, 2007c), serotonin receptor levels (Aldridge, Meyer, Seidler, & Slotkin, 

2005b; Aldridge, Seidler, Meyer, Thallai, & Slotkin, 2003; Aldridge, Seidler, & Slotkin, 2004; Slotkin & 

Seidler, 2005), serotonin reuptake receptor levels (Aldridge, Meyer, et al., 2005b; Aldridge, et al., 2003; 

Aldridge, et al., 2004; Slotkin & Seidler, 2005), serotonin elicited second messenger activity (Aldridge, 

Meyer, et al., 2005b; Aldridge, et al., 2003; Aldridge, et al., 2004), gene expression of serotonin receptor 

and metabolism related genes (Slotkin & Seidler, 2007a), serotonin related behavioral assessments 

(Aldridge, Levin, Seidler, & Slotkin, 2005; W.-Q. Chen et al., 2011; Venerosi, Ricceri, Rungi, Sanghez, & 

Calamandrei, 2010), and behavior after serotonergic drug challenge (Aldridge, Levin, et al., 2005; 

Venerosi, et al., 2010). Chlorpyrifos was administered during gestation (Aldridge, Meyer, et al., 2005a; 

Aldridge, et al., 2003; Aldridge, et al., 2004; Slotkin & Seidler, 2007c; Venerosi, et al., 2010), postnatally 

(Aldridge, Levin, et al., 2005; Aldridge, Meyer, et al., 2005a, 2005b; Aldridge, et al., 2003; Aldridge, et al., 

2004; Slotkin & Seidler, 2005, 2007a, 2007b) or during adolescent (W.-Q. Chen, et al., 2011) and the 

animals were assessed either shortly after dosing ceased (Aldridge, et al., 2003; Slotkin & Seidler, 

2007a), or after a matter of weeks (Aldridge, Meyer, et al., 2005b; Aldridge, et al., 2003; W.-Q. Chen, et 

al., 2011; Slotkin & Seidler, 2007b, 2007c) or in adulthood (Aldridge, Meyer, et al., 2005a; Aldridge, et 

al., 2004; Slotkin & Seidler, 2005; Venerosi, et al., 2010). All the data indicate that there are acute, as 

well as permanent, effects of neonatal chlorpyrifos treatment on the maturation of the serotonergic 

nervous system. The effects are often gender-specific, region-specific and dose-related. For example, a 

group of papers found permanent effects of late gestational chlorpyrifos administration in rats 

(Aldridge, Meyer, et al., 2005a; Aldridge, et al., 2004) or mice (Venerosi, et al., 2010) on the 

development of the serotonergic nervous system. In rats (Aldridge, Meyer, et al., 2005a; Aldridge, et al., 

2004), chlorpyrifos was administered on GD17-20 (1 or 5 mg/kg, sc in DMSO), and the adult brain 

showed changes in brain regional serotonin content (5 mg/kg group), increased serotonin turnover (1 

and 5 mg/kg groups), increases in 5-HT1A and 5-HT2 receptor populations (1 mg and 5 mg/kg groups), 

increases in serotonin reuptake receptor (1 and 5 mg/kg groups), and changes in the adenylate cyclase 

response to serotonin in the cerebral cortex and mid brain (1 mg/kg group)—all indicative of permanent 

changes in the serotonergic tone of the adult brain. In a related study in mice (Venerosi, et al., 2010), 

chlorpyrifos was also administered during late gestation, but at a higher dose and by a different route 

(GD14-17; 6 mg/kg, peanut oil, gavage); the offspring were assessed upon reaching adulthood (≅ 90 

days). The chlorpyrifos-treated rats showed minor differences in the behavioral tests, but when the tests 

were combined with a fluvoxamine (serotonin reuptake inhibitor) pharmacological challenge, there 

were marked differences in responses between the animals that had been exposed to chlorpyrifos 

during gestation and those that had not—another indication that developmental chlorpyrifos exposure 

had permanently altered the serotonergic tone of the adult brain. 

There is ample evidence that chlorpyrifos exposure during development causes permanent changes in 

the serotonergic nervous system; there are, however, few papers that assessed concurrently the 

cholinesterase inhibition (either brain or blood) in those same animals. In some cases, although 

cholinesterase activity was not assessed concurrently, a dosing regimen was used that had been 

characterized previously with regard to cholinesterase activity. It does appear, however, that most of 

the studies on the effects of chlorpyrifos on the serotonergic nervous system were conducted with 

doses of chlorpyrifos that likely produced marked inhibition of cholinesterase activity. There is, however, 

one of the dosing regimens [GD17-20 (daily subcutaneous dosing with chlorpyrifos in DMSO] that has 

been shown to produce no fetal brain cholinesterase inhibition measured 24 hours after the last dose at 

the 1 mg/kg , and inhibition (~20%) at 2 mg/kg (Qiao, et al., 2002). As noted above, this level of 

gestational exposure to chlorpyrifos permanently altered the development of the serotonergic nervous 

system (Aldridge, Meyer, et al., 2005a; Aldridge, et al., 2004).
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In summary, with regard to development of the chlorpyrifos adverse outcome pathway for the 

serotonergic nervous system: the serotonergic system is intregal to mammalian brain development and 

function; perturbation of that system during development leads to lasting effects in mammals; and 

exposure to chlorpyrifos during the perinatal period can permanently change the function of 

serotonergic system. Therefore, as many steps in this chlorpyrifos adverse outcome pathway are 

possible and plausible, and in laboratory animals the serotonergic nervous system is sufficiently 

sensitive to low doses of chlorpyrifos during development to alter its function, it is plausible that 

exposure to chlorpyrifos during development could alter brain development and the function of the 

serotonergic nervous system. Although chlorpyrifos effects on the serotonergic nervous system in 

laboratory animals likely is initiated within 24 hours (Slotkin & Seidler, 2007a), the actual initiating event 

of this potential adverse outcome pathway is unknown. 

3.2.1.6. Tubulin, Microtubule Associated Proteins and Axonal Transport 

Microtubules, one component of the dynamic cytoskeletal scaffolding within each cell, are composed of 

heterodimers of α- and β-tubulin, as well as microtubule associated proteins. The microtubule 

associated proteins appear to have three main functions: (1) to stabilize the microtubules; (2) to aid in 

tubulin dissociation and (3) to act as motor proteins moving substances forward and backward along the 

microtubules (Avila, Dominguez, & Diaz-Nido, 1994; Pellegrini & Budman, 2005; Sánchez, Díaz-Nido, & 

Avila, 2000). Not only does the microtubule cytoskeleton determine neuronal morphology (Matus, 

1988, 1990; Sánchez, et al., 2000), but the dynamic reorganization of the microtubules and microtubule 

associated proteins within a cell may also coordinate neurite extension/retraction, as well as growth 

cone advancement. In addition to these integral roles in brain structure and growth, microtubules and 

the microtubule associated motor proteins kinesin (Hirokawa & Noda, 2008) and dynein (Vallee, 

Williams, Varma, & Barnhart, 2004) also provide a “railway” for transport of materials throughout the 

cell, i.e., axonal transport (Fukushima, Furuta, Hidaka, Moriyama, & Tsujiuchi, 2009), another process 

which is integral to the health of the central and peripheral nervous system, playing a pivotal role in 

neuronal network formation and synapse maturation (Hirokawa & Takemura, 2004). 

During brain development, the pattern of expression of microtubules (Bond & Farmer, 1983; Denoulet, 

Edde, & Gros, 1986; Meininger & Binet, 1989; Moskowitz & Oblinger, 1995), the post-translational 

modification of the tubulin (Fukushima, et al., 2009; M. K. Lee, Rebhun, & Frankfurter, 1990), and the 

spectrum of microtubule associated proteins (Fischer & Romano-Clarke, 1990; Nunez, 1986; Sánchez, et 

al., 2000) are all developmentally regulated. Because the phosphorylation of both microtubules and 

microtubule associated proteins are integral to their functional roles in the nervous system (Avila, et al., 

1994; Fischer & Romano-Clarke, 1990; Fukushima, et al., 2009; Sánchez, et al., 2000), there is concern 

that exposure to OPs may affect the phosphorylation status and consequently the function of these 

molecules.

One of the shortcomings of constructing this adverse outcome pathway is that there are no studies on 

tubulin dynamics, microtubule associated proteins or axonal transport in developing animals, but 

because tubulin dynamics, microtubule associated proteins and axonal transport are so integral to the 

development and maintenance of the nervous system, the available adult data are first given in detail 

and then summarized below.

Adult rats treated for 14 days with chlorpyrifos (sc in peanut oil; 2.5, 10, 18, or 25 mg/kg; all doses 

produced plasma cholinesterase inhibition) show depressed anterograde and retrograde axonal 

transport in the sciatic nerve 6 days after cessation of treatment 2003 (Terry et al., 2003). This same 
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laboratory followed up in 2007 (Terry et al., 2007) with another, slightly different, study where axonal 

transport was assessed in adult rats given chlorpyrifos repeatedly (sc, 3% DMSO, 97% peanut oil vehicle, 

every other day for 30 days; 18 mg/kg). Depressed bidirectional axonal transport was noted after 1 

dose, after the 30 days of dosing or 2 weeks after cessation of treatment. This dose produced about 30-

50% brain cholinesterase inhibition 14 days after last dose. In that same year, the Terry laboratory 

published an in vitro paper (Gearhart, Sickles, Buccafusco, Prendergast, & Terry, 2007) studying the 

effect of chlorpyrifos or chlorpyrifos oxon on kinesin-dependent microtubule motility (the mechanism of 

anterograde axonal transport). Using tubulin and kinesin prepared from bovine brain, they assessed 

kinesin-mediated movement of microtubules. This was done by either preincubating the tubulin or the 

kinesin fraction with chlorpyrifos or chlorpyrifos oxon and then combining the fractions and assessing 

movement of the microtubules. They obtained the same interesting pattern using chlorypyrifos or 

chlorpyrifos oxon: there was no effect if preincubated with microtubules, but marked effect (increased 

detachment of the microtubules) if preincubated with kinesin. This assay contains no cholinesterase 

activity, thereby precluding inhibition assessment. A recently published in vitro paper (MiddlemoreRisher, Adam, Lambert, & Terry, 2011) assessed the effect of chlorpyrifos or chlorpyrifos oxon transport 

of mitochondria in rat cortical neurons. They found dose-related effects of chlorpyrifos or chlorpyrifos 

oxon on total distance moved and number of moving mitochondria. All doses produced cholinesterase 

inhibition except lowest dose of chlorpyrifos (1 uM) or chlorpyrifos oxon (5 nM), and those doses still 

affected movement. Interestingly the toxic effect was not altered by co-treatment of either a nicotinic or 

muscarinic receptor blocker, an indication that the depressed transport was not mediated through 

cholinergic receptors. Subsequent to demonstrating that in vitro chlorpyrifos oxon appeared to alter 

tubulin dynamics (Grigoryan & Lockridge, 2009) and also covalently bound to tubulin (Grigoryan et al., 

2008), the Lockridge laboratory assessed tubulin and microtubule associated proteins in the brains of 

mice that had been dosed in vivo with chlorpyrifos or chlorpyrifos oxon (Jiang et al., 2010). Repeated (14 

days) dosing with chlorpyrifos (sc, 3% DMSO/97% peanut oil, 3 mg/kg) produced a transient decrease in 

blood AChE activity, no change in blood carboxylesterase activity and a steady decline in blood BuChE 

activity. Brain β-tubulin from the animals treated with chlorpyrifos had been covalently modified 

(diethosyphosphorylated) on tyrosine 281 (the residue also found in the in vitro study (Grigoryan, 

Schopfer, et al., 2009). This same paper also described treating adult mice with a single dose of 

chlorpyrifos oxon (3 mg/kg; ip); this dose produced toxic signs and marked AChE and BuChE inhibition in 

the plasma. Brain tubulin was then purified, polymerized, the proteins separated by gel electrophoresis 

and the proteins identified by mass spectrophometry. The polymerized tubulin was also subjected to 

atomic force microscopy nanoimaging to assess the structure (width and height) of the microtubules. 

Microtubules from chlorpyrifos oxon- treated brains were missing 6 microtubule associated proteins. 

This observation was supported by the nanoimaging showing that microtubules from chlorpyrifos oxontreated animals were narrower and shorter than control brain microtubules and also had fewer 

attached (decorated) proteins than do the microtubules from control brains. It interesting to note that 

an in vitro assessment (Prendergast et al., 2007) of the effects of chlorpyrifos oxon tubulin/microtubule 

associated protein dynamics in a 8 day old rat hippocampal slice preparation showed all concentrations 

of chlorpyrifos oxon decreased mircrotubule associated protein- staining and decreased microtubule 

associated protein mediated tubulin polymerization by about 45%; all chlorpyrifos-oxon doses inhibited 

AChE.

In summary, in the late 2000s, a number of papers were published on the in vitro modification of 

various proteins by chlorpyrifos or chlorpyrifos oxon (Grigoryan, Li, et al., 2009; B. Li et al., 2009), 

including tubulin (Grigoryan, Li, et al., 2009; Grigoryan & Lockridge, 2009; Grigoryan, Schopfer, et al., 

2009; Grigoryan, et al., 2008). Although interesting and provocative, these studies were usually 

conducted with exceedingly high concentrations (high micromolar to millimolar) of the OP compound, 
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making the connection to a “real world” human exposure tenuous. In 2010, however, Jiang and 

coworkers published an in vivo study in which they found covalently labeled tubulin in the brains of mice 

treated with chlorpyrifos or chlorpyrifos oxon, showing the real-world possibility of chlorpyrifos-induced 

covalent modification of brain tubulin in mammals at doses that produced minimal blood cholinesterase 

inhibition (Jiang, et al., 2010). Other studies have shown perturbations in tubulin dynamics, 

perturbations in the amount of tubulin protein or microtubule associated proteins, and changes in 

axonal transport elicited by treatment with chlorpyrifos or its oxon. Submicromolar concentrations of 

chlorpyrifos oxon markedly inhibited tubulin polymerization in vitro (Prendergast, et al., 2007), while in 

vitro treatment with chlorpyrifos oxon (but not chlorpyrifos) reduced that amount tubulin in rat C6 

glioma cells (Sachana, et al., 2008). Decreases in the amount of microtubule associated proteins after 

chlorpyrifos or chlorpyrifos oxon treatment has been reported by three separate studies: two in vitro

studies both showed decreases after chlorpyrifos oxon treatment (Prendergast, et al., 2007; Sachana, et 

al., 2008), and one in vivo study showed extensive reduction in specific microtubule associated proteins 

in the brains of mice treated with chlorpyrifos oxon (Jiang, et al., 2010). Depressed axonal transport has 

been noted by many laboratories both in vivo (Terry, et al., 2007; Terry, et al., 2003) and in vitro

(Gearhart, et al., 2007; Middlemore-Risher, et al., 2011) using both chlorpyrifos (in vivo and in vitro

studies) and chlorpyrifos oxon (in vitro studies). In two instances (Gearhart, et al., 2007; MiddlemoreRisher, et al., 2011), the effects of the chlorpyrifos and oxon were noted at concentrations of 

chlorpyrifos or chlorpyrifos oxon that did not inhibition cholinesterase activity (Middlemore-Risher, et 

al., 2011) or in systems that were so basic (purified microtubule/kinesin) that no cholinesterase was 

present (Gearhart, et al., 2007).

The construction of an adverse outcome pathway using chlorpyrifos-induced effects on tubulin and 

microtuble associated proteins is still in its infancy mainly because although it is thought that tubulin, 

microtubule associated proteins and axonal transport are integral to nervous system development and 

maintainence, there is no experimental evidence that perturbations of these endpoints by chlorpyrifos 

during development has neurotoxic outcomes. 

3.2.1.7 Summary/Conclusions 

As can be seen from the section above, there are several lines of evidence for actions of chlorpyrifos 

distinct from the classical mode of action of cholinesterase inhibition. This information has been 

generated from model systems representing different levels of biological organization, and provide 

support for molecular initiating events (binding to the morphogenic site of AChE, muscarinic receptors, 

or tubulin), cellular responses (alterations in neuronal proliferation, differentiation, neurite growth, or 

intracellular signaling) and responses at the level of the intact nervous system (serotonergic tone, axonal 

transport). In some cases these effects occur at chlorpyrifos exposure levels below or equivalent to 

those which result in cholinesterase inhibition in the same test system. When taken in aggregate, data 

from individual studies can be used to develop plausible hypotheses for a mode of action leading to 

developmental neurotoxicity (e.g. inhibition of the morphogenic action of AChE leading to reduced 

neurite outgrowth and subsequent miswiring or dysmorphology of the brain); however, as is the case 

for many other developmental neurotoxicants, most of these studies have not been designed with the 

specific goal of construction or testing an adverse outcome pathway. Thus, there are not sufficient data 

available to test rigorously the causal relationship between effects of chlorpyrifos at the different levels 

of biological organization in the nervous system. This is not surprising in light of the complex, intricate, 

and interrelated dynamic processes ongoing in the developing brain and multiple critical windows for 

exposure. There are a number of known developmentally neurotoxic chemicals with well-established 

relationships between exposure and neurological disorders in humans for which a definitive mode of 
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action has not been established: for example lead, methyl mercury, and ethanol (Alfonso-Loeches & 

Guerri, 2011; Castoldi et al., 2008; Farina, Rocha, & Aschner, 2011; Johansson et al., 2007; Verstraeten, 

Aimo, & Oteiza, 2008). Even today, armed with thousands of published papers on these three, 

accepted, developmental neurotoxicants, no coherent adverse outcome pathway or pathways can be 

constructed, because there are a multitude of initiating toxic events and cellular responses put forth, 

and they are not positively connected with one another. The only adverse outcome pathway available 

for any form of neurotoxicity has been developed for domoic acid in adult animals (Watanabe et al., 

2011). Domoic acid was chosen for this case study due to a large amount of preexisting data and a 

molecular initiating event at a receptor (glutamate) with known consequences to its overactivation in 

the adult nervous system.

3.2.2 Adverse Outcomes in Laboratory Animals: Effects on the Developing Brain 

There is a considerable and growing body of literature on the effects of chlorpyrifos on the developing 

brain of laboratory animals (rats and mice) indicating that gestational and/or postnatal exposure may 

cause persistent behavioral effects into adulthood. These data provide support for the susceptibility of 

the developing mammalian brain to chlorpyrifos exposure. The literature was summarized and a 

preliminary review was provided by the Agency for the 2008 SAP. At that time, the Panel agreed that 

exposure to doses of 1 mg/kg/d and greater, during some developmental period, produced significant 

and long-term effects on behavior. While behavioral changes were consistently reported, they were 

somewhat inconsistent, most likely due to experimental design differences. Such factors include route 

of exposure, developmental period of exposure, test methods, choice of dependent variable, statistical 

analyses, and so on, which are critical features of all developmental neurotoxicity studies. 

In this section the Agency reviews new literature, i.e., papers published since 2008, to build on the 

earlier SAP review. The new information provided is evaluated in terms of concordance with, or 

divergence from, the previous, pre-2008 findings. Overall, these new studies serve to strengthen the 

findings on which the 2008 SAP made their conclusions and further documents the scope of long-term 

behavioral effects resulting from chlorpyrifos exposure during development. 

Papers considered herein by the Agency as addressing long-term outcomes from developmental 

exposure include only those where chlorpyrifos is administered during the preweaning period (either 

gestational and/or postnatal) and the offspring are examined at some time after weaning. That is, 

papers reporting evaluations shortly after birth or during the pre-weaning period do not reflect longterm consequences and may also be confounded by AChE/ChE inhibition during concurrent or recent 

exposure; thus, those data are not discussed further here. In addition, the Agency focused its efforts on 

studies using relatively low doses, and that did not (or would not be expected to) produce a 

considerable degree of brain AChE inhibition and resultant cholinergic toxicity. These constraints aid in 

the unencumbered evaluation of longer-term effects compared to acute impacts of AChE inhibition, but 

it does limit the use of some papers. 

To date, 25 papers are identified and evaluated by the Agency; nine have been published from 2008 to 

the present (Abou-Donia et al., 2006; Aldridge, Levin, et al., 2005; Billauer-Haimovitch et al., 2009; 

Braquenier, Quertemont, Tirelli, & Plumier, 2010; Carr et al., 2001; Chakraborti, Farrar, & Pope, 1993; 

Chanda & Pope, 1996; Dam, Seidler, & Slotkin, 2000; Haviland, Butz, & Porter, 2010; Icenogle et al., 

2004; D. A. Jett, Navoa, Beckles, & McLemore, 2001; Johnson, Chambers, Nail, Givaruangsawat, & Carr, 

2009; Laviola, Adriani, Gaudino, Marino, & Keller, 2006; Levin et al., 2002; Levin et al., 2001; Maurissen, 

et al., 2000; Muto, Lobelle, Bidanset, & Wurpel, 1992; Ricceri et al., 2003; Ricceri et al., 2006; Venerosi, 
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Calamandrei, & Ricceri, 2006; Venerosi et al., 2008; Venerosi, et al., 2010; Venerosi, Ricceri, Scattoni, & 

Calamandrei, 2009; Zhang et al., 2011). It is important to appreciate the various experimental designs, 

encompassing species, gender, exposure scenarios, ages of assessment, and test apparatuses. Of these 

25 papers, four papers and one study from a fifth paper did not meet the criteria stated above, and have 

been excluded from further consideration herein. 

Specifically, these excluded studies are: 

1) Muto et al. (1992) used a chlorpyrifos product that was formulated in xylene, a known 

developmental and reproductive toxicant, yet the vehicle control group received saline, not 

xylene. Thus, the findings of that study cannot be attributed to chlorpyrifos alone. 

Furthermore, all testing took place before weaning, no later than PND16;

2) A gestational-exposure study (GD12-19) used a dose (25 mg/kg/d sc) that produced 90% brain 

inhibition in the dams at the end of dosing, and ~20% brain inhibition in offspring on PND3 

(Chanda & Pope, 1996). The reported motor effects are confounded with the concurrent 

cholinesterase inhibition on PND1 and 3, and no testing was conducted at later ages;

3) A postnatal study (Chakraborti, et al., 1993) used a very high dose (40 mg/kg/d sc) given directly 

to the pups postnatally (PND7, 11, 15, 19), with up to 60% brain inhibition evident four days 

after dosing;

4) A gestational study (Venerosi, et al., 2009) only tested pups up to PND12, reporting altered 

motor ontogeny and changes in ultrasonic vocalizations; and

5) The “postnatal” study described by (D. A. Jett, et al., 2001) involved dosing after weaning, while 

the rats were being tested in the Morris water maze. Note, part of this paper remains in 

consideration, since the “preweaning” study from that paper met the criteria for inclusion. 

3.2.2.1 Summary of Existing Database

Of the 21 papers in the Agency’s review, several trends are apparent. The numbers of rat and mouse 

studies are similar (11 rat, 10 mice). As seen in Table 5 below, this research now spans over a decade, 

with the first publications in 2000. These studies have been generated in ten different laboratories, four 

of which have been added since SAP literature review in 2008. Of the nine studies published in 2008 

and later, seven evaluated mice and the remaining two used rats; however, one of the mouse studies 

was excluded from further discussion since there were no assessments conducted after weaning 

(Venerosi, et al., 2009). Most, but not all, of the studies include testing in both males and females. Two 

laboratories have generated the most research: Istituto Superiore di Sanita (ISS; Rome, six studies), and 

Duke University Medical Center (NC, USA, six studies). The majority of studies (14) used subcutaneous 

administration, eight used oral gavage, one employed an osmotic minipump, and one used a dermal 

route. Of the two major contributing laboratories, Duke University always uses subcutaneous 

administration of chlorpyrifos in DMSO vehicle, whereas the ISS laboratory uses oral dosing of the dams 

with peanut oil vehicle during gestation, and subcutaneous dosing of the pups postnatally with either

peanut oil or DMSO as vehicle. Eight studies involved only gestational exposures, nine used only direct 

postnatal exposures to the pups, and two spanned gestational and lactational (LD) periods. Note that 

two studies actually included both gestational and direct postnatal exposures in the same subjects, and 
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this more complicated experimental design resulted in all possible combinations of doses both pre- and 

postnatally. For the discussion below, only treatments that included chlorpyrifos at one or the other 

dosing period, but not both, are included (i.e., vehicle followed by chlorpyrifos, or chlorpyrifos followed 

by vehicle); however, these data are listed in Appendix 3. All studies used chlorpyrifos except one 

(Laviola, et al., 2006) who used the oxon metabolite. One study is in Chinese, with only the abstract 

available in English, and the route is not specified. Only two of the new papers include any measures of 

AChE inhibition, whereas a few others refer to earlier papers from the laboratory with such 

measurements. 

Table 5 Summary of Developmental Chlorpyrifos Studies by Laboratory

Laboratory Principal 

Investigator(s) Number of Studies Year(s) of publication

Duke University, NC, USA Slotkin, Levin; AbouDonia 6 2000, 2001, 2002, 

2004, 2005, 2006

Mississippi State University, 

MS, USA Carr 2 2001, 2009

Dow, MI, USA Maurissen 1 2000

Johns Hopkins University, 

MD, USA Jett 1 2001

University of Wisconsin 

Madison, WI, USA Haviland 1 2010

Istituto Superiore di Sanita, 

Rome, Italy Calamandrei; Laviola 6 2003, 2006, 2006, 

2006, 2008, 2010

Hebrew University, 

Jerusalem, Israel Yanai 2 2011

University of Liége, Belgium Plumier 1 2010

Central South University, 

Changsha, China Zhao 1 2011

In addition to differences in species, gender, and route and timing of exposure, evaluation of the existing 

database shows a lack of consistency for the specific methods, equipment, variables, and analyses of the 

various tests. For example, locomotor activity has been measured in various chambers of differing size 

and configuration, data are collected differently (e.g. automated or visual scoring), and measures may 

vary (e.g., latency for a response or area of exploration). While conclusions are based on statistical tests 

of significance, differences across the papers range from restricting the alpha level (e.g., p<0.02) to 

prevent type I errors to accepting interaction terms of higher probabilities (e.g., p<0.10) to continue 

with post-hoc determinations. Despite these differences, the discussion below is based on the authors’ 

conclusions and interpretations. While this situation is not unusual in the developmental neurotoxicity 

literature, it does render direct comparisons across studies more difficult (Cory-Slechta et al., 2001). For 

this reason, we and others (e.g., (Prueitt, Goodman, Bailey, & Rhomberg, 2011) have focused more on 

evaluating outcomes in terms of overall domains of neurological function rather than the specific 

endpoints used to measure that function. 
Page 39 of 124

3.2.2.2 Developmental Impacts on Neurological Domains

Because many of these papers report a number of positive as well as negative findings, the Agency had 

previously taken the approach of comparing responses that were observed following various exposures 

to a common dose, 1 mg/kg/d ((FIFRA Scientific Advisory Panel (SAP), 2008a; U.S. Environmental 

Protection Agency, 2011). A more robust approach is taken here, to include important factors such as 

dose-response and differences in exposure scenarios. In order to evaluate effects on domains of 

neurological functions, broken down by exposure period, the papers have been summarized as such in 

Appendix 3. It is well-known that exposures during different critical periods of development can result 

in very different outcomes (Adams et al., 2000; Rice & Barone, 2000). Unfortunately, development of 

the nervous system does not occur in distinct non-overlapping temporal periods with which to attribute 

the apical behavioral changes observed. In the text below, the more recent papers (2008 to present) are 

described and evaluated in terms of their contribution to and concordance with, previous papers using 

similar measures. The Agency’s evaluation follows several Bradford Hill criteria including consistency of 

findings across domains and dose response. The latter was considered as either graded magnitude of 

effects across doses or significant findings at a higher but not lower dose; unfortunately, many of the 

chlorpyrifos studies have evaluated only one dose. 

The new studies include measures that evaluate the following neurobehavioral domains: cognitive 

function, anxiety and emotion, social behavior and maternal interactions, and motor activity. These are 

discussed in greater detail below; overall, the newer data were in general agreement with previous 

studies. Although a few earlier studies addressed neuromotor development (Abou-Donia, et al., 2006; 

Maurissen, et al., 2000), this domain has not been evaluated more recently; with no new information, 

earlier reviews have not changed. 

3.2.2.2.1 Cognitive Function

A few of the new studies evaluated effects on cognitive function, using some of the same procedures 

used in the earlier studies (radial arm maze, Morris water maze). The earlier findings are summarized 

below, along with comparisons to the newer studies. 

Radial Arm Maze

The radial arm maze is a spatial task requiring the subject to enter different arms for food reinforcer, 

which is located in only certain arms. A widely used test of learning and memory, both working errors 

(re-entering baited arms after taking the food) and reference errors (entering arms that are never 

baited) are recorded over days of training. From 2001-2005, a series of behavioral studies from a single 

laboratory (Duke University) described effects on radial arm maze performance in adult rats that were 

treated with chlorpyrifos during gestational or postnatal development. Both working and reference 

memory are altered in rats treated during gestation or early postnatally, and gender differences were 

described in all but one study. Across these studies, significant differences were detected in one or the 

other type of error, or both, and error selectivity was not fully consistent across studies. Following early 

gestational exposure (GD9-12), high-dose (5 mg/kg/d) rats of both sexes showed increased working and 

reference memory errors early in training (Icenogle, et al., 2004). With later gestational exposure 

(GD17-20), only females showed more errors; however, this was observed only at the low dose (1 

mg/kg/d, not 5 mg/kg/d) (Levin, et al., 2002). With early postnatal exposure (PND1-4, 1 mg/kg/d), two 

separate studies showed similar results in that error rate was increased in males (worse performance) 

but decreased in females (better performance) (Aldridge, Levin, et al., 2005; Levin, et al., 2001). Only 
Page 40 of 124

the late postnatal exposure (PND11-14, 5 mg/kg/d) was without effect on this measure (Levin, et al., 

2001). 

Several new studies have confirmed chlorpyrifos impacts on learning and memory. In a new study using 

the radial arm maze and longer postnatal exposures (PND1-21), male, but not female, rats showed more 

working memory errors at all doses (lowest dose, 1 mg/kg/d) (Johnson, et al., 2009). This was evident 

later in training, indicating that the task was not learned as well as in controls. Reference memory 

errors showed the same gender-specific pattern as seen in two Duke studies (Aldridge, Levin, et al., 

2005; Levin, et al., 2001), in that chlorpyrifos-treated females showed fewer errors during training, and 

males showed more. The altered reference memory pattern was only seen in the mid and high dose 

groups, which had incrementally increasing doses (up to 4 or 6 mg/kg/d) during the exposure period. In 

this study, the low-dose group had experienced about 14% brain AChE inhibition on PND20, with no 

inhibition evident when radial arm maze training began (PND36). On the other hand, the mid and highdose rats showed some residual AChE inhibition at that time, making it challenging to separate the 

memory deficits from concurrent AChE holinesterase inhibition.

A study using mice exposed gestationally (GD17-20, 1 or 5 mg/kg/d) also used a radial arm maze, 

although comparisons are difficult due to differences in length of training and data presentation 

(Haviland, et al., 2010). Reference memory errors showed some differences across trial, dose, and sex; 

these are not interpreted by the authors as a meaningful effect, and may indeed be spurious. The 

authors used these data to compare with their novel foraging task, which used a modification of a radial 

arm maze and examined the rate of learning that a reward (food) is present (recognition) as well as its 

location (positional learning). Both of these parameters were altered by chlorpyrifos exposure. The low 

dose (1 mg/kg/d) females were delayed in learning to recognize the reward, but there were no 

statistically significant differences at that dose in positional learning (despite the authors’ claim of an 

effect). On the other hand, low-dose males showed accelerated food recognition, and increased 

positional learning that was evident during only two sessions. The high-dose groups (both sexes) 

showed the same pattern of changes, with somewhat greater magnitude of differences, indicating doseresponse. Thus, effects were observed on a spatial learning task, albeit a different apparatus and 

procedure. Note that the gender effect was reversed from that reported above for rats in the radial arm 

maze – these differences could be due to species or testing apparatus.

Morris Water Maze

The Morris water maze is a different type of spatial learning task, which can be varied to assess different 

types of learning and memory. In this test, the animals are trained over days to swim to the location of 

a submerged platform to escape from the water, and learning is evident by faster latencies and other 

measures of memory for the platform position. A recent paper using gestational exposure (GD9-18) in 

mice resulted in slower learning in the offspring, although a clear dose-response was not evident 

(Billauer-Haimovitch, et al., 2009). The authors report an overall effect of chlorpyrifos treatment, but 

state that only the lower doses (1, 3 mg/kg/d, but not 5, 10 mg/kg/d) were individually significant. 

While the magnitude of effect does not appear to be pronounced in the first study (visual inspection of 

Figure 3 in the paper), the finding was repeated, and was more obvious, in two additional studies using a 

single dose (3 mg/kg/d). Another recent paper from the same laboratory again reported slower learning 

in mice exposed during gestation (3 mg/kg/d; GD9-18) (Turgeman et al., 2011). Thus, there are internal 

replications of this finding across several studies, albeit in the same laboratory.
Page 41 of 124

These recent findings, plus another study available only in Chinese17, extend an earlier report of rats 

that showed deficits in the Morris water maze (beginning on PND23) following exposure on PND7, 11, 

and 15 (0.3, 7 mg/kg/d ) (D. A. Jett, et al., 2001). The 2008 SAP, however, was critical of this paper, for 

reasons described below.

Summary

Taken together, these studies in rats and mice show altered cognitive function using well-accepted tests 

of spatial learning and memory (radial arm maze, Morris water maze). The direction of change may be 

sex-specific and dependent on the timing of exposure. Often these changes suggest impaired learning 

and/or memory. While enhanced function is also apparent in some studies, such changes are evidence 

of alterations in memory processes nonetheless. Several of these findings have recently been replicated 

across studies and laboratories. Effects were also reported in a spatial foraging task (Haviland, et al., 

2010), but direct comparison between that and the radial arm maze is difficult. Earlier papers have 

reported that other cognitive tasks (spontaneous or delayed alternation, passive avoidance) are not 

altered, but there are no new studies using these other tasks. These outcomes are summarized in Table 

6. In this and the following tables, details including dose, route of exposure and age at testing are not 

included but are provided in Appendix 3.

 17 Another new study also evaluated Morris water maze learning, using rats exposed postnatally 

(PND11-14, 5 mg/kg/d, route uncertain) (Zhang et al., 2011). The abstract states that learning and 

memory impairments were observed, but the rest of the paper cannot be critically evaluated at this 

time (article is in Chinese) and so cannot be combined with other findings. 
Page 42 of 124

Table 6 Summary of Outcomes on Cognitive Tests in Male (M) and Female (F) Rodents

Early gestation

GD9-12; 9-18

Late gestation

GD 17-20

Perinatal

GD6-LD10

Early postnatal

PND1-4

Late postnatal

PND11-14; 7,11,15

Postnatal

PND1-21

Radial Arm Maze

Cognitive deficit -

rat, M&F, doseresponse6

Cognitive deficit -

rat, F not M, no 

dose-response4

Cognitive deficit in 

M, improved 

function in F - rat3

Cognitive deficit in 

M, improved 

function in F - rat7

No effect – rat, 

M&F3

Cognitive deficit in 

M, improved 

function in F - rat, 

dose-response9

Morris Water Maze

Cognitive deficit -

mouse, M&F, no 

dose-response8

Cognitive deficit –

mouse, M&F11

Cognitive deficit -

rat, M&F, doseresponse2

Foraging Maze

Cognitive deficit in

F, improved 

function in M –

mouse, doseresponse in F not 

M10

T-maze 

Spontaneous 

Alternation

No effect – rat, 

M&F6

No effect – rat, 

M&F4

No effect – rat, 

M&F3

No effect – rat, 

M&F3

Delayed Spatial 

Alternation

No effect – rat,

M&F1

Passive Avoidance No effect – mouse, 

only M tested5

No effect – mouse, 

only M tested5

1

 Maurissen et al., 2000 2

 Jett et al., 2001 3

 Levin et al., 2001 4

 Levin et al., 2002 5

 Ricceri et al., 2003 6

 Icenogle et al., 2004 7

 Aldridge et al., 2005

8

 Billauer-Haimovitch et al., 2009 9

 Johnson et al., 2009 10 Haviland et al., 2010 11 Turgeman et al., 2011
Page 43 of 124

3.2.2.2.2 Anxiety/Emotion

Anxiety and despair/affect have been modeled in animals using a number of different tests, including 

those that measure avoidance of aversive areas, investigation of novel areas, response to forced activity, 

and response to preferred substances. Several new studies assessed anxiety, despair, and affect. While 

different procedures were used in most cases, there was some concordance in outcomes. These are 

described below. 

Tests of Anxiety

Several studies have suggested that chlorpyrifos alters measures of anxiety. The early studies came 

from the laboratories of Duke University and ISS (Italy). Three newer studies, all in mice, add to the 

findings in this area: two were from the ISS laboratory and one came from a third laboratory (Belgium). 

Earlier studies employed an elevated plus maze, which has open and enclosed areas: being fearful of 

bright open areas, rodents tend to stay in the enclosed arms. Decreased anxiety is inferred from 

changes such as increased time in the open arms, decreased head dipping, and other measures. Using 

this paradigm, lower anxiety was reported in rats (males, not females) exposed postnatally to 1 mg/kg/d 

(PND1-4)(Aldridge, Levin, et al., 2005; Ricceri, et al., 2003) and in mice (females, not males) exposure 

postnatally to 3 mg/kg/d (PND11-14) (Ricceri, et al., 2006). On the other hand, increased anxiety was 

measured in mice exposed gestationally (GD15-18), but only in the low-dose (3 mg/kg/d, not 6 mg/kg/d) 

males (not females) (Ricceri, et al., 2006). There were no changes in anxiety behaviors in rats exposed 

gestationally (GD9-12, 1 or 5 mg/kg/d) (Icenogle, et al., 2004).

Three new studies have used a different apparatus, the light-dark box, to measure anxiety, but the 

principle is the same in that rodents typically prefer the dark chamber. In one study (Braquenier, et al., 

2010), both the elevated plus maze and the light-dark box were used, providing direct comparisons 

between these tests. Mice were exposed both gestationally and lactationally (GD15-LD14, 0, 0.2, 1, or 5 

mg/kg/d), and only female offspring were tested. When tested in the light-dark box, the middle dose 

group only (1 mg/kg/d) showed decreased time in the center of the light side, which was considered to 

reflect increased anxiety. The same dose group moved back and forth between the sides less often, 

which could reflect general activity levels but was also interpreted by the authors as greater anxiety. 

The time in the dark and light sides, however, did not differ. Littermates were tested in the elevated 

plus maze, and again the middle dose group showed less time spent in the open arms and fewer open 

arm entries, supporting a conclusion of increased anxiety, but only at the middle dose. The lack of doseresponse was not addressed by the authors.

A study in which mice were treated gestationally (GD14-17, 6 mg/kg/d) (Venerosi, et al., 2010) showed 

no difference in time spent on either side of the light-dark box, but females (not males) spent more time 

in the tunnel connecting the sides. This finding, along with a few other measures that did not reach 

statistical significance, was interpreted by the authors as increased anxiety. Another study by this group 

(Venerosi, et al., 2008) exposed mice postnatally (PND11-14, 3 mg/kg/d), then bred the female offspring 

as adults and allowed them to deliver normally. On postpartum day 2, the dams were removed from 

the pups and tested in the light-dark box. They reported a decreased latency and higher proportion of 

mice entering the light side, indicating decreased anxiety. Time in the dark or light sides was not 

reported.
Page 44 of 124

Despair/Affect

Only a few studies have also included measures of despair and affect, and the results are not completely 

consistent. Most recently, behavioral despair was measured using a forced swimming procedure 

(Venerosi, et al., 2010). In that study, gestational exposure to chlorpyrifos did not alter any baseline 

responses, and there are no other studies using similar measures with which to compare. There have 

been no new studies that could add to, and aid in the interpretation of, the findings reported by others: 

decreased preference for chocolate milk in rats, or no effect on novelty exploration in mice (Aldridge, 

Levin, et al., 2005; Ricceri, et al., 2003). 

Summary

Taken together, these assessments suggest that, in both rats and mice, changes in anxiety are 

dependent on exposure period. Specifically, postnatal exposure decreases anxiety, whereas increased 

anxiety is observed following late gestational exposure or a longer gestational plus postnatal exposure. 

Some inconsistencies are evident, such as lack of dose-response in a few studies and effects on one sex 

or the other (same laboratory, similar dosing regimen). Thus, while the data are not fully consistent, 

overall there is evidence for long-term changes in anxiety behavior following chlorpyrifos exposure, as 

shown in Table 7.
Page 45 of 124

Table 7 Summary of Anxiety/Emotion Outcomes in Male (M) and Female (F) Rodents

Early gestation

GD9-12

Late gestation

GD14-17; GD15-18

Early postnatal

PND1-4

Late postnatal

PND11-14

Perinatal

GD15-LD14

Elevated Plus Maze 

(anxiety) No effect – rat, M&F1

Decreased anxiety -

mouse, M not F, no 

dose-response5

Decreased anxiety - rat, 

M not F2

Decreased anxiety -

mouse, F not M, doseresponse5

Increased anxiety -

mouse, only F tested, 

no dose-response3

Light-Dark box (anxiety) Increased anxiety -

mouse, F not M7

Decreased anxiety -

mouse, dams6

Increased anxiety -

mouse, only F tested, 

no dose-response3

Despair No effect – mouse, 

M&F7

Novelty/preference

Decreased anxiety - rat, 

M&F2

No effect – mouse, 

M&F4

No effect – mouse, 

M&F4

1 Icenogle et al., 2004

2 Aldridge et al., 2005

3 Braquenier et al., 2010

4 Ricceri et al., 2003

5 Ricceri et al., 2006

6 Venerosi et al., 2008

7 Venerosi et al., 2010
Page 46 of 124

3.2.2.2.3 Social Behaviors/Interactions

Conspecific behaviors are not typically studied in the context of developmental neurotoxicity studies, 

and methods for such assessments are not well-standardized (Cory-Slechta, et al., 2001). Aggressive, 

social, and parental behaviors have been studied following developmental exposure to chlorpyrifos, all 

of which were conducted in the ISS laboratory using mice. Even within this one laboratory, however, 

various testing methods have been used. These include: 1) social investigation of nulliparous 

female:female pairs; 2) social investigation and agonistic behaviors in male:male pairs; 3) maternal 

behaviors towards pups, induced by placing nulliparous females with foster litters; 4) maternal 

behaviors in lactating dams towards their own pups, and 5) agonistic behaviors in lactating dam:male 

pairs. In most studies, measures include social (interactive) as well as nonsocial behaviors (e.g., 

grooming, exploration). Within each grouping, there are often many measures taken, and treatment 

effects have sometimes been reported on just a few, suggesting that the changes are subtle. There are 

few systematic comparisons of chlorpyrifos effects across these varied behaviors using similar dosing 

regimens. 

Maternal Behavior

Two recent papers focused on maternal behavior, measuring actions of the lactating dam towards her 

pups, nesting activity, and agonistic behaviors towards an intruder male. In one study (Venerosi, et al., 

2008), mice were exposed postnatally (PND11-14, 3 mg/kg/d), mated as adults, and the dams were 

tested. They reported that the treated dams showed delayed start of nesting, decreased latency to lick 

pups, as well as fewer defensive postures and more social investigation of an intruder male. This latter 

effect was replicated in that decreased aggressive behaviors were also reported for lactating dams that 

had been exposed in their own fetal period (GD14-17, 6 mg/kg/d) (Venerosi, et al., 2010). These 

findings were interpreted by the authors as impaired maternalistic behaviors.

Social/Agonistic Behavior

Evaluations of male or female same-sex social behaviors generally show no effect of postnatal (PND1-4 

or PND11-14) chlorpyrifos exposures (Ricceri, et al., 2003; Venerosi, et al., 2006), a finding that was 

recently confirmed (Venerosi, et al., 2008; Venerosi, et al., 2010). On the other hand, increased 

investigation of the stranger mouse was reported following gestation exposure (GD15-18)(Venerosi, et 

al., 2006), and increased solicitation behaviors were observed following late postnatal exposure (PND11-

14) (Ricceri, et al., 2003). Thus, there are contradictory reports of changes in female social behavior, 

being either increased or not altered. The only study of male social behavior showed no effect.

Earlier studies reported that male agonistic behaviors were consistently increased in mice exposed 

either postnatally (PND1-4, PND11-14) or gestationally (GD15-18); the lowest effective dose was 1 

mg/kg/d administered during either postnatal period (Ricceri, et al., 2003; Ricceri, et al., 2006). None of 

the newer studies addressed male behaviors. These findings are summarized below (Table 8). 

Summary

Overall, decreased aggressive behaviors of dams that had been exposed during their development are a 

common finding across several exposure periods, as are increased male agonistic behaviors. Social 

and/or maternal behaviors in females are less consistently or convincingly altered. 
Page 47 of 124

Table 8 Summary of Social/Interactive Behavior Outcomes in Male (M) and Female (F) Mice

Exposure/effect Late gestation

GD14-17, 15-18

Early postnatal

PND1-4

Late postnatal

PND11-14

Female:female social Increased investigation, doseresponse3

No effect1 Increased solicitation, dose-response1

No effect3

No effect4

Male:male social Increased aggressive postures, doseresponse2

Increased agonistic behavior, differs 

across time-course, dose-response1 

Increased solicitation, dose-response1

Increased agonistic behavior, over 

time, no dose-response1

Increased attack behaviors, doseresponse2

No effect social4,a

Induced maternal No effect2 Increased maternalism, doseresponse2

Natural maternal Decreased maternalism4

Natural maternal:male Decreased aggression5 Decreased defensiveness4

DR dose-response; if not mentioned, only one dose tested a

 apparently only social, no agonistic, behaviors measured 1 Ricceri et al., 2003

2 Ricceri et al., 2006

3 Venerosi et al., 2006

4 Venerosi et al., 2008

5 Venerosi et al., 2010
Page 48 of 124

3.2.2.2.4 Motor activity

Most of the developmental studies of chlorpyrifos have evaluated motor activity of some sort; these 

were presented and summarized in detail in the 2008 SAP report. In general, activity levels have been 

recorded in terms of: 1) locomotion; 2) response latencies or other activity measures in the course of 

testing in various apparatuses (t-maze, radial arm maze, etc); and 3) habituation of activity over the 

session in activity chambers. In general, the earlier studies reported that activity is increased, 

decreased, or not altered, in either both or just one sex. Recent studies add to this literature, but none 

provide information to better explain these varied and contradictory outcomes. 

Activity Devices

Evaluations of locomotor activity in open fields or other activity devices have been made in numerous 

studies across quite a few laboratories. Looking across developmental period and gender, there have 

been similar reports of increased, decreased, or no change in various measures of activity (e.g., 

exploration, rearing, etc). Generalizations of effect could not be made in the 2008 preliminary review. 

Two recent studies have used an open field to measure activity. Braquenier et al. (2010) reported no 

change of activity levels in female mice that had been exposed from late gestation through late lactation 

(GD15-LD14). While Zhang et al. (Zhang, et al., 2011) reported decreased activity in rats, those data 

cannot be evaluated at this time due to the language barrier. 

Ancillary Activity Measures

As with activity measured specifically in activity devices, various measures of activity during other tests 

have shown increases, decreases, or no effect, in both or only one sex, with no obvious association to 

exposure period. It is important to consider, for example, that the time it takes to visit the arms of a 

radial arm maze is a somewhat different behavior than exploratory movement in a novel chamber. 

Thus, these activity measures may not be fully comparable, and the evaluations described here should 

be interpreted with caution.

Three recent papers reported assessments of activity during cognitive and social behavior testing. There 

was no change noted in postnatally exposed rats during radial arm maze testing (PND1-21) (Johnson, et 

al., 2009), or in gestationally exposed mice during foraging testing (GD17-20) (Haviland, et al., 2010). 

Mice exposed postnatally (PND11-14) showed decreased exploratory behaviors during the 

acclimatization phase of social testing (i.e., before being exposed to another mouse); however, this was 

only significant in the first block of the test session (Venerosi, et al., 2008). 

Habituation

Habituation, or decrease in activity level during the course of a test session, has only been specifically 

evaluated in the a few laboratories. As with other activity outcomes, the data are specific to exposure 

period and gender. Within the same laboratory, habituation is either faster (early gestation, both sexes) 

(Icenogle, et al., 2004), slower (late gestation, females only; late postnatal, both sexes) (Levin, et al., 

2002; Levin, et al., 2001), or not altered (early postnatal) (Levin, et al., 2001). A perinatal study (GD6-

LD10) also reported no change in habituation; however, the data were not statistically analyzed in the 
Page 49 of 124

same way, so results are not comparable (Maurissen et al., 2000). This measure has not been addressed 

in more recent studies.

Summary

As seen in Table 9, there remains to be inconsistencies in the various measures of motor activity, but it is 

important to note the numerous differences in procedures and apparatuses in which activity was 

measured. Given this, generalizable conclusions and summaries cannot be made at this time.
Page 50 of 124

Table 9 Summary of Motor Activity Outcomes in Male (M) and Female (F) Rodents

Early gestation

GD9-12

Late gestation

GD14-16; 14-17; 

15-18; 17-20

Perinatal

GD6-LD10; GD15-

LD14

Early postnatal 

PND 1-4

Late postnatal 

PND11-14 Postnatal PND1-21

Activity 

chambers No effect – rat, M&F6

Increased activity -

mouse, only M 

tested, doseresponse13

No effect - mouse8

No effect – rat, 

M&F10

No effect – rat, 

M&F11

No effect – mouse, 

only F tested2

No effect - mouse, 

M&F12

Decreased activity -

rat, M not F4

No effect – rat, M&F9

Increased activity -

mouse, only M 

tested, doseresponse13

Increased activity -

mouse, M&F, doseresponse12

Increased activity -

rat, M&F4

No effect – rat, M&F9

Decreased activity -

rat, M&F, doseresponse3

Activity 

measures

Increased activity -

rat, M&F, doseresponse6

No effect - mouse, 

only F tested14

No effect - mouse, 

M&F5

Increased activity -

rat, M&F, DR10

No effect - mouse, 

M&F13

No effect – mouse, 

only F tested2

Increased activity -

rat, M not F1

Increased activity -

mouse, M&F12

Decreased activity -

rat, M not F9

Decreased activity -

mouse, M&F14

Decreased activity -

rat, M not F9

Increased activity -

mouse, M&F12

No effect – mouse, 

M&F13

No effect – rat, M&F7

Habituation 

rate 

Increased habituation

- rat, M&F, doseresponse6

Decreased 

habituation - rat, F 

not M, doseresponse10

No effect – mouse8

No effect – rat, 

M&F11 No effect – rat, M&F9

Decreased 

habituation - rat, 

M&F9

1

 Aldridge et al., 2005 2

 Braquenier et al., 2010 3

 Carr et al., 2001 4

 Dam et al., 2000 5

 Haviland et al., 2010 6

 Icenogle et al., 2004 7

 Johnson et al., 2009 8

 Laviola et al., 2006

9

 Levin et al., 2001 10 Levin et al., 2002 11 Maurissen et al., 2000 12 Ricceri et al., 2003 13 Ricceri et al., 2006 14 Venerosi et al., 2008
Page 51 of 124

3.2.2.3 Dose-Response Considerations

One of the key issues for the Agency is the degree to which points of departure based on cholinesterase 

inhibition are protective for the neurobehavioral outcomes reviewed in this chapter following 

developmental exposure. Few of these papers assess AChE inhibition at all, much less at time points 

potentially critical to the ultimate outcomes, e.g., within hours after dosing or after the last of several 

repeated doses. The Agency and the 2008 SAP agreed that doses ≥1 mg/kg/d produce significant 

neurobehavioral changes following gestational and/or postnatal exposure(s) (FIFRA Scientific Advisory 

Panel (SAP), 2008a; U.S. Environmental Protection Agency, 2011). A dose of 1 mg/kg/d has been often 

used across studies and laboratories. Half of the papers published from 2008 on also used this dose, 

reporting significant effects on several different measures; this confirms the earlier conclusions. 

A few studies have reported AChE inhibition when 1 mg/kg/d is administered directly to the pup 

postnatally (Dam, et al., 2000; Johnson, et al., 2009; Ricceri, et al., 2003). However, none of the 

neurobehavioral studies described here tested for fetal AChE inhibition when 1 mg/kg/d is given during 

gestation. A companion study to Maurissen et al. (2000) reported no cholinesterase inhibition in fetuses 

taken 4 hr after dosing to the dam when 1 mg/kg/d had been administered daily since GD6 (Mattsson, 

et al., 2000). Qiao et al. (2002) also reported no brain AChE inhibition in fetuses 24 hr after the last dose 

of 1 mg/kg/d on GD17-20. No other time points or days were assessed in either study. This suggests, 

but does not confirm, that the fetus would not experience AChE inhibition at 1 mg/kg/d, further 

suggesting that the behavioral effects reported in those studies were not due to AChE inhibition.

At the time of the 2008 SAP, there were only two studies in the literature (both in rats) that included a 

dose less than 1 mg/kg/d (D. A. Jett, et al., 2001; Maurissen, et al., 2000); a third study (in mice) has 

since been added (Braquenier, et al., 2010). In Maurissen et al. 2000, exposures of 0.3 mg/kg/d from 

GD6-LD10 produced no significant changes in the offspring across the numerous measurements (motor 

activity, startle response, delayed spatial alternation), whereas the highest dose (5 mg/kg/d) produced 

delayed growth and maturation. Female adults at both 1 and 5 mg/kg/d dose presented decreased 

parietal cortex thickness, yet this measure was not evaluated in the lowest dose group; therefore, 

conclusions about low-dose effects on that outcome cannot be made. 

A different dosing paradigm was used by Jett and colleagues (2001), who dosed pups directly with 0.3 or 

7 mg/k/g/d on PND7, 11, and 15. Rats were tested in the Morris water maze, starting on PND24, and 

the data suggested an effect in the 0.3 mg/kg/d dose group, in that latencies to find the platform were 

longer on the first day of training. With continued training, they were no longer different from control. 

While the authors state in the results that these latencies were “significantly greater” with p=0.05, the 

subsequent discussion describes this as a trend and that the group was “not significantly different” from 

control; thus, even the authors were not consistent in their conclusion. Figure 2 of Jett et al. (2001), 

however, shows that the means and standard errors for these groups do not overlap, suggesting that 

this may be a transient but real effect. The 2008 Panel was critical of this study with regards to the 

postweaning dosing regimen (not discussed here; see Section 3.2.2 for explanation) and the poor 

asymptotic performance of controls. It is critical to note, however, that these rats received a total of 

only six training trials, whereas at least twice that many trials are typically used for control animals to 

reach asymptotic performance. While the authors measured AChE inhibition, there were no changes 

even at the higher dose, 7 mg/kg/d. These findings are not in agreement with the 30% brain inhibition 

measured after a lower dose (5 mg/kg/d sc) on PND11 (Dam, et al., 2000), or with the number of studies 
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reporting brain inhibition at lower doses following acute oral dosing. Thus, uncertainty in the data for 

the high dose raises question about the AChE data altogether.

Braquenier et al. (2010) dosed mice at 0.2, 1, or 5 mg/kg/d from GD15-PND14. The only significant 

findings were increased anxiety in the 1 mg/kg/d dose group only. This lack of dose-response was seen 

using two different apparatuses, and testing different mice (littermates). The authors do not provide a 

discussion or explanation for this aberrant finding. The authors report ~14% brain AChE inhibition in the 

high dose group (only measured on PND1), and imply that the lower doses (not measured) would not 

produce any brain inhibition. Given uncertainty of no effects at 0.3 mg/kg/d (in rat), these new data 

suggest that 0.2 mg/kg/d, administered to mice for several weeks covering late gestation and early 

postnatal development, is a no-effect level. Notably, this no-effect level is similar to the acute point of 

departure and is approximately 10-fold higher than the repeated dosing point of departure. 

3.2.2.4 Conclusions

These studies report a range of neurobehavioral changes in rats and mice following developmental 

exposure to chlorpyrifos. In this ongoing area of research, more studies are being generated in even 

more laboratories, all of which report some form of neurobehavioral alteration. Obvious species 

differences have not emerged, and effective doses are similar (1-6 mg/kg/d). Changes in various aspects 

of cognitive tests indicate perturbations of learning and/or memory, even though in some cases these 

may be manifest as improved function. Likewise, alterations in domains such as anxiety and social 

interactions may differ in direction of change, but are still suggestive of impacts on normal neuronal 

processing. There is replication of some effects across studies, and with some of the newer papers, 

across laboratories as well. Activity measures, on the other hand, still provide results as varied as the 

different measures of assessment. Taken together, these data do not provide evidence for a specific 

profile of effects but instead suggest more global alterations in neurobehavioral function. 

All testing reported herein was conducted after weaning, and there is a presumption that the effects are 

permanent; however, no study has directly addressed this issue, and there is a range in test ages. Doseresponse is not always evident, since many studies only use one dose, and of those using two or more 

doses, there is not always a monotonic response. Furthermore, the summary presented herein 

combines studies of different dosing regimens. While there are demonstrated differences in uptake and 

persistence of chlorpyrifos given subcutaneously vs. oral, with different oils or DMSO as vehicle, the 

developmental literature does not provide obvious differences in outcome based on this. Likewise, the 

literature has not shown that any specific developmental period is critical overall to the long-term 

outcomes, since similar effects are shown with different exposure periods. For example, cognitive 

changes in the radial arm maze were observed following gestational and early postnatal (PND1-4), but 

not late (PND11-14), exposure (Aldridge, Levin, et al., 2005; Icenogle, et al., 2004; Levin, et al., 2002; 

Levin, et al., 2001). However, cognitive deficits were reported with the Morris water maze following 

both gestational and late postnatal exposures (Billauer-Haimovitch, et al., 2009; D. A. Jett, et al., 2001; 

Turgeman, et al., 2011). Likewise, some changes in anxiety and social behaviors were reported at both 

gestational and postnatal exposure periods. Overall, these data do not clearly show specific critical 

periods of exposure, or definitive sensitive behavioral outcomes. Unfortunately, no laboratory has 

provided systematic comparisons across exposure period, dosing regimen, and age of testing; such 

studies would improve understanding of the impact of these critical factors. 

These studies have almost exclusively focused on doses that could produce some degree, however 

minimal, of AChE inhibition. Thus it is not possible to know whether effects would be present at lower 
Page 53 of 124

doses, since they have not been adequately studied; thus far, only one study (Braquenier, et al., 2010)

has tested a dose lower than the point of departure. The broad profile of neurological effects that have 

been reported do not aid in the development of a specific AOP, and as described in section 3.2.1., 

existing experimental studies have not been designed to examine and track possible mechanisms from 

early initiating events to the final neurological outcome. Such studies represent longer term research 

efforts by the different laboratories.
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4.0 Review of Chlorpyrifos Epidemiology Regarding Children’s Health

4.1 Scope and Purpose 

In September 2008, EPA presented to the FIFRA SAP its preliminary review of available epidemiologic 

investigations of prenatal exposure to chlorpyrifos in association with measures of fetal growth and 

adverse neurodevelopmental effects in three major prospective children’s health cohorts in the U.S.18

These are: 1) The Mother’s and Newborn Study of North Manhattan and South Bronx performed by 

Columbia University researchers referred in this document as “Columbia Mother’s and Newborn Study;” 

2) the Mt. Sinai Inner-City Toxicants, Child Growth and Development Study or the “Mt. Sinai child 

growth and development study;” and, 3) Center for Health Assessment of Mother’s and Children of 

Salinas Valley (CHAMACOS) conducted by researchers at University of California Berkeley, the 

“CHAMACOS study.” In this meeting, EPA updates and expands its targeted evaluation of this important 

line of evidence regarding chlorpyrifos developmental neurotoxicity. 

At the previous meeting, the Panel agreed with EPA’s conclusions that “chlorpyrifos likely played a role 

in the birth and developmental outcomes noted in the three cohort studies” (pp. 37 Meeting Minutes). 

In support of this statement, the Panel offered that investigations performed within these three 

epidemiological cohorts utilized a similarly strong study design (prospective cohort); measured exposure 

using several different methods including specific and non-specific biomarkers of chlorpyrifos; 

ascertained developmental outcomes using validated assessment tools common to both clinical and 

research settings; and, analyzed, selected and statistically adjusted for potentially confounding 

variables using reasonable and appropriate methods. Overall, the Panel noted that the epidemiological 

database at that time presented an informative body of evidence with some notable consistencies 

across studies. Areas of inconsistency were also observed, and judged, in part, to be due to differing 

methods of measurement and evaluation, as well as dissimilar exposure profiles (i.e., residential versus 

direct and indirect occupational (farm laborer) exposure). Importantly, the Panel at the 2008 FIFRA SAP 

meeting also stated that it could not conclude that chlorpyrifos was the sole contributor to these 

outcomes, as co-exposure to other organophosphate pesticides and mixtures of environmental 

exposures may also have played a part in these outcomes. 

For the purpose of informing the chlorpyrifos risk assessment, the Panel concurred with the Agency with 

respect to the primacy of the Columbia Mother’s and Newborn Study among the three birth cohort 

studies, although the Panel also encouraged the Agency to comprehensively consider the results of the 

three children’s health cohorts. While the Columbia Mother’s and Newborn Study researchers measured 

the parent compound chlorpyrifos, as opposed to non-specific organophosphate metabolites, the other 

cohorts examined health outcomes (the Brazelton index of neonatal development) and performed 

supplemental analyses (e.g., effect modification by PON1 status) not reflected in the Columbia Mother’s 

and Newborn Study database. Therefore, the Panel expected EPA could strengthen its understanding of 

the potential developmental neurotoxicity of chlorpyrifos by considering the three children’s health 

cohort studies together. Additionally, the Panel in 2008 suggested supplemental statistical analyses to 

enhance understanding of epidemiological study results in the risk assessment context (See Appendix 4). 

The Panel also generally noted both strengths and limitations of these studies, and offered that random 

 18 See Meeting Minutes at: 

http://www.epa.gov/scipoly/sap/meetings/2008/september/sap0908report.pdf.
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or systematic errors in the design, conduct or analysis of these studies were unlikely to fully explain 

observed associations. However, the Panel also noted that absent the available toxicological and 

epidemiological databases was a well understood and defined mode of action as to the role of 

chlorpyrifos exposure in the etiology of adverse infant and child neurodevelopmental outcomes.

At this time, EPA expands and updates its review of the available epidemiologic data concerning the 

effect of chlorpyrifos exposure on children’s environmental health in conjunction with a review of recent 

experimental studies and hypothesized adverse outcome pathways (AOP) (See Section 2.0).

Observational studies published subsequent to the September 2008 FIFRA SAP evaluation extend the 

knowledge base of potential long-term sequalae of prenatal chlorpyrifos exposure. Specifically, in April 

2011, researchers with each of the three prospective children’s health cohort studies concurrently 

published results of their respective investigations of prenatal chlorpyrifos exposure and measures of 

intelligence among school aged children approximately 7-years (Bouchard et al., 2011; Engel et al., 2011; 

V. Rauh et al., 2011). Additionally, researchers with the Mt. Sinai Child Growth and Development study 

contributed their evaluation of organophosphate exposure and mental and psychomotor development, 

and authors with the CHAMACOS study published an evaluation of chlorpyrifos exposure and both fetal 

growth and neurodevelopment measures in young children as modified by paraoxonase-1 (PON1) 

genotype and phenotype (Eskenazi et al., 2010; Harley et al., 2011). 

Researchers have also performed epidemiologic methods research which in many ways reduces 

uncertainties related to key measures within these studies. Within the Columbia Mothers and Newborn 

study, investigators published results of analyses evaluating the validity of prenatal chlorpyrifos 

exposure measures in the time periods immediately after the voluntary cancellation of chlorpyrifos for 

residential use (Whyatt et al., 2009; Whyatt et al., 2007), as well as employed innovative statistical 

techniques to further assess the potential confounding bias of socio-economic status (SES) in the 

relation between prenatal chlorpyrifos exposure and adverse neurodevelopmental health outcomes 

(Lovasi et al., 2011). Overall, these additions to the epidemiologic database concerning children’s health 

effects in combination with other lines of evidence as discussed within this document (see sections 2, 3, 

and 5), add to the body of knowledge available to inform the ways in which chlorpyrifos exposure may 

be related to adverse neurodevelopment outcomes in children. 

4.2 Summary of Epidemiology Findings

In this section, EPA summarizes and critically reviews epidemiologic studies of prenatal chlorpyrifos 

exposure and subsequent fetal growth and child development evaluated within the three prospective 

children’s health cohorts described above. Specifically, in this section the design, conduct and methods 

of analysis of each cohort study are briefly presented; individual study results are summarized by type of 

health outcome investigated; and strengths and limitations of these investigations are discussed. 

Appendix 5 includes evidence tables summarizing details of each investigation, and Appendix 6 includes 

detailed study reviews and critical analysis of each investigation. The following section 4.3 reflects EPA’s 

synthesis and evaluation of the current chlorpyrifos epidemiology database. In accordance with the 

draft “Framework for Incorporating Epidemiology in Risk Assessment,”19 this analysis considers the 

strengths and limitations reflected in each cohort and research study as well as modified Bradford Hill 

considerations for the synthesis of these data. 

 19 http://www.epa.gov/scipoly/sap/meetings/2010/020210meeting.html#materials
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4.2.1 Overview of Design and Methods of Children’s Health Studies

These cohorts were recruited for the purpose of studying the potential health effects of environmental 

exposures during pregnancy on subsequent child development: The Columbia University’s Mother’s and 

Newborn Cohort (Lovasi et al., 2011; Rauh et al., 2011; Rauh et al., 2006; Whyatt et al., 2009; Whyatt et 

al., 2007; Whyatt et al., 2004); The Mount Sinai Hospital Children’s Environmental Health Cohort 

(Berkowitz et al., 2004; Engel et al., 2007; Engel et al., 2011); and UC Berkeley’s the Center for the 

Health Assessment of Mothers and Children of Salinas (CHAMACOS) Cohort (Bouchard et al., 2011; 

Eskenazi et al., 2004; Eskenazi et al., 2010; Eskenazi et al., 2007; Harley et al., 2011; Marks et al., 2010; 

Young et al., 2005). These studies enrolled pregnant women at baseline and prospectively assessed 

associations in their newborns and young children. Multiple reports have been published based on the 

findings in each cohort. 

The three study populations reflect different exposure profiles. The Columbia Mother’s and Newborn 

study and the Mt. Sinai Child Growth and Development study participants were likely exposed to 

chlorpyrifos through residential use of the pesticide for indoor pest control. In the residential setting, 

chlorpyrifos was among the most widely used household pesticides in the US during the time period of 

these epidemiologic investigations. However, pesticide companies voluntary cancelled indoor residential 

uses of chlorpyrifos-containing pesticide products on December 31, 2000, during the time period of 

these cohort investigations. In the agricultural setting, chlorpyrifos was registered for use on commonly 

consumed crops such as corn, almonds, apples and oranges. Therefore, these study populations were 

most likely additionally exposed to chlorpyrifos via the oral route through ingesting residues in the diet 

and from hand-to-mouth contact with in-home chlorpyrifos-contaminated surfaces, as well as possible 

dermal exposure through contact with treated areas in the home environment (Whyatt et al., 2003; 

Whyatt et al., 2009; Whyatt et al., 2007). In contrast, CHAMACOS cohort participants were employed as 

farm laborers or were residing in homes with farm laborers. These participants likely experienced either 

occupational exposure through the inhalation and dermal routes, as well as probable indirect exposure 

through drinking water and take-home exposures (Bradman et al., 2007). In each of the children’s health 

cohorts, the prevalence of pesticide exposure was high; however, reported use of chlorpyrifos in the 

CHAMACOS region was modest (10% of total pesticide use) (Eskenazi et al., 2004). 

4.2.1.1 The Mother’s and Newborns Study of North Manhattan and South Bronx 

(Columbia University)

Researchers with the Columbia Mother’s and Newborn Study evaluated the association between 

prenatal exposure to pesticides including chlorpyrifos and developmental outcomes in children through 

age 7 years. In this birth cohort study, participants were recruited during early pregnancy (≤20th week) 

among African-American and Dominican women age 18-35 years, and registered for prenatal care and 

delivery at either New York Presbyterian Medical Center or Harlem City hospitals. Women who smoked, 

had a history of drug abuse, diabetes, hypertension, or HIV infection were excluded from participation in 

the study, as were women who resided in New York City for less than 1 year. The study samples 

represented in the reports reviewed were recruited between 1998 and 2004, a period which overlaps 

the voluntary cancellation of chlorpyrifos use in the residential environment. 

In this cohort, authors measured chlorpyrifos exposure in several different biological and environmental 

matrices. These include chlorpyrifos parent compound in cord blood; the chlorpyrifos metabolite 3,5,6-

TCPy in maternal and infant urine and meconium; and chlorpyrifos in personal and stationary air 

monitoring samples. In epidemiologic analyses, investigators consistently utilized cord blood measures 
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of chlorpyrifos as the measure of prenatal exposure. Chlorpyrifos levels in umbilical cord blood samples 

were sampled as close to the time of delivery as possible, and within 2 days post-partum. Cord blood 

plasma chlorpyrifos levels were imputed from maternal blood levels for newborns for whom no cord 

blood sample was obtained because correlation was high (>80%). Quantification of chlorpyrifos levels in 

plasma were conducted by the Centers for Disease Control and Prevention (CDC). Information regarding 

basic demographics, socio-economic status, and pregnancy related measures, among other factors was 

collected through self-report questionnaire at the time of enrollment. 

4.2.1.2 Inner-City Toxicants, Child Growth and Development Study (Mt. Sinai 

Hospital)

Researchers with the Mt. Sinai Children’s Health and Development study conducted a prospective birth 

cohort study in which they enrolled primiparous women presenting for prenatal care with singleton 

pregnancies at the Mount Sinai prenatal clinic and two private practices and who delivered their infants 

at Mount Sinai Hospital in New York City between May 1998 and July 2001. Mothers were excluded if 

they had any of the following characteristics: an initial prenatal visit after 26 weeks of gestation, serious 

chronic diseases, a serious pregnancy complication that could affect fetal growth and development. 

Additional, participants were excluded for risky health behaviors including alcohol consumed greater 

than two alcoholic beverages per day or illicit drug use. Mothers and infants were also excluded if the 

child was born with a congenital malformation or severe prematurity. 

To measure prenatal pesticide exposure, researchers implemented a self-report questionnaire to solicit 

information regarding pesticide usage, in-home pest pressure and other exposure characteristics. In 

addition, in the early third trimester, participants were asked to provide a urine sample at the time of a 

routine clinical blood draw. Using this biological sample, authors measured urinary concentration of 

pesticide metabolites including both TCPy (Berkowitz et al., 2004), and also non-specific measures of 

organophosphate exposure, DAPs (Engel et al., 2007; Engel et al., 2011). Authors adapted analytical 

methods to conduct laboratory measurement of TPCy; DAPs were measured by CDC using published 

methods (Barr et al., 2002). Using maternal and infant (cord blood) blood sample, researchers measured 

PON1 enzymatic activity levels, and conducted genotyping analysis to determine prevalence of PON1

variant alleles. 

Potentially confounding variables were measured through self-report questionnaire and included in final 

statistical models if variables were known to be associated with either pesticide exposure or fetal 

growth. Authors also measured concentration of other pesticides in biological matrices including 

metabolites of pyrethroid exposure (PBA), pentachlorophenol (Berkowitz et al., 2004), as well as other 

organophosphates (malathion) and organochlorines compounds including polychlorinated biphenyls 

(PCBs) (Engel et al., 2007). However, authors did not present analyses concerning potential confounding 

effect of co-exposure to other environmental chemicals.

4.2.1.3 Center for Health Assessment of Mother’s and Children of Salinas Valley, 

CHAMCOS (University of California/Berkley)

The Center for the Health Assessment of Mothers and Children of Salinas (CHAMACOS) cohort is 

comprised by participants who live and work in the Salinas Valley, CA. The Salinas Valley is a major 

center of agricultural production in the United States with approximately 500,000 pounds of 

organophosphate pesticides applied annually (California EPA. Pesticide Use Reporting 2001 Summary 

Data, 2002. (www.cdpr.ca.gov/docs/pur/pur01rep/01_pur.htm)), chlorpyrifos, however, was not 
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frequently used in agriculture in this area during the time period of this study. This cohort is comprised 

of low-income, predominantly Mexican-American (or Mexican immigrant) women employed as farm 

laborers or living with someone employed as a farm laborer. Enrollment of the Center for the Health 

Assessment of Mothers and Children of Salinas (CHAMACOS) cohort took place at regional community 

clinics. Women considered eligible for the study were less than 20 weeks gestation, aged 18 years or 

older, Medi-Cal eligible, fluent in English and/or Spanish, and planning to deliver at Natividad Medical 

Center. Excluded from analyses were women with gestational or preexisting diabetes, hypertension, 

twin births, or stillbirths. 

Data collection was performed using several different tools. Authors administered self-report 

questionnaires to study participants to obtain information regarding demographic characteristics, work 

history and health behaviors. To assess exposure, participants were asked to provide biological samples 

(urine, blood). Specifically, maternal and fetal exposure to organophosphate pesticides was assessed by 

measurement of organophosphate DAPs and seven different pesticide-specific metabolites, including

TCPy, in maternal urine during two periods in the pregnancy. Maternal urine samples collected between 

5 and 27 weeks gestation and again between 18 and 39 weeks. The post-delivery urines were collected 

within 1 week of delivery for 73% of the sample, with the remainder obtained up to 176 days 

afterwards. Total DAP, including DMP, and DEP levels were determined for each participant for each of 

the two pregnancy urine samples, and because the measures did not differ substantially, these values 

were averaged in epidemiologic analyses to estimate prenatal pesticide exposure. Quantification of 

organophosphate metabolites was conducted by the Centers for Disease Control and Prevention (CDC)

labs (Barr et al., 2002). 

In this cohort, authors also measured ChE in whole blood and BuChE in plasma as a surrogate for OP 

exposure. PON1 was also measured in blood. Researchers also measured concentration of other 

environmental chemicals including PCBs, lead, DDT/DDE, HCB, and PBDEs; however, these compounds 

were not included in final models as potential confounding variables.

4.2.2 Summary of Research Results

Across these three children’s health cohorts, authors have assessed the relation between measures of 

prenatal chlorpyrifos exposure and various measures of fetal growth and infant and child 

neurodevelopment. In this section, research results are briefly summarized by health outcome category. 

An evidence table summarizing key study features is presented in Appendix 5. A more detailed 

description of the study design, methods and analysis, as well as research results and a critique of 

individual study strengths and limitations are presented in Appendix 6. Table 10 briefly summarizes 

these results.

4.2.2.1 Measures of Fetal Growth

Authors with each of the three respective children’s health cohorts measured prenatal chlorpyrifos 

exposure in association with fetal growth including birth weight, birth length, head circumference, 

gestational age and Ponderal index. To ascertain birth characteristics, authors linked with medical 

records at respective participating hospitals. Across these children’s health cohorts, authors observed 

inconsistent evidence of an association; however, differing exposure profiles as well as dissimilar 

methods of prenatal chlorpyrifos exposure assessment likely played a role in this observation 

(Needham, 2005).
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Table 10 Summary of Findings in the Columbia Mothers and Newborn Cohort, the Mount Sinai Hospital Children’s Environmental 

Health Cohort, and the UC Berkeley Center for the Health Assessment of Mothers and Children of Salinas Studies of Prenatal 

Organophosphate Pesticide Exposure and Child Development.

Columbia Cohort Mount Sinai Cohort UC Berkeley Cohort

Markers of Exposure: Chlorpyrifos

TCPy 

(fetal 

growth)

DAPS 

(BNBAS, Bayley, Wechsler)

TCPy 

(fetal growth, 

Bayley);

DAPS (BNBAS, 

Bayley, Wechsler)

Birth Length Inverse

(Null post 2000) Null Positive (NS)

Birth Weight Inverse

(Null post 2000) Null Positive (NS) Positive*

Head Circumference at 

Birth Null Null

Inverse* Null Positive*

Gestational Age -- Null Inverse* 

BNBAS Newborn 

(Abnormal reflexes) -- Positive* Positive

BNBAS Newborn 

(Neurodevelopment) -- Null Null

Bayley Scores 

6 months (MDI/PDI) -- -- Null/Inverse (NS) Inverse 

(NS)/Inverse (NS)

Bayley Scores 

12 months (MDI/PDI) Null / Null

Positive** (among blacks 

and Hispanics) / Positive 

(NS)

Null / Null

Bayley Scores 

24 months (MDI/PDI) Null / Null Inverse* /Null* Null / Null Inverse / Null

Bayley Scores 

36 months (MDI/PDI) Inverse/ Inverse -- --

Pervasive Development

Disorder (PDD)

Positive 

(36 mo.) -- Positive (24 mo.)

Mental Development 

(WPPSI-III, age 6 years) -- Null --

Mental Development

(WISC-IV, age 7-9 years)

Inverse (Full-scale 

IQ and Working 

Memory);Null 

(Others)

Inverse (NS) (Full scale IQ, 

perceptual reasoning, 

verbal comprehension, 

working memory and 

processing speed)

Inverse

Odds of ADHD/ 

attention and behavior 

problems at age 7 years.

-- --

Positive

(NS 3 years)

Positive (5 years)

*Interaction observed between pesticide markers and PON1 activity or genotype

**Interaction observed between pesticide markers and race

NS=Not statistically significant

Inverse= Higher levels of exposure associated with adverse health outcomes (measurement value or score decreased)

Positive = Higher levels of exposure associated with adverse health outcome (measurement value or score increased)

Null = No association observed

MDI= Mental Development Index

PDI = Psychomotor Development Index 
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Researchers with the Columbia Mothers and Newborn study observed an association between 

decreased fetal growth and prenatal chlorpyrifos measures among 314 mother-infant pairs selected for 

this study (Whyatt et al., 2004). Controlling for potential confounders, for each log unit increase in cord 

plasma chlorpyrifos levels, birth weight decreased by 42.6 g (95% CI: –81.8 to –3.8) and birth length 

decreased by 0.24 cm (95% CI: –0.47 to –0.01). Whyatt et al. (2004) did not report an association with 

head circumference. Combined exposure to both chlorpyrifos and diazinon (adjusted for relative 

potency using US EPA cumulative risk assessment methods) were also significantly inversely associated 

with birth weight and length (p < 0.05). When births were stratified by time period prior to or after the 

voluntary cancellation of chlorpyrifos for residential use, researchers no longer observed evidence of an 

association (p > 0.8) among births which took place in the cancellation period (n=77). In supplementary 

analyses, authors replicated this analysis with additional post-cancellation era births (total n=193) and 

similarly did not observe evidence of an association with chlorpyrifos in the later time period (Whyatt & 

Rauh, 2011) (See Appendix 4). Overall, authors suggest that prenatal chlorpyrifos exposures may have 

impaired fetal growth among this inner-city, low income cohort. While suggestive of a conclusion that 

cessation of chlorpyrifos exposure explains the different associations with fetal growth before and after 

the period of the voluntary cancellation of chlorpyrifos, EPA notes that this finding may also be 

suggestive of a possible threshold effect, or it could also be due to a lack of statistical power to assess 

associations by time period. 

Within the Mt. Sinai Child Growth and Development study, Berkowitz et al. (2004) assessed the 

association between prenatal chlorpyrifos exposure measured as urinary TCPy and subsequent risk of 

impaired fetal growth (Berkowitz et al., 2004). Authors also evaluated potential effect modification by 

PON1 genotype and phenotype in the relation of interest. Among 404 births which took place between 

1998 and 2002, authors found no statistically significant associations between fetal growth including 

birth weight, or birth length or head circumference and chlorpyrifos (estimated as TCPy) exposure. 

However, Berkowitz et al. (2004) did observe evidence of heterogeneity of effect by PON1 activity level. 

Specifically, researchers observed a small, but statistically significant reduction in head circumference 

among children of mothers with levels of chlorpyrifos above the limit of detection and also in the lowest 

tertile of PON1 activity (least able to metabolize exogenous exposures such as OP pesticides). In the 

subgroup of infants whose mothers had TCPy levels greater than the level of detection, those with low 

maternal PON1 had an average (SD) head circumference of 33.3cm (1.5cm) which was significantly 

smaller than those with medium (34.0cm (1.5cm)) and high (34.1cm (1.6cm)) maternal PON1 activity 

after adjusting for race/ethnicity, infant sex, and gestational age (p = 0.014), although the statistical 

interaction was not significant. Authors did not observe evidence of heterogeneity of effect by PON1 in 

the relation between other measures of birth outcomes including birth weight and birth length and 

pesticide exposure. Neither maternal PON1 genetic polymorphisms nor infant paraoxonase levels were 

associated with reduced head size. 

Within the CHAMACOS cohort, among 488 participants enrolled between 1999-2000 Eskenazi et al. 

(2004) did not observe a significant adverse relationship between fetal growth and any measure of in 

utero organophosphate pesticide exposure (Eskenazi et al., 2004). On the contrary, investigators 

reported positive associations between birth length and head circumference associated with nonspecific organophosphate exposure measures (DAPs). Researchers observed decreases in gestational 

age associated with measures of in utero pesticide exposure: urinary DMP metabolites (β = –0.41 weeks 

per log10 unit increase; 95% CI: –0.75––0.02; p = 0.02), which reflects exposure to DMP 

organophosphate compounds such as malathion, but not chlorpyrifos. Authors did not observe an 

association using TCPy as a measure of chlorpyrifos exposure; however they did report an increased risk 
Page 61 of 124

of preterm delivery with decreasing cholinesterase concentrations, a third biomarker of 

organophosphate exposure. To identify “critical windows” of fetal development when exposure may 

have a greater impact, the authors analyzed the associations of outcomes and metabolite levels 

measured during moving 6-week windows of pregnancy (e.g., 5–10 weeks, 6–11 weeks, 7–12 weeks) 

using a series of multiple regression analyses. No period of greater impact was observed in this largely 

null study.

In a follow-up analysis within this cohort, Harley et al. (2011) performed the same analysis however 

evaluating the potential effect modifying role of PON1 genotype and phenotype in the relation between 

prenatal pesticide and chlorpyrifos exposure (DAP, TCPy, ChE) and fetal growth (Harley et al., 2011). In 

this study, infants’ (but not mother's) PON1 genotype and PON1 activity modified the association 

between gestational age and head circumference. Infants with the susceptible PON1-108 TT genotype had 

shorter gestational age (beta =-0.5 weeks, 95%CI: -0.9, 0.0) and smaller head circumference (beta=-0.4 

cm, 95% CI: -0.7, 0.0) than those without the susceptible genotype (PON1-108 CC genotype). Infants’ 

arylesterase and paraoxonase activity were positively associated with gestational age. Maternal DAP 

concentrations were associated with shorter gestational age among infants of the susceptible PON1-108

TT genotype, although only the interaction between PON1-108 genotype and DEP metabolite 

concentrations was statistically significant (p-value for interaction = 0.09). However, maternal DAP 

concentrations were associated with larger birth weight (p-value for interaction = 0.06) and head 

circumference (p-value for interaction =0.01) in infants with non-susceptible genotypes. The authors 

conclude that infants with certain PON1 genotypes (e.g., PON1-108 TT) may be more susceptible to 

effects of in utero organophosphate pesticide exposure. 

4.2.2.2 Brazelton Neonatal Neurological Functioning

Researchers with both the Mt. Sinai Child Growth and Development study and the CHAMACOS cohort 

evaluated neonatal neurological functioning in association with prenatal chlorpyrifos exposure. To 

measure indices of abnormal neonatal behavior and/or neurological integrity authors used outcome 

measures derived from the Brazelton Neonatal Behavioral Assessment Scale (BNBAS). The BNBAS 

includes 28 behavioral items and 18 primitive reflexes which assesses the infant across several different 

developmental areas. This tool was administered to infants within days after birth and before they left 

the hospital (2-5 days post-partum). Examinations were conducted by trained neonatologists in the 

hospital setting using similar environmental conditions. The Mt. Sinai Child Development study and the 

CHAMACOS cohort evaluated this outcome measure; researchers with the Columbia Mother’s and 

Newborn study did not measure the relation.

Among the 438 infants eligible to participate in the Mt. Sinai evaluation, Engel et al. 2007 observed an 

association between generic (DAPS) biomarkers of prenatal organophosphate pesticides and an 

increased number of abnormal primitive reflexes which are considered a critical marker of neurologic 

integrity. Controlling for confounding factors, subjects with prenatal total DEP levels above the median 

(24.7 nmol/L) delivered infants who were 2.3 times more likely to have at least two abnormal reflexes 

(95% CI: 1.1, 5.0). Associations with other DAPs were also reported. Notably, no statistically significant 

adverse associations were found for any of the DAP metabolites and other domains of the Brazelton 

assessment tool in the author’s primary analyses, e.g., habituation, range of state, orientation. Authors 

also observed evidence of potential effect modification by PON1 activity level in the relation between 

DAPs and neonatal neurodevelopment. Specifically, in the first tertile of paraoxonase 1 expression 

(slowest metabolizers), the relative risks (RR) of having abnormal reflexes were significantly related to 

DAP and DMP levels; the risk estimate for DEP and the number of abnormal reflexes within the lowest 
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PON1 level was not significant. In the CHAMACOS cohort, Young et al. (2005) observed a statistically 

significant association between organophosphate pesticide exposure and the reflex cluster of the BNBAS 

(Young et al., 2005). The proportion of infants with more than three abnormal reflexes was 3-fold 

increased among those with prenatal DEP exposure (DEP: OR = 3.4, 95% CI = 1.2, 9.9). Similar to Engel et 

al. (2007), no other associations between prenatal DAPs and other aspects of the Brazelton assessment 

were noted (other neurodevelopmental domains) (Engel et al., 2007). No adverse associations were 

found between postnatal urinary metabolite levels and any of the developmental outcomes.

4.2.2.3 Bayley Scale of Mental and Psychomotor Development

Researchers across the three children’s health cohorts utilized the Bayley Scales of Infant Development 

II (BSID-II) to generate a Mental Development Index (MDI) and a Psychomotor Development Index (PDI) 

to assess neurodevelopment in early childhood. As a complement to this measure of mental 

development and behavior, authors also used the 99-item Child Behavior Checklist (CBCL). In addition, 

because the quality of the home environment is a key determinant of child development, authors also 

used the Home Observation for Measurement of the Environment (HOME) instrument. This instrument 

collects data regarding the physical and interactive qualities of the child’s home as a measure of mental 

stimulation and interactions. Results from each of these measurement tools were used to model the 

relation between prenatal chlorpyrifos exposure and adverse neurodevelopmental outcomes in infancy 

and toddlerhood (6-36 months of age).

Rauh et al. 2006 investigated Mental Development Index (MDI) and a Psychomotor Development Index 

(PDI) at 12, 24, and 36 months of age within the Columbia Mother’s and Newborn Study. Children were 

categorized as having either high (>6.17pg/g) or low (<6.17pg/g) prenatal exposure, using categories 

informed by results of the previous study on birth characteristics (Whyatt et al., 2004). Authors reported 

that the difference in MDI scores was “marginally significant” (p = .06) with the exposed group scoring 

an average of 3.3 points lower. When the same multivariate regression models were calculated 

regarding the PDI scores, none of the 12 or 24 month PDI scores showed significant effects, but the 36 

month score was significantly related to chlorpyrifos exposure. Investigators also calculated estimates of 

adjusted risk for developmental delays (MDI and PDI) related to chlorpyrifos exposure and illustrated 

that before 36 months of age, delays were no more likely in the highly exposed group, but, at the 36 

month milestone, the likelihood of highly exposed children developing mental delays were 2.4 times 

greater (95% CI: 1.12-5.08, p = .02) and motor delays were 4.9 times greater (95% CI: 1.78-13.72; p = 

.002) than those with lower prenatal exposure. Using general linear modeling (GLM), authors analyzed 

developmental trajectories of the effects and results were consistent with the 12, 24, and 36 month 

milestone analyses indicating that the age effects most significantly occurred at the later phase of the 3 

year period. In supplemental analyses suggested by the 2008 FIFRA SAP, authors illustrated that 

diazinon was a strong confounding variable in this association (correlation with chlorpyrifos 0.63), 

increasing the magnitude chlorpyrifos risk estimate for MDI and PDI 50-200% in the same direction 

(away from the null) (Whyatt & Rauh, 2011).

Within the Mt. Sinai Children’s Environmental Health study, authors administered the BSID-II to 

participating children at 12 and 24 months. Among 404 women originally enrolled between May, 1998 

and July, 2001 (n = 404), children of these mothers returned for neurodevelopment assessments at ages 

12 months (n = 200), 24 months (n = 276) of age. Using generalized linear models, authors found that 

prenatal total DAP metabolite level was associated with a decrement in mental development at 12 

months among blacks and Hispanics children. The associations appeared to be strongest among children 

of mothers who carried the PON1 Q192R QR/RR genotype. The authors concluded that their findings are 
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suggestive of an association between prenatal exposure to organophosphates and decrements in 

cognitive development, particularly perceptual reasoning, with evidence of effects beginning at 12 

months and continuing through early childhood, with PON1 being a potentially important susceptibility 

factor for these deleterious effects. The authors did not observe any effect modification by PON1 

enzyme activity level. In general, associations observed during the 12-month follow-up were either 

attenuated or non-existent at the 24-month visit.

In the CHAMACOS cohort, Eskenazi et al 2007, the authors report on the relationship between prenatal 

and child urinary organophosphate metabolite levels and child neurodevelopment at ages 6, 12, and 24 

months of age. Controlling for indicators such as the psychometrician conducting the assessment and 

the location of assessment, age at assessment, sex, duration of breast-feeding, HOME score, and 

household income, parity and indicator of maternal intelligence (PPVT score), authors observed that 

prenatal DAP levels were adversely associated with MDI, while early life DAP levels were positively 

associated with MDI. At 24 months of age, these associations reached statistical significance (per 10-fold 

increase in prenatal DAPs: β = –3.5 points; 95% CI: –6.6 to –0.5; child DAPs: β = 2.4 points; 95% CI: 0.5 to 

4.2). In a subsequent study, investigators did not observe evidence of effect modification by PON1 status 

in the relation between prenatal DAPs and child neurodevelopment as measured by the Bayley Scale 

(Eskenazi et al., 2010). Neither prenatal nor child DAPs were associated with PDI or CBCL attention 

problems. Both prenatal and postnatal DAPs were associated with risk of pervasive developmental 

disorder (per 10-fold increase in prenatal DAPs: OR = 2.3, p = 0.05; child DAPs OR = 1.7, p = 0.04). TCPy 

as a biomarker of chlorpyrifos exposure was not associated with any neurodevelopment outcome in this 

study. 

4.2.2.4 Attention Problems

Also within the CHAMACOS cohort, Marks et al (2010) conducted a study to investigate the association 

between urinary DAP metabolites in pregnant women and their children, as a marker of 

organophosphate exposure, and attention-related outcomes among 348 children who had available 

data at 3.5 and/or 5 years and met inclusion criteria (Marks et al., 2010). Attention-related health 

outcomes were measured through maternal report of child behavior at 3.5 and 5 years of age; direct 

assessment of the child at 3.5 and 5 years; and by a psychometrician’s report of the behavior of the child 

during testing at 5 years. To identify children whose behaviors were most suggestive of possible ADHD 

within the cohort, a composite ADHD variable was defined that combined the results of the maternal 

report (CBCL), child testing (K-CPT), and the psychometrician report (Hillside). In this study population, 

higher concentrations of organophosphate metabolites in the urine of pregnant women were associated 

with increased odds of attention problems and poorer attention scores in their children at age 5 years. 

Prenatal DAPs were non-significantly associated with maternal report of attention problems and ADHD 

at age 3.5 years but were significantly related at age 5 years (CBCL attention problems: β = 0.7 points; 

95% CI: 0.2–1.2; ADHD: β = 1.3; 95% CI: 0.4–2.1). Prenatal DAPs were associated with scores on the KCPT ADHD Confidence Index > 70th percentile (OR = 5.1; 95% CI: 1.7–15.7 and with a composite ADHD 

indicator of the various measures (OR = 3.5; 95% CI: 1.1–10.7). Some outcomes exhibited evidence of 

effect modification by sex, with associations found only among boys. Children’s concurrent total DAP 

and DMP metabolite levels at 3.5 years and 5 years were unrelated to attention outcomes, and but child 

DEP concentrations at 5 years were adversely associated with the composite measure of attention (OR = 

2.0; 95% CI:1.1–3.6). The results of this investigation by Marks et al. (2010) in the CHAMACOS cohort are 

suggestive of a detrimental association between prenatal organophosphate exposure, as measured by 

maternal urinary metabolite levels, and attentional difficulties at age 5 years using three different 

measures of this neurodevelopmental outcome. 
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4.2.2.5 Intelligence Measures

To measure intelligence among school aged children, authors from each of the three children’s health 

cohorts used the Wechsler Intelligence Scale for Children, 4th edition (WISC-IV). The instrument 

measures four areas of mental functioning: the Verbal Comprehension Index, the Perceptual Reasoning 

Index, the Working Memory Index, and the Processing Speed Index. These indices are associated with, 

but distinct from, overall intelligence quotient (IQ) and are sensitive to cognitive deficits related to 

learning and working memory. A Full-Scale IQ score combines the four composite indices. A General 

Ability Index score is a summary score of general intelligence, similar to Full-Scale IQ, but excludes 

contributions from both Working Memory Index and Processing Speed Index. WISC-IV scores are 

standardized against U.S. population-based norms for English and Spanish-speaking children.

Rauh et al. (2011) evaluated the relationship between prenatal chlorpyrifos exposure and 

neurodevelopment among 265 of the cohort participants who had reached the age of 7 years and had a 

complete set of data including prenatal maternal interview data, prenatal chlorpyrifos marker levels 

from maternal and/or cord blood samples at delivery, postnatal covariates, and neurodevelopmental 

outcome data (Rauh et al., 2011). While models were developed using continuous measures of both 

prenatal chlorpyrifos exposure and Wechsler scores, for ease of interpretation, investigators reported 

that for each standard deviation increase in exposure (4.61pg/g) there is a 1.4% reduction in Full-Scale 

IQ and a 2.8% reduction in Working Memory. 

In the Mt. Sinai Child Growth and Development study, Engel et al. 2011 used generalized linear models 

to analyze the relationship between total DEPs, total DMPs, total DAP metabolites measured in 

maternal urine during the third trimester, and subsequent cognitive/intelligence development evaluated 

at age 7 years. In this study, prenatal maternal DEP urinary metabolite concentrations were associated 

with slight decrements in FSIQ, Perceptual Reasoning, and Working Memory between the ages of 6 and 

9 years. Among children of mothers with the susceptible PON1 genotype, DAP and DMP urinary 

metabolite concentrations were associated with poorer scores on Perceptual Reasoning and FSIQ. 

In the CHAMACOS cohort, Bouchard et al. (2011) observed evidence of an association between prenatal 

exposures to organophosphate pesticides as measured by urinary DAP metabolites in women during 

pregnancy, and decreased cognitive functioning in children at age 7 (Bouchard et al., 2011). Authors 

observed this finding using total DAP, DMP, and DEP metabolites to estimate pesticide exposure. 

Children in the highest quintile of maternal DAP concentrations had an average deficit of 7.0 IQ points 

compared with those in the lowest quintile. Authors reported the associations were linear with no 

threshold. Urinary DAP concentrations in childhood (postnatal) were not associated with cognitive 

scores in this cohort of children. Of note, the following known or suspected neurotoxicants were 

measured: polybrominated diphenyl ethers (PBDEs), polychlorinated biphenyls (PCBs), p,p’-

dichlorodiphenyltrichlorethane (DDT), p,p’- dichlorodiphenyltrichlorethylene (DDE), and lead. Lead was 

measured in maternal blood at 26 weeks gestation, in cord blood for a subset of participants, and 

children’s blood a 2 years of age. These other environmental exposures including blood lead 

concentration did not change the magnitude and direction of the associations observed between 

DAP/DEP and IQ measures among school-aged children.
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4.2.3 Strength and Limitations of Children’s Health Cohort Studies 

There are several strengths reflected in the design, conduct and analysis decisions made by investigators 

with each of the three children’s health studies, as well as some limitations to consider. Strengths of 

these studies include the prospective nature of these investigations, extensive data collection including 

several different measures of prenatal chlorpyrifos exposure which have been validated, and use of 

neurodevelopmental outcome ascertainment tools which have been validated in different populations 

and commonly used in both clinical and research settings, among other strengths noted in this section. 

Key limitations reflect the difficulty in measuring multi-dimensional characteristics such as socioeconomic status accurately, the challenge of conducting environmental epidemiology studies in 

populations exposed to several different compounds and mixtures, and the ability to assess effect 

modification (e.g., genetic variability) with precision and accuracy. Considering both the strengths and 

limitations of the respective investigations, EPA puts forward epidemiologic inference at the conclusion 

of this section. 

Study authors from each of the respective long-term studies hypothesized a role for environmental 

exposures, including but not limited to pesticides generally and chlorpyrifos specifically, in the etiology 

of adverse fetal growth and neurodevelopmental outcomes. Study authors stated their motivation was 

the knowledge that these populations were highly exposed to pesticides and other environmental 

contaminants, and that prior research had linked in utero environmental exposure to adverse 

neurodevelopmental outcomes. Cohorts were each funded through the EPA and National Institute of 

Environmental Health Science (NIEHS) Children’s Environmental Health and Disease Prevention Research 

Centers20. Cohort studies were comprised of several hundred mother-infant pairs (range: n=102 to 

n=488) and were likely sufficiently statistically powered to detect hypothesized main effects, but studies 

may have lacked statistical power to perform stratified analyses, e.g., effect modification by genetic 

variant. Similarly, statistical methods utilized across these investigations were appropriate to the data

collected and research question hypothesized, however statistical analyses were discussed in varying 

degrees of detail by separate study groups. Statistical model selection was generally parsimonious in 

nature, and investigators performed multiple sensitivity analyses (not always shown) within each study 

to further explore and eliminate alternative explanations for observed results, e.g., confounding by 

blood lead, variable transformations. Concerning the main effects hypothesized, both design and 

analysis decisions made by researchers enhanced the probability of identifying an association, if an 

association exists. For the association under study, measurement errors likely occurred due to the 

challenge of estimating exposure during the critical window of development.

Selection criteria were clearly defined and appropriate across the three cohorts. Investigators limited 

participants to individuals within a similar racial and ethnic group characteristics, income and education 

level, and geographic area, and who reported low prevalence of risky health behaviors (alcohol use, 

smoking during pregnancy), and who did not report major medical co-morbidities which may have 

adversely affected fetal growth and development (e.g., gestational diabetes). Therefore, authors sought 

to control through design (restriction) variability in the sample population to isolate the effect of 

chlorpyrifos exposure on fetal outcomes. However, the many selection criteria applied may have 

affected the generalizability of study results. The selected study population is narrowly defined and 

dissimilar to the general U.S. population across many demographic and exposure related characteristics. 

 20 http://epa.gov/ncer/childrenscenters/
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However, as detailed in Section 5.0, comparison with biomonitoring levels in the U.S. general 

population, exposure among these cohorts was greater, but within a similar range. In addition, the 

external validity of the study findings may be limited if the organophosphate pesticide exposure-fetal 

development association is modified by factors that are more, or less, prevalent in the study populations 

relative to the population(s) to which inference is being made, or to populations with a substantially 

different exposure range. 

All three studies are prospective cohort studies in which exposure measurements were obtained 

prenatally and/or at delivery, and fetal growth and neurodevelopmental outcomes were assessed 

subsequently. This study design eliminates or reduces several potential sources of bias. This design 

eliminates temporal bias, i.e., uncertainty as to whether exposure precedes the adverse outcome, and, 

and also reduces the effect of differential exposure measurement error (a type of information bias) as it 

is unlikely exposure measurement would be differentially assessed by (unmeasured, future) health 

outcome. 

Researchers utilized similar tools to measure adverse health outcomes across the studies. The use of 

similar outcome measures across studies aids comparability of studies and assessment of the 

consistency of results. Neurodevelopmental outcome assessment tools utilized were common to both 

clinical and research settings and most have been validated in both English and Spanish speaking 

populations. While measures of fetal growth are somewhat objective to measure, neurodevelopmental 

outcomes measured within these studies are more difficult to assess. While administered consistently 

within these investigations, tests of mental development, cognition, psychomotor development, and 

intelligence tools are somewhat subjective in nature, and may be affected by child’s anxiety (Bayley 

Scales) and influenced by cultural factors (Wechsler). However, the homogeneity of the study 

population may reduce the potential influence of a cultural bias or differences in assessment tool 

response based on external factors, and these are among the best (gold standard) of tools to assess 

these types of developmental issues. Authors employed appropriate quality assurance and control 

measures such as training of test administrators, periodic evaluation of adjudicators, consistent 

environments in which tests were administered, and in many instances included an indicator variable for 

test administrator in statistical analysis to adjust for slight differences in examiner effect. With the 

exception of the child behavior checklist (CBCL), errors in the measurement of health outcomes were 

likely non-differential in nature, leading to an attenuation of the risk estimates. Overall, across these 

studies researchers utilized the best available outcome measurement tools for neurodevelopment 

health effects, and implemented the evaluations in a consistent, standardized manner with trained 

health professionals and/or study staff. 

Across the three children’s health cohorts, study authors measured parent chlorpyrifos, TCPy and DEP to 

estimate chlorpyrifos and/or organophosphate exposure. Chlorpyrifos metabolites (TCPy and DAPs) are 

likely more accurate and objective indicators of organophosphate pesticide exposure than other 

exposure ascertainment methods such as self-report, but uncertainty remains as to the extent 

measurement of non-specific metabolites reflects chlorpyrifos exposure. TCPy is a metabolite specific to 

chlorpyrifos, but can also be produced as a result of chlorpyrifos-methyl exposure or environmental 

exposure to TCPy itself (Morgan et al., 2005; Wilson, Chuang, Lyu, Menton, & Morgan, 2003). Urinary 

DAP metabolites of organophosphorus insecticides are markers of exposure to organophosphate 

pesticides generally; chlorpyrifos and diazinon are metabolized to DEP, while other organophosphates 

are metabolized to DMP. For risk assessment purposes, it is difficult to infer chlorpyrifos effects 

specifically from urinary DAPs and to a lesser extent the TCPy metabolite. Different exposure 

measurements may in part explain inconsistencies across study findings, as noted by others (Needham, 
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2005). The three birth cohorts recruited participants who likely have chlorpyrifos exposures that are 

higher than the general population, and the exposures in these cohorts may also be more consistent 

over time. Inner-city populations report frequent use of pesticides, most applied at least once per week, 

during the time period of these studies (Whyatt et al., 2003; Whyatt et al., 2002). In New York City, 

location of two of the three children’s health cohort studies, housing authority reported common use of 

chlorpyrifos for indoor pest control during the time period of these studies, prior to the time period of 

the voluntary cancellation of these uses. In the CHAMACOS cohort, chlorpyrifos metabolite (TCPy) was 

detected in 77% of urine samples (Bradman et al., 2005; Eskenazi et al., 2004). Despite being considered 

to have greater opportunities for exposure relative to the general population, the average levels of 

chlorpyrifos biomarkers in the three cohorts were generally comparable to the U.S. general population, 

which may indicate a difference between actual and measured exposure. Nevertheless, it is not clear to 

what extent the use of one or two biomarker measurements conducted in the reported studies reflects 

exposure(s) over critical windows of development during pregnancy. This remains an uncertainty within 

these studies. As previously mentioned, chlorpyrifos exposure was assessed by quantification of 

biomarkers primarily in one- or two-samples of maternal urine taken during the third trimester, and 

also, in one cohort, in a single maternal blood or cord blood sample obtained at the time of delivery. If 

the exposure is chronic, a biomarker measured at a single time point may provide a representative 

dosimeter, even if the toxicant has a short half-life, as is the case for chlorpyrifos and organophosphate 

pesticides generally. However, if pesticide exposures are sporadic or otherwise vary over short time 

scales, the biomarker measurement may not be representative of “usual” exposure or of the exposure 

during critical periods of fetal development. In the UC Berkeley CHAMACOS cohort (Eskenazi et al., 

2004), prenatal urine was collected between 5 and 27 weeks gestation and again between 18 and 39

weeks. The within-person standard deviation in the DAP metabolites was approximately three times 

larger than the between-person standard deviation, and concluded that a single biomarker assessment, 

as was conducted in the other two cohorts, “may not accurately reflect exposures over the entire 

pregnancy.” 

This potential limitation was also assessed in two validation studies conducted by Whyatt et al within 

the Columbia Mother’s and Newborn Study. The first of these studies (Whyatt et al., 2007) assessed

within- and between-home variability in indoor-air insecticides over the final 2 months of pregnancy 

among a subset of participants. Authors observed little within-home variability and no significant 

difference in air concentrations within homes over time (p ≥ 0.2); between-home variability accounted 

for 92% of the variance in the indoor air levels of chlorpyrifos. Indoor and maternal personal air 

insecticide levels were highly correlated (r = 0.7–0.9, p < 0.001). While this study provides some 

evidence that assessment of chlorpyrifos exposure at a single time point may be reasonable, the study 

took place during a period of rapid decreasing indoor use of chlorpyrifos, and may not reflect variability 

in indoor air concentration prior to the period of cancellation. In addition to inhalational exposure, 

participants were likely also exposed by ingesting residues in the diet and from hand-to-mouth contact 

with surfaces contaminated with the pesticide. If true, then high correlation of chlorpyrifos air exposure 

over time may not indicate a true consistency in exposure levels. 

The variability in exposure measures and the validity of different biomarkers of exposure was evaluated 

in a second study (Whyatt et al., 2009) which evaluated trends over time in multiple biomarkers of 

prenatal chlorpyrifos exposure. Authors measured TCPy levels in repeat prenatal urine samples and 

determined they were positively, but only moderately correlated (r = 0.23–0.56), and within-subject 

variability exceeded between-subject variability (intraclass correlation coefficient = 0.43). This indicates 

that variability in individual exposure over time may be considerable. Indoor air levels explained only 

19% of the variance in prenatal urine TCPy (p = 0.001) in this study. However, these researchers also 
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demonstrated the presence of TCPy in meconium; meconium is fecal matter accumulated in intestines 

of developing fetus from week 13 of gestation and released immediately prior to or within a few days 

after delivery. This matrix is considered to be an integrated measure of exogenous exposures to the 

fetus during the gestational period. Importantly, investigators reported a moderate and significant 

correlation between maternal blood and urine collected in the later part of the pregnancy period, infant 

cord blood collected at delivery, and meconium (Whyatt et al., 2009). As noted in earlier sections, the 

critical windows of effect for these specific outcomes is unknown at this time, and may span the period 

of early pregnancy through early childhood (Rice & Barone, 2000; Rodier, 2004). The correlation among 

several biomarkers of chlorpyrifos exposure reflecting different windows of exposure from week 13 

through delivery (e.g., correlation between meconium and infant cord blood r=0.33, p=0.01, n=56), 

suggests a one-time measure of exposure may accurately rank participants with respect to prenatal 

exposures, at least within a 2- to 3-fold level of variation as suggested by authors (Whyatt et al., 2009). 

Additionally, the correlation between the analytic measure of exposure and a measure reflective of 

exposure during a large proportion of the gestational period suggests the exposure measures may also 

reflect exposure during the several likely critical windows of development in the mid- to latter period of 

gestation. However, to the extent participants experienced “peak” exposures at key periods of 

development, this information would not likely be reflected in the one-time measure, and likely lead to 

non-differential measurement error. 

Bias due to confounding occurs primarily when risk factors for the outcome are unequally distributed 

among exposure groups, but are not themselves caused by the exposure, and these factors are not 

controlled in either the design or the analysis of the study. The greatest concern is for unmeasured 

confounders, those variables not measured within the study that may be related to both the exposure 

and disease of interest but are unknown to the investigator. Residual confounding can also arise due to 

errors in the measurement or categorization of confounders, even after apparent adjustment for these 

factors in the study, i.e., these are factors not measured well enough. Among the major potential 

unmeasured or poorly measured confounding variables in the chlorpyrifos and neurodevelopmental 

association are: 1) difficult to measure aspects of the social environment (e.g., chronic stress, 

socioeconomic conditions), and 2) environmental co-exposures that may be determinants of 

neurodevelopment (e.g., lead exposure, mercury exposure, air pollution, tobacco smoke exposure, 

maternal alcohol intake, and exposure to other pesticides). The selection of cohorts that are 

homogenous with respect to many demographic and social factors significantly reduces the potential for 

residual or unmeasured confounding factors related to socio-economic status to have influenced these 

results as they have been controlled through restriction (another method of adjustment). For example, 

socioeconomic status cannot confound the main association in a cohort of individuals living under the 

same socioeconomic condition, i.e., there is little variation in the study population on this factor. 

However, close examination of descriptive tables in these publications indicates some degree of 

variability remains. Therefore, authors appropriately evaluated and included in final models several 

individual-level markers for socio-economic status such as education, income, and race. Additionally, 

within the Columbia Mother’s and Newborn study, authors performed new analyses using hierarchal 

regression techniques to model variability due to SES in the relation between prenatal chlorpyrifos 

exposure and mental and motor development (the Bayley Scale) (Lovasi et al., 2011). These techniques 

utilize individual- as well as group-level variables to model SES. In this research, authors did not observe 

a significant difference in risk estimates, indicating that the role of SES as a major confounder in the 

Columbia cohort is appropriately adjusted. However, others argue that the multi-dimensional nature of 

a characteristic such as SES can never be fully captured through commonly used variables, and suggest 

within a term such as SES there could be components that act as both positive and negative confounding 
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variables (Bellinger, 2011). However, at this time, there is no compelling evidence that lack of control of 

the confounding influence of SES significantly biased the reported risk estimates.

Additionally, authors have measured several other environmental chemicals and in some instances 

considered these other exposures as potential confounding variables (not always shown). Researchers 

collected information from participants regarding biomarkers of other organophosphates, 

environmental tobacco smoke (ETS), blood lead, PAH, methylmercury (Whyatt & Rauh, 2011), 

pyrethroids/pyrethrins, organochlorines such as DDT/DDE and hexachlorobenzene, PBDE, and 

polychlorinated biphenyls. While authors evaluated these environmental exposures in relation to both 

chlorpyrifos and outcomes, none were included as confounders in final models due to lack of statistical 

evidence of confounding. Within the Columbia Mothers and Newborn study, authors performed 

additional analyses suggested by the 2008 FIFRA SAP to evaluate the role of other organophosphate 

exposures, specifically diazinon and propoxur, in the relation between chlorpyrifos and both birth 

characteristics and mental and motor delays (Bayley scores) (Whyatt & Rauh, 2011). Chlorpyrifos and 

diazinon biomarkers were highly correlated in this cohort (r=63%). While the effect sizes reported for 

the reported relation between both birth length and also birth weight remained unchanged from 

published reports when diazinon was added to the model, significant differences were observed in the 

relation with neurodevelopmental functions (mental (MDI) an PDI (motor) scores). Specifically, authors 

reported a 50-200% increase in effect sizes with MDI and PDI, respectively, i.e., observed adverse effects 

became more pronounced with diazinon added to the model (See Appendix 4). Limited evidence of 

confounding by propoxur was also observed, but the correlation between chlorpyrifos and propoxur was 

moderate (23%), the confounding effect was not significant. In addition, neither pre- nor post-natal 

blood lead levels or methyl mercury levels measured in cord blood were significantly correlated with 

chlorpyrifos, and was therefore not considered a confounding variable in the association of interest 

within the Columbia cohort (Whyatt & Rauh, 2011; Rauh et al., 2006). 

However, uncertainties remain as to the role of other (unmeasured) environmental compounds as 

potential confounding variables in this association. None of the authors evaluated the possible effect 

modifying role of other environmental chemicals in the relation between chlorpyrifos and 

neurodevelopment, nor were studies suited to examining the effect of mixtures (simultaneous 

combined exposures at potentially relevant critical windows of development). New investigations are 

needed to address these hypotheses, if relevant. While the possibility of unknown, unmeasured positive 

confounding bias can never be completely excluded, given the evidence available it is unlikely this 

potential bias may entirely explain observed associations.

Selection bias occurs when study participants are either selected or lost to observation as a result of a 

third, unmeasured factor that is associated with both the exposure and outcome of interest. In these 

prospective cohort studies, selection bias is most likely induced not by selection into the study, but by 

selection out of the study, i.e. due to attrition of study participants and missing data. Because data are 

often not obtained on those that are lost to follow-up or for whom data is otherwise missing, it is 

difficult to determine whether or not, and to what extent, the observed associations are biased due 

these factors. The remedy is simple, in theory – follow the entire cohort and obtain all relevant data –

but often difficult in practice, particularly for long-term cohort studies such as those reviewed here. 

Across these studies, the amount of missing data varied but was great in some studies. To address 

missingness in data, authors imputed values in several instances, and performed sensitivity analyses 

with and without imputed data to illustrate comparability in reported results. In addition, , authors were 

able to illustrate differences between those included and excluded were comparable on several major 

characteristics, but some were not evaluated such as blood lead levels. Therefore, while it is difficult to 
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ascertain the degree to which selection bias due to missing data may have influenced these study 

results, authors employed appropriate analytic tools to address missingness in data to the extent 

possible. 

Information bias arises due to misclassification or error in the measurement of either the exposure or 

outcome of interest, or the accurate measurement of confounding variables. Qualities of the outcome 

ascertainment method, the exposure assessment method and the measurement of confounding 

variables were reviewed above. In summary, regarding exposure measurement error the true relevant 

exposure in these studies is prenatal chlorpyrifos to the developing fetus. Sources of exposure 

measurement errors include 1) the single measurement chlorpyrifos or its metabolites during the third 

trimester, 2) error arising from differences between measured biomarker levels and actual chlorpyrifos 

or chlorpyrifos oxon exposure, even at the one time point, 3) unmeasured time- and space- dependent 

patterns of chlorpyrifos exposure, 4) uncertainty regarding the critical period for chlorpyrifos effects on 

development, 5) missing exposure data, 6) laboratory errors, and 7) imputation of missing exposure 

levels. Measurement errors in the ascertainment of chlorpyrifos exposure are likely to have occurred in 

the studies reviewed. However, because of the prospective study designs employed, the errors are 

unlikely to result in falsely positive findings, because the probability and magnitude of these errors are 

likely to be independent of the outcome status of participants. Measurement errors in the 

ascertainment of the health outcomes were also likely to have occurred, perhaps more for the 

neurocognitive outcomes, than for the indicators of fetal growth and development, and also are unlikely 

to have spuriously positive findings. These measurement errors are more likely to have resulted in false 

negative findings, if a causal association truly exists. 

4.2.3.1 Epidemiologic Inference of Combined Children’s Health Cohorts

As stated earlier, the three children’s health cohorts considered herein have several strengths as well as 

limitations to consider in the interpretation of these studies. Within these studies, there are several 

factors that would tend to under-estimate the actual association (possibly leading to false negative 

association), as well as some characteristics that may lead to over-estimation (possibly leading to a false 

positive association). However, it must also be noted that methodological research and supplemental 

analyses (primarily within the Columbia Mothers and Newborn study) performed subsequent to the 

2008 FIFRA SAP deliberations concerning these children’s health cohort studies have reduced and 

further characterized important sources of uncertainty. As noted in the Introduction to this paper, 

currently scientists cannot determine with accuracy the critical window of exposure for these outcomes 

(early gestation through early childhood). Across the children’s health cohorts, researchers used one or 

two measure(s) of exposure to estimate gestational exposure during the critical window of 

development, and investigators assessed the main exposure using non-specific biomarkers of 

chlorpyrifos. Undoubtedly, exposure measurement error occurred. However, exposure validation 

studies illustrate some degree of correlation across exposure measures made at different periods of 

gestation (e.g., meconium, week 13 through delivery; maternal urine during last 8-weeks of pregnancy; 

and cord blood at delivery). Additionally, given the prospective nature of these studies, it is unlikely the 

measurement error was differential by outcome, i.e., non-differential exposure misclassification leading 

to biased estimates toward the null is anticipated. Finally, as noted above the degree of missingness in 

some key variables across these studies may have resulted in a form of selection bias which may have 

also lead to an under-estimation of effects. However, given the data are missing, it is difficult to assess 

the magnitude and direction of this error on study results. Low sample size may have limited the ability 

of researcher to identify sources of effect modification or perform stratified analyses with accuracy.
Page 71 of 124

Conversely, there may have been factors at play which led to an inflated estimate of the true 

association. Factor that may lead to false positive associations include unmeasured or poorly measured 

confounding factors which are positively associated with both chlorpyrifos use and adverse 

neurodevelopmental outcomes (positive confounding variables). The measurement of socio-economic 

factors and other environmental exposures experienced either pre- or post-natal environment are most 

likely among these potential factors, although others may exist. However, when additional adjustment 

was made for these factors using both individual- and group-level variables, the magnitude and direction 

of the associations remained stable (diazinon, propoxur, blood lead, methyl mercury, SES, post-natal 

exposure) (Lovasi et al., 2011; Rauh et al., 2006; Whyatt et al., 2004; Whyatt & Rauh, 2011). In the 

instances in which authors assessed the role of early childhood exposure to chlorpyrifos or other 

environmental contaminants as potentially confounding variables in the relation between prenatal 

exposure and neurodevelopmental outcomes, confounding bias was not observed. However, not all 

investigations evaluated the role of post-natal pesticide exposure, and authors note the limitation. 

Selection bias due to missing or drop outs from the study could also have influenced the observation of 

a false positive association, and this is again difficult to assess in the absence of data. However, 

researchers uniformly discussed and described the comparison of those included and excluded from 

individual analyses based in past of missing data using several important factors, and in many instances 

reported a level of comparability which is reassuring. The observation of positive associations as a result 

of multiple comparisons may also be a factor to consider; however the a priori identification of research 

questions of interest and the consistency of findings across several neurodevelopmental domains argues 

against a large role for multiple statistical testing to explain positive findings. 

Overall, these are well performed studies which are shielded from several major sources of bias in the 

interpretation of results due to the strong design, conduct and analyses utilized in these investigations. 

While factors are present across these studies which may have led to either false positive or negative 

associations, it is notable that positive associations were observed as EPA believes the possibility of 

under-estimation of effect size is more likely than factors that would lead to over-estimation of effect 

size. Authors have taken significant steps to address major sources of uncertainty that may lead to overestimation of effects, i.e., positive confounding bias as a result of poorly measured factors related to 

socio-economic status and other environmental chemical exposures. 

4.3 Chlorpyrifos Epidemiology Synthesis and Evaluation

In this chapter, EPA reviewed the results of epidemiological investigations of the association between 

prenatal chlorpyrifos exposure and adverse effects upon fetal growth and neonatal and early childhood 

neurodevelopmental outcomes across three major prospective children’s health cohort studies in the 

U.S. In accordance with the OPP’s draft “Framework for the Incorporation of Epidemiology into 

Regulatory Risk Assessment”21

 21 http://www.epa.gov/scipoly/sap/meetings/2010/020210meeting.html#materials

, including the use of the modified Bradford Hill considerations in judging 

the potential causal nature of observed associations, in this section EPA considers the totality of the 

epidemiological evidence from these cohorts. To perform this analysis, EPA considered the strength of 

the associations observed and the presence of exposure-response trends, the temporality of the 

observed associations, and the degree to which alternative explanations have been considered and 

eliminated as explanatory factors, among other considerations. Issues of biological plausibility and 

specific mechanisms of action which may explain a causal role for chlorpyrifos in adverse 
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neurodevelopmental outcomes, while broadly considered within the context of these investigations, are 

discussed in depth in accompanying chapters (See Sections 2, 3, 6). 

As noted previously, each investigation is a prospective cohort study in which prenatal exposures 

occurred prior to the developmental of either fetal growth anomalies or neonatal or early childhood 

neurological delays. Therefore, in each study exposure preceded the health effect, risk may be directly 

calculated, and measurement error is more likely non-differential in nature, i.e., likely under-estimation 

rather than over-estimation of risk effect. Table 10 summarizes the associations observed for 

chlorpyrifos markers and neurodevelopmental outcomes in the three cohorts. Within this database, 

moderately strong associations have been observed, and in many instances measured with precision. 

The relation between both neonatal neurological development as measured using the Brazelton index 

(number of abnormal reflexes) and also mental and psychomotor development among toddlers (Bayley 

Scale for 24-36 months) with prenatal chlorpyrifos exposure is elevated approximately 2- to 4- fold 

among those more highly exposed. This observation was consistent across cohorts with respect to the 

neonatal period, but not across measures using the Bayley Scale. Considering adverse effect in the 24-36 

months period of development, the effects became more pronounced over time in one cohort (the 

Columbia Mother’s and Newborn study), and became less pronounced over time in the Mt. Sinai Child 

Development study using non-specific biomarker of chlorpyrifos exposure (DAPs). However, it is notable 

the significant effects were seen when chlorpyrifos parent compound was measured directly; and, the 

effect size became significantly greater when diazinon was considered in the final statistical model. 

Notable decrements in intelligence measures were consistently observed across all three cohorts. While 

effect sizes measured as beta-coefficients are reflective of average change in intelligence score per unit 

change in chlorpyrifos exposure, and may appear modest in effect size, EPA notes that according to 

these models, the adverse influence at the extremes of the chlorpyrifos exposure distribution, if truly 

present in nature, would be more deleterious. Strong, consistent evidence of a positive association 

between prenatal chlorpyrifos and attentional problems, pervasive developmental disorder, and ADHDlike symptoms, although not measured with precision.

An association is consistent when a similar magnitude and direction of results are replicated in studies in 

different settings using different methods. If a relationship is causal, one may expect to observe it 

consistently in different studies and among different populations. Several consistencies were noted 

across these studies. Higher levels of the chlorpyrifos-specific metabolite TCPy were associated with 

prevalence of abnormal reflexes among newborns, as assessed using the Brazelton Neonatal Behavioral 

Assessment Scale (BNBAS), in both the Mount Sinai Hospital and UC Berkeley (CHAMACOS) Cohorts. The 

BNBAS assessment was not conducted in the Columbia Mothers and Newborns cohort. TCPy levels 

were consistently not associated with BNBAS indicators of newborn neurodevelopment such as 

behavioral domains other than those observed for number of abnormal reflexes. In addition, 

associations between pesticide markers and the Mental Development Index (MDI) portion of the Bayley 

Scores at 12, 24, and 36 months varied between small, largely non-statistically significant decrements, 

and null. However, in supplemental (post-publication) analyses, authors illustrated that adjustment for 

other organophosphate pesticides in the relation between prenatal chlorpyrifos and MDI strengthened 

the magnitude of the effect significantly at the 36 months time point (Whyatt & Rauh, 2011). 

Associations between early life organophosphate pesticide exposure biomarkers and decrements in 

mental development in childhood (age 7-9 years) were observed in all three studies, although the 

relationships were not consistently statistically significant.

Some notable inconsistencies were also observed. These may be explained by differing exposure 

biomarkers, timing of biomarker collection, or other limitations of study design and measurement, e.g., 
Page 73 of 124

the magnitude of non-differential exposure misclassification. However, inconsistency in results may also 

argue against a true causal association. Associations between markers of prenatal pesticide exposure

and fetal growth were not consistent across the three studies, as noted in the FIFRA SAP 2008 meeting. 

In the Mount Sinai Child Growth and Development cohort, no statistically significant associations were 

observed in primary analyses, although a modest decrease in average head circumference with 

increasing TCPy was noted among those with low PON1 activity. In the Columbia Mothers and Newborn 

Study (Whyatt et al., 2004), chlorpyrifos levels in maternal blood were modestly associated with 

decreased fetal growth. In contrast, among the UC Berkeley CHAMACOS study participants, modest 

increased birth weight and head circumference were associated with in utero DAP concentrations 

(Eskenazi et al., 2004; Harley et al., 2011). In fact, stratification by PON1 status seemed to enhance these 

positive associations, such that in the non-susceptible groups (i.e., PON1-108 CC, PON1192 RR, and high 

arylesterase activity), increasing DAP concentrations were significantly associated with increased birth 

weight and head circumference. The pesticide biomarkers differed between the two studies; cord blood 

chlorpyrifos levels were assessed in the Columbia Mothers and Newborn study, while DAPs were 

assessed in the UC Berkeley CHAMCOS cohort. It is possible that high urinary DAP concentrations may 

be an indication of rapid detoxification and excretion of organophosphate pesticides, rather than a 

marker of high exposure, particularly among the participants that are less potentially less susceptible to 

pesticide effects due to their PON1 genotype or PON activity level, as suggested by study authors.

Although consistency of associations across studies set in different populations, and employing different 

methods has been used as a criterion for causality, it is not the only compelling factor. In fact, 

heterogeneity across studies is expected to occur between studies in the presence of effect 

modification, if the prevalence of the effect modifier varies between the populations being assessed. 

Apparent inconsistencies could also be due to chance differences, and due to the presence of biases 

operating in the studies under consideration.

There was evidence for an exposure-response relationship in several of the studies reviewed, whereby 

increasing levels of exposure were associated with increasingly large decrements in measures of 

neurodevelopment (Rauh et al., 2011; Rauh et al., 2006). For the analyses in which continuous 

distributions of exposure markers were used, the levels were often, but not always, log transformed 

prior to entry in statistical models. The estimated association between the biomarker level and the 

outcome in these cases is not linear; rather, the association between the log-transformed exposure and 

the outcome is linear. In such cases, results are interpretable as changes in mean level/risk of the 

outcome of interest for a given percentage increase in the exposure. Examination of departures from 

linearity of exposure-response relationships was reported in only a small subset of the reviewed articles. 

For example, Rauh et al. (2011) reported that, in the Columbia cohort, “the dose-effect relationships 

between CPF [chlorpyrifos] exposure and log-transformed Working Memory Index and Full-Scale IQ 

scores are linear across the range of exposures in the study population, with no evidence for a 

threshold” (Rauh et al., 2011). Departures from linearity were not statistically significant. Two possible 

explanations for this finding are that 1) the exposure response is linear on the scale assessed, or that 2) 

the studies did not have sufficient power to detect departures from linearity in the shape exposureoutcome relationship. The smoothed exposure-response curve superimposed upon the scatter plot of 

log-transformed working memory appears to be monotonically decreasing; however, a similar curve 

superimposed upon the scatter plot of log-transformed IQ score and chlorpyrifos levels appears flat if 

not slightly positive, before turning negative beyond an inflection point at about 5 pg/gram of 

chlorpyrifos. As such, it may be that both explanations for the null finding regarding departures from 

linearity are operating in this study. As another example, Bouchard et al used cubic splines to evaluate 
Page 74 of 124

the shape of dose-response curves, test the linearity assumption, and investigate potential thresholds 

(Bouchard et al., 2011). Again, no statistically significant departures from linearity were observed. 

Due to the voluntary cancellation by registrants of indoor residential uses of chlorpyrifos-containing 

pesticide products, children born after the year 2000 were likely exposed to far lower levels of 

chlorpyrifos, on average, than children born prior to 2001. In their report, Whyatt et al. (2004) took 

advantage of this “natural experiment” by stratifying their analyses on birth date prior to, or after, 

January 1, 2001. Although reported household use of pesticides in general did not change over the 

same time period in this study, maternal chlorpyrifos in maternal air samples were significantly lower 

after the voluntary cancellation (4.9 ng/m3

), relative to those taken before 2001 (8 ng/m3; p<0001). 

Associations between birth outcomes and organophosphate insecticide levels in maternal personal air 

samples were not statistically significant in both un-stratified analyses and among subgroups stratified 

by birth prior to or after January, 2001. However, the associations between birth weight and length and 

cord plasma chlorpyrifos were highly significant (p ≤ 0.007) among newborns born before the 2001 

cancellation of registrants. Among newborns born after January 2001, no association with fetal growth 

was observed (p > 0.8) (Whyatt et al., 2004). Supplemental analysis suggested this observation persisted

among additional number of infants born after the period of the voluntary cancellation (n=193) (Whyatt 

& Rauh, 2011). This finding by Whyatt et al provides evidence that the association “can be altered 

(prevented or ameliorated) by an appropriate experimental regimen,” as advocated by AB Hill in the 

assessment of causality in observational studies. Changes in exposure were reflected in decreasing 

chlorpyrifos levels, and appear to have been associated with changes in developmental outcomes.

However, as noted by the 2008 FIFRA SAP Panel in their report, this study was not designed to test this 

hypothesis, specifically.

It must also be considered that given the many selection criteria applied within each of the three 

children’s health cohorts, the generalizability of study results may have been affected. These studies 

reflect a somewhat narrowly defined sub-population who likely experienced a greater range of 

exposures and higher peak exposures than the general U.S. population; a comparison of maternal 

urinary concentration of TCPy in the three children’s health cohorts with the general population 

(NHANES) illustrates higher exposures occurred among the cohort participants. The exception is with the 

urinary concentration data collected by Whyatt et al. (2009) which were lower than the U.S. general 

population; however, these data were collected 2001-2004, a period after the voluntary cancellation of 

indoor residential uses of chlorpyrifos (see Section 5.0). In addition, the external validity of the study 

findings may be limited if the organophosphate pesticide exposure-fetal development association is 

modified by factors that are more, or less, prevalent in the study populations relative to the 

population(s) to which inference is being made, or to populations with a substantially different exposure 

range. In the studies reviewed, the exploration of gene-by-environment, and phenotype-byenvironment interactions is a strength of many of the analyses, although as noted above the studies are 

uniformly underpowered to assess such interactions. There was scant assessment of interaction 

between chlorpyrifos and other environmental co-exposures conducted in these investigations including 

blood lead levels.
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4.4 Conclusions

The FIFRA SAP panel convened in September 2008 concluded that given the strengths of these studies, 

the effect of biases likely present, and rigorous statistical analyses performed, the observation of 

associations across these studies can be interpreted as a conclusion that chlorpyrifos likely played a role 

in the neurodevelopmental outcomes observed among children more highly exposed to chlorpyrifos in 

the in utero environment. Since that time, researchers with each of the three children’s health cohorts 

have published several new etiologic and methodological studies and performed supplemental analyses 

as suggested by the 2008 Panel which extend the knowledge base, and reduce several major sources of 

uncertainty present in the 2008 database (i.e., confounding by SES, exposure validation in the Columbia

Mothers and Newborn study, further evaluation of PON1 genotype and PON activity level). In the 

current document, EPA utilized the draft “Framework for Incorporating Epidemiology into Risk 

Assessment”22, including use of modified Bradford Hill considerations, to assess the strengths and 

limitations of the studies in this database on the association of interest, and to clarify the epidemiologic 

evidence in support of a causal role for chlorpyrifos in infant and child adverse neurodevelopmental 

outcomes. 

In EPA’s analysis of the strengths and weaknesses of these studies, the chance that positive associations 

observed are false positives due to systematic errors in the studies cannot be excluded; however, it is 

more likely that error present in these studies would lead to the under-estimation of the true 

association. Therefore, while alternative explanations for positive association can be hypothesized (e.g., 

additional unmeasured or poorly measured positive confounding variables), these explanations are 

judged to be less plausible than the alternative that associations have been missed or under-estimated 

due to non-differential measurement error and low sample size across exposure strata. In occupational 

settings, exposure measurement error has been shown to more greatly influence epidemiology study 

results than unknown or unmeasured confounding variables (Blair et al., 2011). Additionally, the 

elimination of temporal bias due to the prospective study design employed within each of the three 

children’s health cohorts assures that prenatal exposures preceded neurodevelopmental outcomes 

measured at birth, and in early and later childhood through age 7 years. The strength of the associations 

measured in some studies was notably strong. However, associations in many instances were weak to 

moderate, possibly due to measurement error. Associations with neurodevelopmental outcomes were 

consistently identified with respect to the number of abnormal reflexes in the neonatal period, the 

presence of mental and behavioral issues as well as gross motor delays were pronounced especially in 

later toddler years of 24-36 months, and the observation of intelligence decrements were seen across 

the three cohorts using different measures of prenatal chlorpyrifos exposure, although not consistently 

statistically significant. EPA notes that other organophosphates may be involved as well as the 

CHAMACOS and Mt. Sinai Child Development study utilized DAP which measure exposure to several 

different organophosphates, and that use of chlorpyrifos in the CHAMACOS study region was low at the 

time of the study. However, links with DEP which result from either chlorpyrifos or diazinon exposure 

were observed as well. In general, given that the intended mode of action of chlorpyrifos toxicity is 

neurological dysfunction, i.e., inhibition of acetyl cholinesterase, a biologically plausible role for 

chlorpyrifos in adverse neurodevelopmental outcomes in the developing brain can be posited. However, 

this issue is discussed in greater detail elsewhere in this paper. 

In summary, while the strengths and limitations of the studies would be more likely to lead to an underestimation of the true effect, the possibility of false positive associations cannot be entirely ruled out. 

 22 http://www.epa.gov/scipoly/sap/meetings/2010/020210meeting.html#materials
Page 76 of 124

The temporal association, strength of the statistical associations observed and presence of exposure 

response across some (but not all) of the investigations, as well as some notable consistencies in 

findings within this epidemiologic database tend to support a role for chlorpyrifos in this relation. It is 

true that the Columbia Mothers and Newborn study is the only cohort to have measured chlorpyrifos 

parent compound directly; replication in other cohorts that use this exposure metric would aid causal 

inference. The presence of factors which may have over-estimated effects, the lack of consistency in 

many neurodevelopmental domains, and the lack of a clear mechanism of action may argue against a 

true association. Overall, the current database supports the conclusion from 2008 Panel that

“chlorpyrifos likely played a role in the birth and developmental outcomes noted in the three cohort 

studies.” Additionally, the subsequent studies made available since the time of the 2008 evaluation 

helps to clarify the potential role for chlorpyrifos in adverse neurodevelopmental health effects. 
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5.0 Exposure Profile & Biomonitoring Research

In the previous Section, the Agency performed a targeted evaluation of recent epidemiologic results 

from the Columbia Mother’s and Newborn Study, Mt. Sinai Child Growth and Development study, and 

CHAMACOS children’s health cohorts. Each of these prospective cohorts utilized exposure assessment 

methodologies based on observed levels of chlorpyrifos or metabolites in environmental media and 

biological samples; however, only in the Columbia Mothers and Newborn study were etiologic analyses 

were all performed using biological markers of chlorpyrifos exposure. 

In this section, EPA presents a comparison of biomonitoring studies involving the general population and 

specific sub-populations, including children, farm workers, and pregnant women. The purpose of the 

comparison is to establish a frame of reference to help characterize and evaluate the range of exposures 

to chlorpyrifos observed in each of the children’s health cohorts. Specifically, TCPy biomonitoring results 

from the three children’s health cohorts are compared with other biomonitoring results from other 

major observational exposure studies in the United States. Section 5.1 first provides an overview of the 

main chlorpyrifos biomarkers and important scientific considerations that should be evaluated when 

interpreting biomonitoring results. Section 5.2 then provides a comparison of the children’s health 

cohorts with other biomonitoring studies involving pesticide applicators, farm families, children, and the 

general population, which expands the Agency’s biomonitoring review presented during the September 

2008 SAP on chlorpyrifos. Section 5.3 updates the Agency’s previous review of the association between 

chlorpyrifos biomonitoring data and AChE levels. Finally, Section 5.4 discusses various approaches for 

interpreting biomonitoring data and provides a discussion of PBPK modeling efforts for chlorpyrifos.

5.1. Overview of Biomarkers and Scientific Considerations

A summary of biomarkers of chlorpyrifos exposure are summarized in Table 11 below. The biomarkers 

most widely monitored to evaluate exposure to chlorpyrifos are the urinary metabolites TCPy and the 

non-specific DAP metabolites DEP and DETP. In addition to these urinary biomarkers of exposure, 

chlorpyrifos can also be measured directly in blood and other biologic media.

Table 11 Common Chlorpyrifos Biomarkers of Exposure. 

Biomarker/Analyte Biologic Matrix Population Group

3,5,6-trichloro-2-pyridinol 

(TCPy)

Urine All

Meconium Pregnant Females, Neonates

Diethylphosphate (DEP) Urine All

Diethylthiophosphate (DETP) Urine All

Chlorpyrifos Blood All

Cord Blood Pregnant Females, Neonates

The National Academy of Sciences established a biomonitoring framework that recommends assessing 

biomonitoring data quality based on scientific and interpretive considerations that include: biomarker 

specificity; pharmacokinetics and intra-individual variability; and study design considerations (NAS, 

2006). The remainder of this section provides an evaluation of these scientific and interpretative

considerations with respect to the primary urinary and blood biomarkers that have been used to 

evaluate chlorpyrifos exposure in exposure studies and epidemiologic research. Several review articles 
Page 78 of 124

in the peer-reviewed literature have attempted to characterize the data quality of chlorpyrifos 

biomonitoring. Barr and Angerer (2006), in particular, provide a detailed review of chlorpyrifos 

biomarkers that discusses several of the considerations outlined by NAS. Similarly, Eaton et al. (2008) 

provides discussion of chlorpyrifos biomarkers in their extensive review of the toxicology of chlorpyrifos. 

Likewise, Egeghy et al. (2011) of the Agency’s Office of Research and Development (ORD) reviewed 

pesticide urinary biomarkers from ORD observation exposure studies with an emphasis on 

pharmacokinetic factors. 

Urinary Biomarkers of Exposure

As previously discussed in Section 4, TCPy, DEP, and DETP are commonly used as urinary biomarkers of 

chlorpyrifos to exposure in etiologic research. DEP and DETP are considered non-specific biomarkers 

because they are metabolites of multiple of OP pesticides, including chlorpyrifos-methyl, coumaphos, 

diazinon, disulfoton, ethion, parathion, phorate, sultotepp, and terbufos (CDC, 2009). As such, it is more 

difficult to attribute urinary levels of DEP and DETP to chlorpyrifos exposure without additional 

evaluation of exposure routes involving all potential OP parent compounds. 

In contrast, the main parent compounds of TCPy are chlorpyrifos, chlorpyrifos-methyl, and trichlopyr, so 

urinary TCPy levels are attributable to exposure to fewer parent compounds that the DAP metabolites. 

For this reason, TCPy is generally considered a more specific biomarker of exposure than the DAP 

metabolites. However, studies have demonstrated that chlorpyrifos degrades into TCPy in 

environmental media and food. As such, urinary TCPy concentrations may reflect exposure to the 

parent compound chlorpyrifos and/or to its degradate, TCPy, in the environment. Morgan et al. (2005, 

2011) and Wilson et al. (2003) have observed that TCPy concentrations were at least 10 times higher 

than chlorpyrifos concentrations in 48-hour duplicate diet samples (Eaton et al., 2008). Similarly, DEP, 

DETP, and the other non-specific DAPs have also been detected in the environmental media and foods. 

For example, Lu et al. (2005) observed that OP pesticides in fortified fruit juice samples degraded into 

DAPs. Based on these findings, and as noted by epidemiology study authors, use of urinary 

measurements of TCPy (or DEP and DETP) to assess chlorpyrifos exposure may overestimate chlorpyrifos 

exposure if it is assumed that exposure to the parent compound is the only source of the urinary 

biomarkers (Eaton et al., 2008). It is not clear, however, the extent to which environmental exposure to 

the metabolites contribute to urinary concentration of DAPs or TCPy. Thus it is important to consider the 

relative contribution of exposure to the parent chlorpyrifos compound and the environmental 

metabolite when interpreting epidemiology results.

Longitudinal variability is another important consideration when interpreting chlorpyrifos biomonitoring 

results. Studies evaluating chlorpyrifos exposure typically measure TCPy or DAP levels through 

collection of a single spot urine samples, 24-hour voids, or repeated samples over an extended period of 

time. Spot urine samples generally provide less reliable estimates of exposure if the goal of the study is 

to estimate the periodicity of exposure over an extended period of time (Barr and Angerer, 2006); 

however, if exposure is relatively constant, a spot urine sample may accurately reflect chronic exposure. 

Barr and Angerer (2006) suggest that previous studies indicate that spot-urine samples cannot 

accurately quantify pesticide exposure, but also indicate that more recent studies have shown that a 

single sample may be adequate for assigning broad exposure classifications in epidemiology studies. This 

conclusion was supported by methodological research performed by investigators with children’s health 

cohort studies. Specifically, epidemiology investigators have illustrated some correlation among 

biomarkers reflecting short and long exposure durations, suggesting a one-time exposure measure may 

accurately reflect a longer period of exposure (Whyatt et al. 2009).
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Blood Biomarkers of Exposure

Monitoring chlorpyrifos directly in blood can reduce the uncertainty associated with urinary biomarkers 

because blood measurement provides a more direct measure of internal dose that is specific to 

chlorpyrifos (Barr and Angerer, 2006). While monitoring of chlorpyrifos in blood can provide a better 

measure of exposure in some cases, blood measurement can place a greater burden on study 

participants and require more sensitive analytical methods. Additionally, chlorpyrifos has a short 

biological half-life (<27 hours), therefore, blood concentrations of chlorpyrifos may represent a much 

narrower exposure period relative to urinary concentrations (NAS, 2005). These challenges may explain 

why researchers have typically monitored TCPy and other urinary biomarkers to evaluate OP exposure. 

While collecting and monitoring blood samples for chlorpyrifos may be challenging, collection of 

umbilical cord blood can be more feasible because larger quantities of blood can be collected without 

invasive collection procedures (Barr and Angerer, 2006). 

It appears that only a limited number of biomonitoring studies in the United States have measured 

chlorpyrifos in blood, perhaps for the reasons suggested by Barr and Angerer (2006). Two of these 

studies have previously been reviewed by the Agency during its 2008 SAP on chlorpyrifos (USEPA, 2008) 

and involved administration of a dose of chlorpyrifos to adult volunteers in order to examine the 

association between chlorpyrifos blood levels and AChE activity (Nolan et al., 1982; Kisicki et al., 1999).23 

Additionally, one study by Whyatt et al. (2003), also previously reviewed by the Agency, measured 

chlorpyrifos in both maternal and cord blood of the Columbia Mother’s and Newborn study. 

5.2 Chlorpyrifos Biomonitoring Study Comparison

This section presents a summary comparison of selected TCPy urinary biomonitoring studies in order to 

help characterize the range of potential exposures in the both the general U.S. population and study 

populations that have been targeted by researchers. The comparison focuses on urinary TCPy because 

it has been widely monitored in a range of different studies and has greater specificity to chlorpyrifos 

than the DAP analytes. Urinary TCPy concentrations are commonly reported on a volume and/or a 

creatinine-adjusted basis. Creatinine is a common correction method to adjust for variable dilutions in 

urine samples. However, there can be considerable variation in creatinine concentrations in urine 

amongst different age and ethnic populations, due partially to differences in lean muscle mass (Barr et 

al., 2005). For example, children generally have less muscle than active adults and African-Americans 

generally have more muscle mass than Caucasians. As a result, interpretation of creatinine-adjusted 

concentrations requires additional adjustment of typical creatinine urinary levels by age, sex, and race. 

Additionally, it has also been suggested that creatinine adjustment may not improve the interpretation 

of biomonitoring when urinary creatinine is highly variable (Kissel et al., 2005). For these above 

reasons, only TCPy concentrations on urine volume basis (μg/L) are presented in this section.

The studies selected are not the entire universe of all available biomonitoring studies. Rather, the 

majority of selected studies have previously been evaluated as part of biomonitoring review conducted 

as part of the Agency’s 2008 SAP on the toxicity profile of chlorpyrifos (USEPA, 2008), reviewed by Eaton 

et al. (2008), or discussed in more detail in Section 4 of this report. In addition to these studies, a 

 23 EPA has determined that it will not rely on Kisicki et al. (1999) because it was intentional human 

chlorpyrifos dosing study involving children less than 18 years old.
Page 80 of 124

literature review was performed to identify relevant studies that were published following the Agency’s 

previous 2008 SAP.

While the selected studies do not represent the entire universe of available studies, they are generally 

considered the major monitoring and epidemiologic studies that have monitored exposure to 

chlorpyrifos. CDC’s National Health and Nutrition Examination Survey (NHANES) provides reference 

ranges of TCPy urinary levels in the general population during the 1999-2000 and 2001-2002 survey 

cycles (CDC, 2009).24 Similarly, EPA’s Office of Research and Development (ORD) has conducted or 

funded a number of observational studies to investigate children’s exposure to pesticides and other 

contaminants, including: National Human Exposure Assessment Survey (NHEXAS); Minnesota Children’s 

Pesticide Exposure Study (MSCPES); Children’s Total Exposure to Persistent Pesticides and Other 

Persistent Organic Pollutants Study (CTEPP); Biological and Environmental Monitoring for 

Organophosphate and Pyrethroid Pesticide Exposures in Children Living in Jacksonville, Florida Study 

(JAX); and Children’s Pesticide Post-Application Exposure Study (CPPAES) (Egeghy et al., 2011). The 

sample collection period of these studies ranged from 1995-1997 (NHEXAS) to 2001 (CTEPP), 

representing time periods when chlorpyrifos was registered for residential applications and the start of 

the phase-out. 

Dow AgroSciences, the primary registrant of chlorpyrifos products, has also submitted a number of 

observational studies to the Agency that involved biomonitoring of TCPy following application of 

chlorpyrifos by adults in occupational and residential use scenarios. The majority of these studies have 

been incorporated into the Agency’s assessment of occupational and residential human health risk 

assessments for the Chlorpyrifos Reregistration Eligibility Decision (EPA, 2006) and, more recently, the 

2011 updated Occupational and Residential Exposure Assessment (EPA, 2011]. Dow’s studies were 

intended to quantify environmental levels and exposures to chlorpyrifos following pesticide application 

in occupational and residential scenarios. For the purposes of comparison, a subset of Dow 

biomonitoring studies is presented in the section (MRID 44458201, MRID 43013501, MRID 43062701, 

MRID 43138102). The subset of studies was selected because it represents the high and low range of 

TCPy urinary levels observed in all the occupational and residential use scenarios evaluated by Dow. 

In addition to the NHANES, ORD research, and registrant studies described above, a number of 

biomonitoring studies help further characterize the range of exposures in different population groups 

and the role of different exposure pathways. Several of the studies have previously reviewed during the 

Agency’s 2008 SAP (EPA, 2008). Eaton et al. (2008) has also performed a review of exposure studies that

have measured urinary TCPy levels. The studies reviewed include several that have measured to TCPy 

levels to investigate differences in pesticide exposure in farm and non-farm communities (Curwin et al., 

2005, 2007; Alexander et al., 2006; Fenske et al., 2002). These studies help characterize the role of takehome exposure in agricultural communities. Similarly, biomonitoring studies have been conducted to 

investigate the role of dietary exposure (Lu et al., 2006, 2008). In these studies, children’s conventional 

diets were replaced with organic foods to determine if an organic diet could lower children’s dietary 

exposure to OP pesticides. 

 24 CDC has not publically released TCPy urinary data from more recent NHANES survey cycles.
Page 81 of 124

Comparison of Study Results

The monitoring studies described above can help characterize the exposure levels in the CHAMACOS, 

Columbia University, and Mt. Sinai cohorts described previously in Chapter 3. Specific studies from 

these cohorts include Eskenazi et al. (2007) and Castorina et al. (2010) from the CHAMACOS cohort, 

Whyatt et al. (2009) from the Columbia University cohort, and Berkowitz et al. (2003) from the Mount 

Sinai cohort. Based on the studies described above, comparisons of TCPy biomonitoring results for adult 

and child study populations are provided in Figure 4 and Figure 5, respectively. These figures provide 

information on the range of observed TCPy levels based on summary results reported by the 

investigators. As such, traditional boxplots could not be generated independently because raw study 

data from individual studies were not available. Whenever possible, boxplots are presented in which 

the box length indicates 25th and 75th percentiles, central line indicates median, and whiskers indicate 5th

and 95th percentiles. However, several studies only reported central tendency values and the range of 

observed values. In these cases, study summary data are represented by a line that indicates the central 

tendency value (median unless noted) and whiskers indicating the range of observed values. In 

addition, the horizontal axis of each figure provides the study reference and description of study 

population. As part of the study description, the type of population is described, often based upon

Table 17 of Eaton et al. (2008), along with the year of sample collection and sample size. 

Child Study Populations

Figure 4 first provides a general reference range of urinary TCPy levels of 6 to 12 year old children in the 

general U.S. population based on NHANES 1999-2000 and 2001-2002, respectively (CDC, 2009). Based 

on comparison of these survey periods, 50th percentile TCPy levels were 2.8 and 3.1 μg/L, respectively. 

Likewise, 95th percentile urinary TCPy levels were fairly similar during the two survey periods and were 

16.0 and 15.3 μg/L, respectively. 

As has been previously described by Egeghy et al. (2011), median urinary TCPy levels were fairly similar 

across ORD monitoring studies of children and ranged from 5.1 μg/L in CTEPP to 12.0 μg/L in NHEXASAZ. These median levels were roughly 2-4 times higher than the median TCPy levels of 6-12 years old 

children in the 1999-2000 and 2001-2002 survey periods of NHANES. Comparison of studies that 

monitored children living in farm families or involved in farm activities (Alexander et al., 2005; Fenske et 

al., 2002), yielded results that were more similar to NHANES, although these studies were smaller in 

scale, making it more difficult to make direct comparisons.

The results of longitudinal studies examining the contribution of dietary pesticide exposure (Lu et al., 

2006, 2008) were also comparable to the general range of TCPy levels observed in 6-11 year old children 

in NHANES. While both of these studies were comparable to NHANES, both studies also reported 

significant decreases in TCPy levels in their cohort following replacement of diet with organic foods. Lu 

et al. (2003) observed that median TCPy levels decreased from 7.2 μg/L during the initial conventional 

diet phase of the study to 1.7 μg/L during the subsequent organic diet phase. Following the reintroduction of conventional foods, the median TCPy level in children increased to 5.8 μg/L. Similarly, 

while not represented by the data presented in Figure 4, Lu et al. (2008) had study participants switch to 

organic diets during part of the summer and fall periods of their study and reported similar trends. 
Page 82 of 124

Adult Study Populations

As with the previous figure, Figure 5 first provides a general reference range of TCPy levels in the general 

adult U.S. population based on NHANES 1988-1992,1999-2000 and 2001-2002, respectively (Hill et al., 

1995; CDC, 2009). Based on comparison of these three different survey periods, 50th percentile TCPy 

levels were 3, 1.5, and 1.9 μg/L, respectively. While this suggests that 50th percentile concentration may 

have declined relative to 1988-1992, the 95th percentile TCPy levels were 13, 9.9, and 10.9 μg/L, 

respectfully. As such, there appear to be smaller differences at the upper end of the distribution of TCPy 

levels in adults during the three survey periods.

Comparison of the range of TCPy levels observed in NHANES with other monitoring yielded mixed 

results. In studies that specifically monitored farmers or their families, some studies reported TCPy 

levels in adults that were within the range observed in NHANES (Curwin et al., 2005; Alexander et al., 

2005), whereas Curwin et al. (2007) observed geometric mean TCPy levels in adults that were higher 

than the 95th percentiles in all three NHANES survey periods. On the other hand, TCPy levels in the 

registrant-submitted worker and residential applicator studies were generally higher than the range 

observed in NHANES. This was true for registrant studies that represented the both the lowest and 

highest TCPy levels in all eleven monitoring studies that have been submitted to the Agency.

TCPy levels of mothers that were part of the CHAMACOS, Columbia, and Mount Sinai cohorts also varied 

in comparison with the general range of TCPy levels observed in NHANES. Eskenazi et al. (2007) from 

the CHAMACOS cohort observed median levels that were higher than all NHANES adults aged 20-59 

years old. Similarly, Castorina et al. (2010) performed a subanalysis of NHANES, which suggests that 

median TCPy levels observed in prenatal spot urine samples were slightly higher than the median levels 

TCPy levels of pregnant women in NHANES 2001-2002. Berkowitz et al. (2003) of Mount Sinai also 

observed median TCPy levels that were higher than NHANES. In particular, Berkowitz et al. observed the 

median TCPy level of their cohort was 7.6 μg/L, which was higher than the median TCPy levels observed 

in NHANES during the three survey periods. On the other hand, the TCPy levels reported by Whyatt et 

al. (2009) of the Columbia cohort appear to be lower than the TCPy levels observed in adults in the 

general U.S. population. Specifically, Whyatt et al. (2009) reported median TCPy levels that ranged from 

<LOD (0.26 μg/L) to 0.61 μg/L during pre/postnatal sample collection; however EPA notes that in 

contrast to the other cohorts mentioned these data were collected, in part, during a time period of rapid 

decline in use of chlorpyrifos (subsequent to the voluntary cancellation of indoor residential uses).
Page 83 of 124

Figure 4 Summary of Comparison of TCPy Biomonitoring Studies Involving Child Study Populations.a

a Boxplots are presented in which the box length indicates 25th and 75th percentiles, central line indicates median, and whiskers indicate 5th and 95th percentiles. However, 

several studies only reported central tendency values and the range of observed values. In these cases, study summary data are represented by a line that indicates the central 

tendency value (median unless noted) and whiskers indicating the range of observed values.

GM – Geometric Mean

0.1

1

10

100

1000

NHANES 1999-2000 (n=481)NHANES 2001-2002 (n=573)

NHEXAS-AZ 1995-97, 5-12 years (n= 21)

MNCPES 1997, 3-13 years (n = 263)

CTEPP-NC 2000-2001, 2-5 years (n=129)

CTEPP-OH 2001, 2-5 years (n=123)

JAX 2001, 4-6 years (n = 9)

CPPAES 1999-2000, 2-4 (n=81)

CPES-WA, Summer 2003-2004 (n= 21) 

CPES-WA, Fall 2003-2004 (n= 21) 

CPES-WA, Winter 2003-2004 (n= 21) 

CPES-WA, Spring 2003-2004 (n= 21) 

Conventional Diet, Pre, 2003 (n=87)

Organic Diet, 2003 (n=116)

Conventional Diet, Post, 2003 (n=155)

Child 2001, farm (n=65)

Child 2001, nonfarm (n=51

Farm families

—pre-appl 2000-2001 (n=50) 

Farm families

—post-appl 2000-2001 (n=50)

CDC, 2009 (6-11 

years)

Egeghy et al, 2011 Lu et al., 2008 (3-11 years) Lu et al., 2006 (3-11 years) Curwin et al, 2007 Alexander et al., 2006 

(4-18 years) 

TCPy Urinary Concentration (μg/L)

Reference 

Range

GM GM
Page 84 of 124

Figure 5 Summary of Comparison of TCPy Biomonitoring Studies Involving Adult Study Populations.a

a Boxplots are presented in which the box length indicates 25th and 75th percentiles, central line indicates median, and whiskers indicate 5th and 95th percentiles. However, 

several studies only reported central tendency values and the range of observed values. In these cases, study summary data are represented by a line that indicates the central 

tendency value (median unless noted) and whiskers indicating the range of observed values.

GM – Geometric Mean

0.1

1

10

100

1000

NHANES III, 1988-1992 (n= 993)

NHANES 1999-2000 (n=832)

NHANES 2001-2002 (n=1,113)

Residential Low- Crack & Crevice/Spot (n=)

Residential High- Turf Treatment (n=)

Worker Low- Re-Entry (n=)

Worker High- Mix/Load/Apply/Clean (n=)

Nonfarmers, chlorpyrifos not sprayed (n= 45)

Farmers, chlorpyrifos not sprayed (n=45)

Farmers, chlorpyrifos self-sprayed (n=2)

Father, nonfarm (n=23)

Father, farm (n=24)

Mother, nonfarm (n=24)

Mother, farm (n=24)

Farm families

—Applicators pre-appl. (n=34)

Farm families, Applicators post-appl (n=34

Farm families, Spouses pre-appl 

Farm families, Spouses post-appl (n=34)

Mother–child pair cohort 1998-2001 (n=365)

Prenatal 1 (n=95)Prenatal 2 (n=83)Prenatal 3 (n=54)Prenatal 4 (n=21)

Postnatal (n=73)

1st Pregnancy (n=442)

2nd Pregnancy (419)

Prenatal 1 (n=538)Prenatal 2 (n=481)

NHANES Subsample 1999-2002 (n=224)

Hill et al, 1995; CDC, 

2009

Dow AgroSciences Curwin et al, 2005 Curwin et al, 2007 Alexander et al., 2006 Berkowitz et 

al., 2003

Whyatt et al, 2009 Eskenazi et 

al, 2007

Castorina et al, 2010 

TCPy Urinary Concentration (μg/L)

Reference 

Range

Maternal Cohorts

GM

GM

GM

GM

GM GM

GM
Page 85 of 124

Table 12 Summary Results of TCPy Biomonitoring Studies (Modified from Eaton et al., 2008)

Reference Cohort Type of Study 

Population Collection Period Collection Method n % Det.

Urinary TCPy Concentration (μg/L)

LOD AM GM Min 5th 25th 50th 75th 95th Max

Hill et al.,

1995 NHANES III 20-59 year olds 1988-1992 Spot 993 82 1 4.5 - - <LOD 1.3 3 5.9 13 77

CDC, 2009

NHANES 6-11 year olds 1999-2000 Spot 481 - 0.4 - 2.88 - 0.56 1.3 2.8 7.09 16 -

NHANES 6-11 year olds 2001-2002 Spot 573 - 0.4 - 2.67 - <LOD 1.1 3.09 6.36 15.3 -

NHANES 20-59 year olds 1999-2000 Spot 832 - 0.4 - 1.53 - <LOD 0.76 1.5 2.9 8.9 -

NHANES 20-59 year olds 2001-2002 Spot 1,113 - 0.4 - 1.91 - <LOD <LOD 1.91 4.44 10.9 -

Egeghy et al.

2011

ORD - NHEXES, AZ 5-12 year olds 1995-1997 Spot – Morning Void 21 100 - 12 9.3 2 - 5.7 12 14 26 30

ORD - MNCPES 3-13 year olds 1997 Spot – Morning Void 263 92 - 9.2 6.6 <1.4 - 4 7.2 12 23 45

ORD - CTEPP,NC 2-5 year olds 2000-2001 Spot – Three Voids 129 98 - 7.5 5.5 <1.0 - 3.8 5.3 8.4 16 100

ORD - CTEPP, OH 2-5 year olds 2001 Spot – Three Voids 123 100 - 5.9 4.9 1.2 - 3.1 5.2 7.8 12 15

ORD - JAX 4-6 year olds 2001 Spot – Morning Void 9 100 - 11 9.1 2.9 - 7.5 9.8 15 21 21

ORD - CPPAES 2-4 year olds 1999-2000 Spot – Morning Void 81 93 - 8 6.4 <1.0 - 4.5 7.7 11 18 20

Lu et al., 2008 

CPES-WA 3-11 year olds 2003, Summer Spot - Longitudinal 23 86 0.2 6.4 - - 0 0 2 4.8 9.4 -

CPES-WA 3-11 year olds 2003, Fall Spot - Longitudinal 21 88 0.2 2.6 - - 0 0 0.6 1.7 3.7 -

CPES-WA 3-11 year olds 2004, Winter Spot - Longitudinal 20 96 0.2 5.1 - - 0.2 0.6 1.7 3.5 7.7 -

CPES-WA 3-11 year olds 2004, Spring Spot - Longitudinal 19 97 0.2 5.6 - - 0.4 1.3 2.5 4.6 8 -

Lu et al., 2006 

WA Children 3-11 year olds 2003 Spot - Longitudinal 87 78 0.2 7.2 - <LOD - - 6 - - 31.1

WA Children 3-11 year olds 2003 Spot - Longitudinal 116 50 0.2 1.7 - <LOD - - 0.9 - - 17.1

WA Children 3-11 year olds 2003 Spot - Longitudinal 155 78 0.2 5.8 - <LOD - - 4.3 - - 25.3

Alexander et 

al., 2006 

Farm families—pre 

appl 4-18 year olds 2000-2001 Spot 50 - - - 7.6 1.9 - - 6.5 - - 31.5

Farm families—postappl 4-18 year olds 2000-2001 Spot 50 - - - 7.6 1.1 - - 6.7 - - 77.1

Farm families—pre 

appl Applicators 2000-2001 Spot 34 - - - 7.8 1.6 - - 6.6 - - 44.1

Farm families—postappl Applicators 2000-2001 Spot 34 - - - 19 4.1 - - 15.3 - - 293

Farm families—pre 

appl Spouses 2000-2001 Spot 34 - - - 4.7 1.6 - - 4.4 - - 24.1

Farm families—postappl Spouses 2000-2001 Spot 34 - - - 5 1.2 - - 4.7 - - 34.7

Fenske et al., 

2002

Applicator children ≤ 6 years olds 1995 Spot 49 20 8 - - <LOD - - - - - 100

Farm worker children ≤ 6 years olds 1995 Spot 12 33 8 - - <LOD - - - - - 53

Agricultural children ≤ 6 years olds 1995 Spot 61 23 8 - - <LOD - - - - - 100

Reference children ≤ 6 years olds 1995 Spot 14 29 8 - - <LOD - - - - - 27

Curwin et al., 

2005

Nonfarmers, not 

sprayed 

2001, SpringSummer Spot - Longitudinal 45 89 0.5 - 3.3 - - - - - - -

Farmers, not sprayed 2001, SpringSummer Spot - Longitudinal 45 89 0.5 - 3.5 - - - - - - -

Farmers, self-sprayed 2001, SpringSummer Spot - Longitudinal 2 100 0.5 - 54.2 - - - - - - -
Page 86 of 124

Reference Cohort Type of Study 

Population Collection Period Collection Method n % Det.

Urinary TCPy Concentration (μg/L)

LOD AM GM Min 5th 25th 50th 75th 95th Max

Curwin et al., 

2007

Father, nonfarm 2001, SpringSummer Spot 23 94 3.32 - 13 3.8 - - - - - 47

Father, farm 2001, SpringSummer Spot 24 100 3.32 - 17 6.5 - - - - - 58

Mother, nonfarm 2001, SpringSummer Spot 24 95 3.32 - 11 1.8 - - - - - 35

Mother, farm 2001, SpringSummer Spot 24 100 3.32 - 14 5.6 - - - - - 52

Child, nonfarm 2001, SpringSummer Spot 51 100 3.32 - 15 5.4 - - - - - 54

Child, farm 2001, SpringSummer Spot 65 100 3.32 - 17 6.1 - - - - - 87

Berkowitz et 

al., 2003 Mt. Sinai 1998-2001 Spot 365 - - - - - - 1.8 7.5 25.7 - -

Whyatt et al., 

2009

Columbia Prenatal 1 2001-2004 Spot 95 52 0.26 - - - - <LOD 0.46 2.1 7.8 -

Columbia Prenatal 2 2001-2004 Spot 83 48 0.26 - - - - <LOD <LOD 1.3 3.5 -

Columbia Prenatal 3 2001-2004 Spot 54 56 0.26 - - - - <LOD 0.57 2.4 6.9 -

Columbia Prenatal 4 2001-2004 Spot 21 62 0.26 - - - - <LOD 0.61 3.3 6.8 -

Columbia Postnatal 2001-2004 Spot 73 40 0.26 - - - - <LOD <LOD 1.02 4.8 -

Castorina et 

al., 2010 

CHAMACOS Prenatal 1 1999-2000 Spot 538 71.2 0.3 - - - <LOD <LOD 2.1 5.4 16.9 56.1

CHAMACOS Prenatal 2 1999-2000 Spot 481 81.9 0.3 - - - <LOD 1 3.2 7.1 17.9 38.1

NHANES, 1999-2002 Preg. Subsample 1999-2002 Spot 224 79 0.4 - - - <LOD 0.6 1.6 3.1 7.4 61

Eskenazi et al., 

2007 

CHAMACOS 1st Preg. 2000-2003 Spot 442 71 - - - - - - 3.76 - - -

CHAMACOS 2nd Preg. 2000-2003 Spot 419 82 - - - - - - 4.6 - - -
Page 87 of 124

5.3 Chlorpyrifos Biomonitoring and AChE Inhibition

Monitoring of ChE has been performed in both research and state-based occupational surveillance to 

measure potential changes in blood ChE levels following exposure to OP and carbamate pesticides (Barr 

and Angerer, 2006; Lessenger, 2005; WA State, 2011). The most common biomarkers used to monitor 

ChE activity include erythrocyte AChE and plasma BuChE. Inhibition of erythrocyte AChE is generally 

considered the most relevant measure of potential toxicity of chlorpyrifos and other compounds that 

act as ChE inhibitors (Eaton et al., 2008). While AChE is believed to have greater toxicological relevance, 

plasma BuChE has been shown to be inhibited at lower levels of exposure than erythrocyte AChE, 

making plasma BuChE a more sensitive biomarker for the purposes of monitoring potential physiological 

change. 

Although analytical methods are available to measure AChE and BuChE levels, evaluation of inhibition 

requires careful study design and analysis. Most notably, AChE and BuChE activity can have relatively 

high inter- and intra-individual variability. Due to this variability, reference ranges of activity have not 

been established to evaluate inhibition in the general population or occupational groups. Rather, 

researchers and state-based occupational surveillance programs evaluating inhibition have typically 

utilized longitudinal study designs that compare AChE/BuChE activity during a “baseline” period where 

only background exposure is expected to occur with an “exposure” period where exposure may occur in 

excess of background levels (e.g., season in which pesticide applicators apply chlorpyrifos). AChE/BuChE 

activity can also be influenced by a number of other factors, including pregnancy status, disease status, 

exposure to other ChE inhibitors, and illegal drug use. As such, it may be important to control for these 

types of factors when evaluating inhibition. Another important consideration is that changes in AChE 

activity may be less sensitive to lower levels of environmental exposure that may be experienced by the 

general population (Barr and Angerer, 2006). For this reason, examination of AChE inhibition has 

generally been more common in occupational studies involving chemical manufacturing or agricultural 

study populations. While it may be difficult to monitor AChE inhibition in non-occupational study 

populations, some investigators have measured AChE and BuChE as indicators of pesticide exposure in 

epidemiologic investigations. Eskenazi et al. (2004), for example, measured AChE and BuChE in 

maternal and umbilical cord blood and evaluated the association between AChE and BuChE and length 

of gestation and fetal growth. 

Previous Agency Evaluation and Findings

The Agency previously evaluated the relationship between both chlorpyrifos blood and urinary TCPy 

levels and AChE activity during its 2008 SAP (EPA, 2008).25

 25 This section does not consider Kisicki et al. (1999), an intentional exposure study that monitored both 

TCPy and AChE. EPA has determined that it will not rely on Kisicki et al. (1999) because it was 

intentional human chlorpyrifos dosing study involving children less than 18 years old.

 In its evaluation, the Agency considered 

human toxicity studies that measured chlorpyrifos blood levels and AChE/BuChE inhibition in adults that 

were administered a single dose of chlorpyrifos (Nolan et al., 1982). For additional context, the Agency 

also compared Nolan et al. (1982) with other biomonitoring studies in order to determine how the 

chlorpyrifos blood levels of study participants compared to blood levels reported in epidemiologic and 

biomonitoring studies in the literature. Based on its evaluation, the Agency noted AChE inhibition in 

some study participants, but indicated that there was large variability in response to the administered 

dose of chlorpyrifos, making it difficult to draw definitive conclusions about the relationship between 
Page 88 of 124

urinary TCPy levels and AChE inhibition. In addition, the Agency also noted that the chlorpyrifos blood 

levels were orders of magnitude lower than cord and maternal blood levels observed by Whyatt et al. 

(2009) in the Columbia Cohort which included births during the pre- and post-voluntary cancellation 

period. 

Recent Studies on the Association between Urinary TCPy and AChE Inhibition

Following the 2008 SAP, two observations studies have been published that further evaluate the 

relationship between TCPy on a creatinine-adjusted basis and AChE inhibition (Galabrant et al., 2009;

Farahat et al., 2011). Both of these studies involved occupational populations exposed to chlorpyrifos at 

levels that were much higher than the general U.S. population, based on comparison with the NHANES 

1999-2000 and 2001-2002 survey cycles (See Table 13 below).26 While these studies involved study 

populations that were exposed to much higher levels than the general U.S. population, both studies 

utilized effective study designs for evaluating the association between urinary TCPy and AChE/BuChE 

inhibition.

Table 13 : Comparison of Creatinine-Adjusted Urinary TCPy levels (μg /g creatinine) reported by

Galabrant et al. (2009) and Farahat et al. (2011) with the General U.S. Population.

Population Group Time Period n

TCPy (μg /g creatinine)

Average Median (95% 

CI)

95th Percentile 

(95% CI)

NHANES (CDC, 2009)

General Population 1999-2000 1,994 NR 1.5 (1.2-1.7) 8.4(6.3-11.6)

2001-2002 2,508 NR 1.9 (1.6-2.2) 9.2 (6.9-12.3)

Galabrant et al, 2009

Exposure Group

Fall 1999 53 NR 37.5 432.6

Spring 2000 53 NR 40.7 436.3

Turnaround 50 NR 59.2 7,403.8

Fall 2000 52 NR 51.5 987.2

Reference group

Fall 1999 60 NR 3.2 13.9

Spring 2000 57 NR 4.8 11.2

Turnaround 57 NR 6.6 13.0

Fall 2000 58 NR 6.3 15.5

Farahat et al, 2011

Applicators 1- to 2- week 

Exposure 

Period

14 6,437 NR NR

Technicians 12 184 NR NR

Engineers 12 157 NR NR

NR – Not reported

In Galabrant et al. (2009), the association between AChE/BuChE was evaluated in a group of exposed 

workers at a chlorpyrifos manufacturing facilities and a reference group of chemical manufacturing 

workers. The study utilized a longitudinal design over a year-long period (1999-2000) that involved 

collection of blood and urine samples during two physical examinations and two additional first morning 

void samples that were collected during the spring and fall of the study period, respectfully. Based on 

 26 Investigators only reported results on a creatinine-adjusted basis, so the results could not be 

incorporated into the TCPy comparison provided in the previous section.
Page 89 of 124

their statistical analysis, the investigators reported that no relationship was between urinary TCPy levels 

and AChE activity was observed over the entire range of TCPy levels. While no association was reported 

in relation to AChE, the investigators reported that there was a statistically significant association 

between TCPy levels and BuChE levels above a cutpoint of 110 μg TCPy/g creatinine. 

The second recent study that has examined the association between TCPy and AChE/BuChE involved 

cotton field workers in Egypt (Farahat et al., 2011)27. In the study, daily urine samples and weekly blood 

samples were collected before, during, and after 9 – 17 consecutive days of chlorpyrifos application to 

cotton fields. All study participants were observed to have detectable levels of urinary TCPy at the start 

of the study, which ranged from 4.1 to 4,080 μg TCPy/g creatinine. Following the start of application, 

TCPy levels increased and were reported to be significantly higher in applicators (6,437 μg/g creatinine)

than in technicians (184 μg/g) and engineers (157 μg/g). Baseline BuChE activity was reported to range 

widely during the baseline period of the study. Following the start of the chlorpyrifos exposure period, 

the investigators reported that BuChE activity was suppressed, but a specific quantification of 

suppression was not reported. The AChE activity of study participants was reported to be less variable 

than BuChE during the baseline period of the study. Following exposure, the investigators reported that 

average AChE activity was inhibited by 43, 73, and 70% during the July-2, July-10, and July-24 collection 

time points, respectively. 

Based on their statistical analysis, Farahat et al. (2011) reported that they identified an inflection point 

representing a significant decrease in AChE activity at TCPy levels greater than 3,148 μg TCPy/g 

creatinine. In discussing their results, the investigators noted that few studies have attempted to 

identify a dose-effect relationship between urinary TCPy and blood ChE activity. They also discussed 

their results in comparison with Galabrant et al. (2009) and indicated that both studies observed similar 

inflection point of inhibition of BuChE activity. In particular, Galabrant et al. (2009) reported an 

inflection point of BuChE inhibition at 110 μg TCPy/g creatinine and Farahat et al. (2011) reported a 

similar reflection point at 114 TCPy/g creatinine. 

Galabrant et al. (2009) and Farahat et al. (2011) provide evidence that suggests that there is an 

association between TCPy and BuChE at higher levels of exposure. Both studies utilized effective study 

designs for evaluation of the association between urinary TCPy and AChE/BuChE inhibition. In 

particular, each study utilized a longitudinal design. Farahat et al. had both a “baseline” period where 

occupational exposure was not reasonable expected to occur and an “exposure” period with repeated 

measurement of urinary TCPy levels and blood AChE/BuChE activity. Galabrant et al. had a similar 

longitudinal design, but did not establish “baseline” AChE/BuChE levels for the purpose of evaluating 

inhibition. While Galabrant et al. were unable to establish baseline AChE/BuChE activity, their inclusion 

of a reference population enabled them to evaluate the relationship between absolute AChE activity and 

TCPy. Limitations were also noted by the investigators. Galabrant et al., for example, reported that 

there study population only included workers and did not include potentially sensitive populations, such 

as children or elderly. Their study population did include some female study participants (22.6% and 

26.7% of the exposure and reference groups, respectively).

 27 While Farahat et al. (2011) provides information on the relationship between TCPy levels and AChE 

inhibition, the levels of exposure observed in the Egyptian farmworker population may not be 

representative of U.S. farm practices. As such, it may not be appropriate to use Farahat et al. (2011) to 

make conclusions about farmworker exposure in the U.S. 
Page 90 of 124

5.4 Approaches to Interpreting Biomonitoring Data

There is a need in environmental exposure assessment to quantitatively interpret the results of 

biomonitoring studies. Quantitative analysis involves the reconstruction of human exposure estimates 

and scenarios most likely associated with measurements of analytes in human biological specimens such 

as blood and urine. The previous section provided an overview of the available TCPy data and discussed 

some of the challenges in its interpretation. The TCPy data are presented in this document as published. 

The Agency will be soliciting comments on approaches to best interpret such biomonitoring data given

the discrete and intermittent nature of human environmental exposures and chemical-specific 

pharmacokinetic behavior. As part of the consideration for the SAP, the following section summarizes

various approaches that can be used in the interpretation of the TCPy data. The approaches are 

generally discussed in order of increasing sophistication (i.e., qualitative to quantitative) including the 

most sophisticated physiologically based pharmacokinetic (PBPK) modeling approach. The strengths 

and limitations of each approach are briefly discussed.

5.4.1 Consideration of Approaches for Interpretation of Biomonitoring Data

• Qualitative-Noncomparative: 

This approach is based on simply laying out biomonitoring data as in the previous 

section. Interpretation is in the context of measures of exposure (for example, pre- and 

post- mitigation periods) without any quantitative link to exposure levels to chlorpyrifos, 

exposure route, lifestage, or health outcome.

• Qualitative-Comparative: 

Similar to Qualitative-Noncomparative except that the analysis will also include gross 

comparisons to the available database of human deliberate dosing studies with 

chlorpyrifos. The primary comparison will be to a controlled human study performed by 

Nolan and co-workers in which human volunteers were administered a single oral dose 

of 0.5 mg/kg bw chlorpyrifos and the resulting blood and urinary TCPy levels were 

measured as well as plasma cholinesterase inhibition as a potential marker of early 

effects (Nolan et al., 1984). Uncertainties arise when comparing biomonitoring data 

lacking critical information on timing of exposure to those obtained from a controlled 

study involving a single oral dose of chlorpyrifos in a single gender (males) (Nolan et al., 

1984). In general, biomonitoring data collected in observational studies reflect repeated 

environmental exposures, multiple exposure sources and routes, unknown timing of 

exposures, and unless specified, from both genders and different life-stages. 

Historically, the Nolan data have been used in the evaluation of occupational 

chlorpyrifos exposures. A number of biomonitoring studies were submitted by Dow 

AgroSciences in which pre- and post-exposure urine measures were conducted and 

measured for TCPy. Measures of total chlorpyrifos (μg) absorbed were calculated using 

the "kinetic method." This method estimated the amount of chlorpyrifos absorbed by 

fitting the amount of TCPy excreted in the individual urine specimens to a one 

compartment pharmacokinetic model that described the time course of TCPy in urine of 

volunteers following application of the chlorpyrifos to their forearm (Nolan et al., 1984). 

Based on the Nolan et al., 1984 study, it was assumed that 72% of the absorbed 

chlorpyrifos was excreted into the urine as TCPy. This methodology is more fully 

described in the example biomonitoring study report (Honeycutt and Day 1994). 
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• Semi-Quantitative Screening:

This approach would be based on the derivation of “biomonitoring equivalents (BE)” as

a screening tool for comparison to chemical levels in biomonitoring data. The BE 

approach does not generate human environmental exposure estimates per se, but 

rather uses pharmacokinetic information and modeling for forward dosimetry analysis 

to estimate biomarker levels of exposure likely to be associated with regulatory human 

external exposure guidelines (e.g., RfD). The resulting estimates can then be compared 

to those in biomonitoring data. The main caveat with this approach is that uncertainty 

arises if the biomarker lacks specificity as biomarker of exposure and/ordoes not closely 

align with a toxicologically relevant internal dose metric. Both of these represent the 

case with the use of TCPy as a biomarker of exposure. TCPy can be formed as a result of 

direct oxidative dearylation of chlorpyrifos (i.e., without the formation of chlorpyrifosoxon as intermediate) and also happens to be an environmental degradate to which 

human direct exposures take place. Therefore, caution should be exercised in the 

interpretation of biomonitoring data of TCPy based on the BE approach.

• Semi-Quantitative Reverse Dosimetry: 

This approach, like that proposed by Mage and co-workers (Mage et al., 2004), makes 

use of information on body surface area, daily creatinine excretion rate to normalize 

urinary analyte levels, and most importantly for bounding dose estimates, the ratio of 

moles of excreted analyte expected per mole of pesticide intake. The resulting human 

dose estimate represents an estimate from all relevant routes of exposure without 

accounting for route-specific bioavailability. The Mage et al. (2004) approach has been 

applied to the interpretation of at least two large datasets: urinary measurements in the 

NHANES III study (Mage et al., 2004) and in a comparative analysis of pesticide dose 

estimates for children of Iowa farmers and non-farmers (Curwin et al., 2007). The 

Agency has utilized the Mage et al., 2004 approach, in combination with other 

approaches that do not use creatinine normalization, in the analysis of biomonitoring 

data associated with triclosan (RED 2008). A major limitation of the Mage et al. (2004)

approach is that the human dose estimate does not account for route-specific 

bioavailability and the contribution of each route of exposure would need to be 

carefully elucidated for any type of interpretation. Moreover, as with any approach, the 

contribution of direct dietary exposures to TCPy as an environmental degradate would 

need to be carefully considered in any analysis associated with environmental exposures

to chlorpyrifos.

• Quantitative PBPK Approach:

By far, the preferred approach for reconstructing exposure estimates is a well-calibrated 

and evaluated PBPK model. These models provide an anatomically/physiologically 

bounded quantitative structure within which the pharmacokinetics of chemicals can be 

described. Because PBPK models can predict internal dosimetry, they can be used in 

reconstructing human exposures (including scenarios) that may be associated with 

biomonitoring data. The major disadvantage of PBPK models is that their development 

including parameterization, calibration, and evaluation requires a significant amount of 

data.

There is a well-developed PBPK model available for the AChE inhibiting mode of action 

of chlorpyrifos that provides the critical internal dose metrics associated with AChE 
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inhibition. The model was developed based on adult non-pregnant rats and was then 

scaled and calibrated for the human counterpart (i.e., adult, non-pregnant human) 

(Timchalk et al. 2002). The model in its current state can only simulate oral exposures to 

chlorpyrifos and does not account for any lifestage-related changes applicable to 

fetuses, infants, or children. 

Table 14 Comparison of Approaches for Interpretation of Chlorpyrifos (CPF) Biomonitoring Data

Approach Strengths Weaknesses/Uncertainties

Qualitative- Noncomparative

- Simple

- No need for pharmacokinetic 

information or modeling

-Gross comparison with no 

link to exposure (levels or 

route), lifestage, or health 

outcome

Qualitative-Comparative

- Simple

- No need for pharmacokinetic 

information or modeling

-Comparison is to 

controlled human oral 

studies irrespective of 

dosing regimen, exposure 

route, lifestage, or gender

Semi-Quantitative Screening

- Simple

- Makes use of some pharmacokinetic 

information/modeling

- Dose estimates can be compared to a 

exposure regulatory guideline (e.g.,RfD)

- Can incorporate some lifestagespecific information in clearance 

estimates

-Screening tool

- Urinary levels of TCPy may 

not be the most relevant 

dose metric 

Semi-Quantitative Reverse 

Dosimetry

-Simple

- Can incorporate some lifestagedependent changes in body surface area 

and creatinine excretion rate.

- Provide quantitatively bound 

estimates based on expected moles of 

urinary analyte per mole of parent 

chemical

- Provides absorbed dose 

estimates without routespecific information

- Assumes that analytes are 

derived from a single 

parent chemical

PBPK-PD model

- Provides critical internal dose metrics 

associated with AChE inhibition (levels 

of CPF-oxon and AChE inhibition)

- Provides internal dose metrics for 

potential non-cholinergic effects (levels 

chlorpyrifos and oxon metabolite)

- Does not address all 

susceptible lifestages

- Does not incorporate all 

relevant routes of exposure

Chlorpyrifos is a unique in that its database is extensive; there is a wealth of information on metabolism, 

mechanism of cholinesterase inhibition, markers of exposure and effects, and a robust PBPK-PD model 

developed for adult non-pregnant adults. Furthermore, its biomonitoring database is also extensive and 

continues to grow. The Agency is considering approaches to best evaluate the current biomonitoring 

information and is soliciting input from the SAP. There is no ideal reverse dosimetry analysis approach. 

Each approach needs to be carefully considered for its strengths and limitations. Table 14 provides a 

brief synposis of several approaches discussed above. The following section 5.4.2 provides a more 

extensive background of the PBPK-PD model available for chlorpyrifos.
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5.4.2 Physiologically Based Pharmacokinetic-Pharmacodynamic Model for 

Chlorpyrifos: State of the Science

A physiologically based pharmacokinetic model (PBPK) model for chlorpyrifos that was originally 

developed by Timchalk and coworkers in 2002 (Timchalk et al., 2002) has been refined over the years as 

more data has become available (Busby-Hjerpe et al., 2010; Cole et al., 2005; Garabrant et al., 2009; Lee 

et al., 2009; Lowe et al., 2009; Lu et al., 2010; Marty et al., 2007; Timchalk and Poet, 2008; Timchalk et 

al., 2005; Timchalk et al., 2006). The model is based on the mechanistic mode of action information that 

chlorpyrifos exerts its cholinergic toxicological effects through metabolic activation to chlorpyrifos-oxon 

and detoxification to TCPy and DAP metabolites DEP and DMP. Chlorpyrifos-oxon inhibits AChE through 

covalent modification of an active site serine active site, resulting in build-up of acetylcholine and 

hyperstimulation at synapses in brain, spinal cord, and peripheral nervous system (Timchalk et al., 

2002). The chlorpyrifos PBPK model has evolved to include the key event in the cholinergic toxicity mode 

of action: AChE inhibition. Thus, the PBPK model includes pharmacodynamics (i.e., a PBPK-PD model) 

that can be used to predict the critical dose metrics associated with cholinergic toxicity following 

chlorpyrifos exposure: CPF-oxon levels and the resulting AChE inhibition in target tissues. As CPF-oxon is 

very reactive and can modify AChE at non-target sites such as red blood cells (RBCs) as well as other 

proteins with esterase activity, the model also includes inhibition of AChE activity in RBCs and plasma 

cholinesterase as potential biomarkers of early effects. Also predicted by the model are TCPy levels (in 

plasma and excreted urine) as a biomarker of exposure given that this metabolite is associated with 

detoxification of chlorpyrifos.

Although variations of the original Timchalk et al. (2002) PBPK-PD model have been published, the 

fundamental model structure has remained largely the same. The Agency’s SAP recommended in 2008 

that the Agency consider further refinement of the Timchalk et al. (2002) model for use in the human 

health risk assessment of chlorpyrifos (FIFRA SAP, 2008). In recent collaborative research effort between 

Battelle and Dow Agrosciences (DAS et al., 2011), the chlorpyrifos PBPK-PD model was linked to the 

probabilistic dietary exposure model CARES and expanded it to include different age groups: infants (6 

months), children (3 year-olds), and adults (30 year olds). The final composite model represents a 

“source-to-outcome” model that, if robust enough, could be used to assess population risk to 

cholinesterase inhibition starting with data on crop residues levels of chlorpyrifos. The Agency’s Science 

SAP reviewed the final composite model at its February 2011 meeting (FIFRA SAP, 2011). Although the 

Panel commended the effort, confidence was not expressed in using the model developed for the 

typical adult non-pregnant human to infer on the pharmacokinetics of chlorpyrifos (and its metabolites) 

in earlier lifestages. One concern reason cited by the Panel was that the model did not include 

gestational and lactational exposures which require explicit attention. Further, the Panel expressed 

concern about the in vitro-based parameterization of metabolism in the model based on a limited 

number of age-appropriate samples (FIFRA SAP, 2011).

5.4.2.1 Lifestages considered in the current chlorpyrifos PBPK-PD model

The current PBPK-PD model for chlorpyrifos is based on the typical non-pregnant adult rat and human 

and does not inform about lifestage-related susceptibility. This is a key deficiency in the model for use in 

human health risk assessment because there is a large body of evidence indicating that the developing 

nervous system may be a particular susceptible target of toxicity associated with chlorpyrifos. The 

evidence comes from rodent studies from multiple laboratories indicating neurobehavioral changes 

resulting from peri-natal exposures to chlorpyrifos. Epidemiological studies have also supported 

associations between prenatal exposures to OPs including chlorpyrifos, and cognitive delays in children. 
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At the February 2011 SAP meeting, the Panel specifically noted the lack of maternal and fetal 

pharmacokinetics and pharmacodynamics in the current PBPK-PD model as a deficit to inform about 

pregnancy and early lifestages (FIFRA SAP 2011) and criticized the use of a simple multiplicative factor to 

account for pregnancy-related changes in order to make comparisons to biomonitoring data (DAS et al. 

2011). 

Attempts have been made to extrapolate the current chlorpyrifos PBPK-PD model to include earlier 

lifestages. In another effort by Timchalk and coworkers (Timchalk et al., 2007), the existing adult rat 

PBPK-PD model was extrapolated to preweanling rats by incorporating age-dependent changes in tissue 

volumes, blood flows, and CYP450 and PON-1 metabolic activities. The modeling results supported the 

notion that the increased susceptibility of young rats may due to pharmacokinetic differences; 

chlorpyrifos-oxon levels in blood and brain increased disproportionally with chlorpyrifos dose as 

compared to adult non-pregnant rats. With the exception of not incorporating in utero and lactational 

exposures, the age-dependent rat model provided a good starting point for developing a human 

equivalent PBPK model that would address exposures to infants and children (Timchalk et al., 2007).

A more recent effort to extrapolate the model to earlier lifestages was published in the peer-reviewed 

literature by Lu and co-workers (Lu et al., 2010). The modeling effort focused on developing a PBPK 

model to predict urinary levels of TCPy in children. The model was implemented using the Agency’s 

Exposure Related Dose Estimating Model (ERDEM) to accommodate for aggregate exposures to 

chlorpyrifos including bolus ingestion, inhalation, and ingestion based on non-dietary routes (hand-tomouth)(Blancato et al., 2006). The dermal route was not included in the analysis because chlorpyrifos 

was not detected on any of the children’s hands. Although a significant amount of data based 

chlorpyrifos measurements in indoor air, food, and house dust was used in the effort, the resulting 

model performed poorly by significantly under-predicting levels of TCPy in children (Lu et al., 2010). The 

Agency is also aware of a PBPK modeling effort led by Dr. Dale Hattis of Clark University in collaboration 

with the Columbia University epidemiology team. This PBPK model may, in the future, be useful in 

clarifying the exposure concentrations that correspond to the chlorpyrifos levels in umbilical cord blood 

associated with statistically significant neurodevelopmental effects. However, the development of the 

Hattis effort is further behind than other research groups (Hattis, 2011)

In summary, there is a robust PBPK-PD model available for chlorpyrifos for describing its AChE-inhibition 

mode of action; however, the model is restricted to the “typical” non-pregnant adult human and does 

not inform about the most susceptible lifestages which includes pregnancy, infants, and children. 

Although several modeling efforts have been undertaken to extrapolate the current model to different 

lifestages, the efforts have yet to produce a robust PBPK model that can be used for human health risk 

assessment applications (as opposed to for AChE-inhibition mode of action). Modeling future efforts

aimed PBPK modeling for chlorpyrifos needs to include in utero exposures along with dermal and/or 

inhalation routes of exposure. The absence of gestational stages in the model was regarded by the SAP 

as a major deficit that prevents the chlorpyrifos PBPK-PD model from being an optimal prototype for risk 

assessments (FIFRA SAP, 2011).
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5.4.2.2 Routes of Exposure simulated by the PBPK-PD model

The current PBPK-PD model for chlorpyrifos is limited to oral exposures to chlorpyrifos with particular 

emphasis on dietary exposures (excluding nursing infants), even though dermal and inhalation are also 

important routes of exposure, particularly in occupational settings. The pharmacokinetics of chlorpyrifos 

(and its metabolites) and resulting cholinesterase inhibition profile via other relevant routes of exposure 

are expected to be distinct as compared to the oral route profile predicted by the model. For a rapidly 

and extensively metabolized chemical such as chlorpyrifos whose cholinergic toxicity is dictated by 

bioactivation, route-specific metabolism can have a dramatic influence on the site, extent, and 

temporality of toxicity. For instance, in a rat study where dosing was performed via the oral route, 

cholinesterase inhibition occurred in plasma far earlier and at a lower dose than in AChE inhibition in 

brain tissue (Mendrala et al. 1998). In contrast, when chlorpyrifos was administered to adult nonpregnant rats via an aerosol through the inhalation route, the result was cholinesterase inhibition 

occurring far earlier in lung tissue followed by plasma, while no AChE inhibition was reported in brain 

tissue at any of the exposure levels tested (Hotchkiss et al. 2010). A dermal toxicity study in rats resulted 

in a similar profile as via the oral route (plasma cholinesterase inhibition occurred to a greater extent 

than RBC AChE inhibition). However, it is important to note that, via the dermal route, blood flow to the 

liver, the major site of metabolism of chlorpyrifos, will be limited by the input of the hepatic artery 

which can lead to temporal changes in cholinesterase inhibition as compared to other routes of 

exposure.

Therefore, the uncertainties associated with the pharmacokinetics of chlorpyrifos via dermal and 

inhalation routes significantly limit the utility of the current chlorpyrifos PBPK model for risk assessment 

applications. The SAP also supported this conclusion at its February 2011 meeting by stating: “the 

important pharmacokinetic differences between oral dosing and other routes of exposure, first pass 

metabolism, can have a dramatic influence on the extent and timing of the functional changes under 

consideration following OP exposure.” (FIFRA SAP 2011)

5.5 Conclusions

A large number of biomonitoring studies have evaluated exposure to chlorpyrifos by measuring urinary 

TCPy levels. While the biomonitoring database is extensive and continues to grow, the relationship 

between urinary TCPy levels and external chlorpyrifos exposure remains uncertain, making it difficult to 

directly evaluate biomonitoring data in a risk assessment context. One important interpretative 

limitation is that TCPy lacks the desired specificity as biomarker because urinary TCPy reflects 

concurrent human exposures to chlorpyrifos and TCPy as an environmental degradant. Further, reverse 

dosimetry approaches including a PBPK model are available, but each has limitations that need to be 

carefully considered for any type of interpretation. The current PBPK model for chlorpyrifos is limited to 

simulating oral exposures to chlorpyrifos and thus cannot be used to evaluate urinary TCPy which is 

likely also reflective other routes of exposure such as inhalation and dermal as well as direct exposures 

to TCPy. Further, the model has limited utility for evaluating maternal/umbilical cord blood 

measurements of chlorpyrifos since it does not incorporate the critical physiological and chemicalspecific changes associated with pregnancy and in utero exposures. The ideal PBPK/PD model would 

accommodate oral, dermal, and inhalation routes and be informative of multiple lifestages (e.g., 

gestation, birth, childhood, and adulthood). This ideal model may not be available for the revised risk 

assessment. In the absence of such a model, the Agency is soliciting comments from the Panel on 

approaches for interpreting the biomonitoring data currently available for chlorpyrifos.
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6.0 Summary & Preliminary Conclusions

Chlorpyrifos is a widely studied pesticide using an expansive suite of toxicity tests conducted in 

accordance with harmonized, scientifically peer-reviewed study protocols and published in the scientific 

literature. Thus, chlorpyrifos has a robust database across many scientific areas (e.g. metabolism, mode 

of action, biomonitoring, and human health toxicity studies) and lines of scientific evidence (i.e. in vitro,

laboratory animal, in vivo, epidemiology), reflective of different levels of biological organization, and 

across various lifestages (fetuses, postnatal, pregnant and non-pregnant adult). 

It is well established that AChE inhibition is the primary mode of action/adverse outcome pathway for 

OPs, like chlorpyrifos. Because AChE inhibition is the initiating event for this mode of action/adverse 

outcome pathway, using AChE inhibition as a regulatory endpoint is protective of downstream 

cholinergic effects. Moreover, historically, given the sensitivity of AChE inhibition data for OPs, these 

data have been considered to be protective of other potential toxicities and/or modes of action for OPs. 

As discussed in detail throughout this draft issue paper, newer lines of research on chlorpyrifos, 

including epidemiological studies in mothers and children, have posed the issue of whether AChE 

inhibition is the most sensitive endpoint, and thus have raised some uncertainty in the chlorpyrifos risk 

assessment. In order to determine the degree to which these recent studies are appropriate for 

incorporation into risk assessment (qualitatively and/or quantitatively) the Agency is taking a stepwise, 

objective and transparent approach to evaluate and interpret all the lines of scientific information 

related to the potential for adverse neurodevelopmental effects in infants and children as a result of 

prenatal exposure to chlorpyrifos, as well as to characterize thoroughly the strengths and uncertainties 

associated with these studies. 

The evaluation of these recent studies is stepwise process which began with the September 2008 FIFRA 

SAP involving a preliminary review of the literature for chlorpyrifos, with a particular focus on women 

and children. Subsequent Agency activities have involved developing approaches for performing risk 

assessment of semi-volatile pesticides, and developing the draft “Framework for Incorporating Human 

Epidemiologic & Incident Data in Health Risk Assessment” for integration of epidemiology with other 

types of experimental data (USEPA, 2010a, b; FIFRA SAP 2010a,b). In summer 2011, the Agency released 

its preliminary human health risk assessment that focused on the AChE inhibiting potential of 

chlorpyrifos. This focus is consistent with the recommendation from the 2008 SAP that AChE data 

provide the most appropriate endpoint and dose-response data for deriving point of departure for 

purposes of risk assessment. Moreover, because of the Agency’s long experience with assessing the 

potential risk to chlorpyrifos and other OPs, and because both the dose response approach based on 

AChE inhibition and also the exposure methodologies used (as noted in previous sections of this issue 

paper) in the 2011 assessment have been vetted by numerous SAPs, there is high confidence in those 

analyses. 

The 2011 risk assessment noted that a full weight of the evidence analysis that explicitly considers 

uncertainty and implications of modes of toxic action related to experimental and epidemiologic lines of 

evidence using factors such as biological plausibility, strength, consistency, dose-response and temporal 

concordance, will be conducted in the future. A full weight of the evidence including characterization of 

uncertainty has not yet been conducted; however, the April 2012 FIFRA SAP is an important step toward 

this effort. Unlike the dose-response and exposure analyses conducted for the 2011 chlorpyrifos risk 

assessment, the types of information reviewed here (e.g., epidemiology studies, biomonitoring data, 

modes of action/adverse outcome pathways for neurodevelopmental outcomes, etc) are less often 
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available and approaches for their use in risk assessment (qualitatively or quantitatively) are less well 

established. The input from the SAP will be helpful as the Agency moves to the next stages of the 

weight of evidence analysis and uncertainty characterization including presenting various risk 

assessment alternatives or options and describing the range of potential risks. 

This 2012 draft issue paper builds on the 2008 SAP review and is informed by the public comments 

received in 2011 in response to the preliminary chlorpyrifos human health risk assessment. In previous 

sections of this issue paper, the state of the science with respect to lifestage susceptibility, modes of 

action/adverse outcome pathways for AChE inhibition and for neurodevelopmental outcomes, 

epidemiology studies in mothers and children, biomonitoring data, and PBPK modeling have been fully 

reviewed. Each of these lines of evidence is important for the pending weight of the evidence analysis 

and uncertainty characterization.

• AChE inhibition: 

In preparation for the 2008 SAP, the Agency performed a comprehensive review of the 

literature for AChE inhibition. This review has been supplemented with the newest data; 

together, the preliminary 2008 literature review and the newest studies inform the selection of 

point of departures for risk assessment (Section 3.1). The approach of using AChE inhibition 

data for derivation of point of departures is consistent with the advice of the SAP from 2008 in 

that AChE data remain the most sensitive and robust dose-response data in experimental 

toxicology studies. Consistent with the concept that younger juvenile animals are more 

sensitive due to immature metabolic systems, PND11 rat pups provide the most sensitive 

cholinesterase inhibition data for deriving the acute oral POD (BMDL10 0.36 mg/kg). As the 

metabolic systems mature at or near adult levels, this age-dependant sensitivity disappears. In 

repeated dosing studies, pups exhibit similar sensitivity to adult rats and it is pregnant dams 

which provide the most sensitive data for deriving a repeated exposure oral point of departure 

(BMDL10 0.03 mg/kg/d ).

• Neurodevelopmental Outcomes in Laboratory Animals:

In the 2008 preliminary review, the Agency evaluated the neurobehavioral studies available at 

that time. Since then, the database of studies with laboratory animals (rats and mice) indicating 

that gestational and/or early postnatal exposure to chlorpyrifos may cause persistent effects 

into adulthood has grown (Section 3.2.2). This literature consistently shows effects related to in 

utero and/or early post-natal exposure; however, there is considerable variability in the 

outcomes. This literature includes a range of exposure periods, dosing scenarios, testing 

strategies, and specific methodologies and equipment used. Given the wide array of testing that 

has been conducted, some variability is not unexpected and in fact, the consistency of finding 

neurological effects is more striking. 

Many studies report effects at a dose of 1 mg/kg/d, a dose that produces some amount of brain 

ChE inhibition when given directly to the pups postnatally, but may or may not alter fetal brain 

ChE activity when given to the dams gestationally. One study (Braquenier et al., 2010) using 

lower doses, administered to the dam on GD15-LD14, reported a NOEL of 0.2 mg/kg/d. The 

AChE inhibition studies described in section 3.1 provide robust dose-response data with which 

to conduct BMD modeling for derivation of points of departure, but the animal data on 

outcomes such as behavior do not provide dose-response data amenable to such modeling. 

Many studies use only one, maybe two doses, and in several cases the lower but not higher dose 
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level produced significant effects. Comparing a NOEL of 0.2 mg/kg/d to a repeated dosing 

BMDL10 of 0.03 mg/kg/d suggests that AChE inhibition is a sensitive and protective endpoint; 

however, ideally studies that thoroughly evaluate neurodevelopmental effects at lower, noninhibitory doses could be helpful. Additional studies that thoroughly characterize (time-course, 

dose-response) the potential for fetal brain AChE inhibition during gestational exposures and/or 

or pup brain ChE inhibition during postnatal exposures, could be combined with behavioral 

studies with which to quantitatively evaluate potential AChE inhibition compared with long-term 

effects. 

In addition, numerous studies on the possible mechanistic aspects of neurodevelopmental 

effects have been published (Section 3.2.1). The results have led some research groups to 

propose that changes in brain connectivity and/or neurochemistry may underlie the long-term 

in vivo neurobehavioral changes observed into adulthood. While multiple biologically plausible 

hypotheses are being pursued by researchers, no one pathway has sufficient data to be 

considered more credible than the others. There is some evidence of possible effects which are 

similarly sensitive or more sensitive than AChE inhibition. Because of insufficient data 

establishing the causal linkages among different levels of biological organization to adversity, a 

mode of action or adverse outcome pathway leading to effects on the developing brain cannot 

be established. For example, while there is in vitro evidence relating binding of 

chlorpyrifos/chlorpyrifos oxon to AChE and the subsequent decrease in neurite outgrowth at the 

cellular level, the relationship between neurite outgrowth and neurodevelopmental 

consequences has not been established. As described in the 2007 NRC report on Toxicity Testing 

in the 21st Century, to develop an adverse outcome pathway, not only does one have to 

establish plausible relationships among the key events, but quantitative relationships also need 

to be established, i.e., how much of a change in one key event is needed to result in an adverse 

effect at the next level of biological organization? Thus, certain exposure to a chemical may 

impact normal physiological responses in a way that may not necessarily be adverse and thus, 

the AOP concept requires an understanding of adaptive/homeostatic capacity of biological 

systems and their limits, relative to concentration and duration of exposure.

To follow the example above, the concentration response relationship relating the amount of 

chlorpyrifos/chlorpyrifos oxon binding to AChE, the inhibition of neurite outgrowth, and the 

level of neurodevelopmental deficit has not been established. Some molecular initiating events 

and some cellular effects occur below exposure levels that inhibit cholinesterase. Most of the 

data, however, are limited to reports from one laboratory and are not anchored to an adverse 

neurodevelopmental effect at the tissue or organism level. 

While it is difficult to establish an adverse outcome pathway in the developing nervous system 

due to on-going dynamic developmental processes and multiple critical windows of 

development, the following experimental design aspects are critical to establishing a 

relationship between chlorpyrifos exposure and adverse effect.

o Concentration/dose-response relationship for each level of biological 

organization of the adverse outcome pathway. Ideally, this should include 

concentrations/doses that do not inhibit cholinesterase. 

o In vitro studies in neural cell cultures need to be translated to effects in more 

complex biological systems, i.e., developing, intact nervous system.

o It would be desirable to look at the molecular initiating event and the 

downstream events in the same biological system. 
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o Demonstration of a relationship between an alteration(s) of a key 

neurodevelopmental event at one level to that in the intact nervous system.

For example, how much of a change in neuronal proliferation causes 

neurodevelopmental deficits? 

The Agency will be soliciting comment from the SAP on the Agency’s preliminary conclusions 

(see below) and on the strengths and uncertainties associated with the in vivo and in vitro

experimental toxicology studies with a particular focus on the extent to which these studies 

suggest potential for long-term effects of chlorpyrifos from early-life exposure and/or the extent 

to which they suggest neurodevelopmental effects below doses which elicit 10% AChE 

inhibition.

• Epidemiology studies in mothers and children: 

In Section 4.0 of this issue paper, EPA summarized the available epidemiologic data concerning 

the relation between gestational exposure to chlorpyrifos and adverse neurodevelopmental 

outcomes in infants and children from three major children’s health cohort studies in the U.S. 

The currently available database expands upon those studies available at the time of the 2008 

FIFRA SAP meeting on chlorpyrifos adding both new epidemiologic evaluations and new 

methodological studies concerning key measures in these studies. In addition, supplementary 

analyses which address some areas of uncertainty identified by the 2008 FIFRA SAP are also 

available for consideration. Overall, the newly available data support and strengthen the 

conclusion of the 2008 FIFRA SAP that prenatal exposure to chlorpyrifos likely plays a role in 

adverse neurodevelopmental outcomes measured in children. Specifically, high-quality etiologic 

studies from each of the children’s health cohorts consistently illustrate evidence for a putative 

role of prenatal chlorpyrifos exposure on intelligence measured at school age (7-years). In these 

concurrent publications (Spring 2011), authors performed numerous model checking exercises 

and sensitivity analyses to reduce or eliminate alternative explanations for the findings, 

strengthening the overall interpretation of these data. Additionally, primarily within the 

Columbia Mother’s and Newborn study, investigators performed exposure validation research 

suggesting the use of a measure of prenatal chlorpyrifos concentration in cord blood collected 

at delivery may be a reasonably accurate measure to categorize (rank) exposure among cohort 

participants. Additionally, the contribution of the assessment of infant neurodevelopment as 

measured by Bayley Scores in relation to prenatal organophosphate exposure by researchers 

with the Mt. Sinai Child Development study as well as the evaluation of effect modification by 

PON1 genotype or PON1 phenotype by researchers with two of the three children’s health 

cohorts adds to the totality of the database. Overall, epidemiology studies available since the 

2008 FIFRA SAP review strengthen the conclusion that prenatal chlorpyrifos exposure likely 

played a role in these adverse neurodevelopmental outcomes observed in children through 

school-age. 

While both strengths and notable limitations are present in these studies, EPA believes that the 

limitation present in the design, conduct, and analyses of these observational studies would 

likely under-estimate, rather than over-estimate measured associations (i.e., measures of 

relative risk). Specifically, given the prospective study design many sources of error are 

eliminated or minimized, including temporal bias and most instances of differential 

measurement error. Errors in the measurement of both neurodevelopmental outcomes and 

exposure most likely occurred in these studies. While the magnitude of this error is difficult to 
Page 100 of 124

quantify, this source of error would unlikely be related to the either the outcomes or exposures 

measured in these studies, respectively, and therefore, would tend to under-estimate measured 

risks (non-differential measurement error). Alternatively, factors which may have biased the 

reported risk estimates away from the null (over-estimated risks) may also be present including 

unmeasured or poorly measured factors related to socio-economic status or other 

environmental exposures both of which may be positively correlated with the outcome (adverse 

neurodevelopmental effects) and the exposure (chlorpyrifos). Considering both supplemental 

analyses and methodological research published since the 2008 FIFRA SAP review, available 

evidence does not suggest a substantial bias of this nature; however, the possibility cannot be 

completely excluded. 

The epidemiology studies reviewed are high-quality, well-conducted studies with numerous 

strengths, as well as some limitations. These limitations include use of non-specific biomarkers 

of chlorpyrifos exposure in some of these studies, lack of measurement of postnatal chemical 

exposure, and the proportion of missing data among several key variables in these 

investigations. There is evidence of consistent associations, of a moderately strong magnitude 

with some evidence of an exposure-response across many of these investigations; however, 

instances of inconsistency across studies of the same association are also observed, which may 

reflect exposure measurement variability or true inconsistency in observed effects. Currently, 

while biologically plausible explanations can be posited, a clear mode of action for these effects 

has not yet been elucidated, as discussed in detail in the paper. The lack of an established mode 

of action for neurodevelopmental effects is problematic for both causal inference as well as 

determining appropriate points of departure among other uses in quantitative risk assessment.

The Agency will be soliciting comment from the SAP on the strengths and uncertainties 

associated with the epidemiology and its incorporation into risk and uncertainty 

characterization.

• Biomonitoring data: 

Chlorpyrifos is one of a small set of pesticides for which there is a large database of 

biomonitoring studies available for a variety of lifestages and subpopulations (Section 5.0). 

Biomonitoring data provide real world information on exposed individuals; however, their 

interpretation is challenging given that the timing, magnitude, and source of exposures are most 

often unknown or only partially known. Since the 2008 SAP, there has been significant growth 

in the scientific literature on approaches for interpreting biomonitoring data. Such approaches 

range in their level of sophistication and the availability of the pertinent data to inform them 

(Section 5.4). In the epidemiology studies, measured exposure likely underestimates actual 

exposure experienced by participants, in most instances, therefore direct use of exposure 

estimates from epidemiologic analyses for use in establishing points of departure or modeling 

exposure-response associations in quantitative risk assessment is less than optimal. Therefore, 

exposure estimates from observational studies would likely require supplementary information 

or use of PBPK/PD modeling-type approach to approximate actual exposure. Given the findings 

of the children’s health cohorts and the Columbia cohort in particular, the Agency has an 

interest in considering the degree to which other groups of individuals in the U.S. may be 

exposed to levels of chlorpyrifos comparable with those from Columbia University cohort,

particularly prior to the cancellation of the indoor residential uses. The Agency will be soliciting 

comment from the SAP on the strengths and uncertainties associated with such approaches for 

interpreting biomonitoring data for use in risk characterization and uncertainty analysis. One 
Page 101 of 124

approach would be to use a PBPK-PD model which accounts for all relevant routes of exposure 

and incorporates a gestational component combined with a dose reconstruction analysis. At 

this time, the existing PBPK-PD models are not equipped for this sophisticated analysis. It is 

proposed that this is an area where additional research could significantly improve the Agency’s 

ability to determine whether actual exposures experienced by participants in the epidemiology 

studies were above or below AChE inhibition concentrations. 

As discussed in previous sections of this draft issue paper, two of the key scientific questions being 

considered in preparation for the weight of the evidence are: 1) the degree to which scientific data 

suggest that chlorpyrifos causes long-term neurodevelopmental effects from fetal or early life exposure 

and 2) the degree to which adverse effects can be attributed to doses lower than those which elicit 10% 

inhibition of AChE, i.e., the dose levels previously used for regulatory decision making. The evaluation 

of these scientific questions are multi-faceted and require integration of multiple lines of data and an 

understanding of the nature and degree of the uncertainties around the data and impact on hypotheses 

or alternative interpretations. At this time, this evaluation is incomplete, and thus in the absence of a 

full weight of the evidence, the Agency has not yet made firm conclusions. The Agency has, however, 

made preliminary interpretations and conclusions on the information summarized above which will be 

reviewed by the SAP. 

• Experimental toxicology studies in rodents suggest that long-term effects from chlorpyrifos 

exposure may occur. Due to the dose selections in most of these in vivo studies evaluating 

effects such as behavior and cognition, it is not known whether such adverse effects would be 

shown at doses lower than those used for derivation of point of departures. Despite this 

uncertainty, however, comparing a NOEL of 0.2 mg/kg/d to a repeated dosing BMDL10 of 0.03 

mg/kg/d suggests that ChE inhibition is a sensitive and protective endpoint.

• With respect to adverse effects in humans, the Agency has not yet developed a full weight of 

the evidence analysis to consider the extent to which causality can be postulated based upon all 

of the available evidence including the observational studies in the human population. 

Consistent with the 2008 SAP, however, the Agency has preliminarily concluded that the 

epidemiologic data support and strengthen a finding that chlorpyrifos likely played a role in the 

neurodevelopmental effects observed in children in association with exposure during gestation. 

Although actual level of such exposure during the critical window(s) of susceptibility is not 

known, the cord blood and other (meconium) measures from the Columbia University study 

provide qualitative, strong evidence of exposure to the fetus during gestation. While there are 

significant uncertainties remain about the actual exposure levels experienced by participants in 

the three children’s health cohorts, particularly during the time period prior to the voluntary 

cancellation of indoor residential uses of chlorpyrifos containing pesticide products, and the 

degree to which those levels were above, at, or below those which would elicit 10% AChE 

inhibition, exposures measured in the range reported in the epidemiology studies (pg/g plasma) 

are likely low enough that is unlikely to result in AChE inhibition.

• At this time, a mode of action/adverse outcome pathway has not been established for 

neurodevelopmental outcomes. This growing body of literature does demonstrate, however, 

that chlorpyrifos and/or its oxon are biologically active on a number of processes that affect the 

developing brain. Although these studies are not appropriate for deriving point of departures

and quantitative relationships between these perturbations and adversity is not known, it is 

notable that some of these effects are similarly sensitive or more sensitive than AChE inhibition. 
Page 102 of 124

Moreover, the lack of established mode of action/adverse outcome pathway does not 

undermine or reduce the confidence in the qualitative conclusions of the epidemiology studies. 

The context of quantitative risk assessment, the lack of an established mode of action/adverse 

outcome pathway complicates the way in which the epidemiological studies can be used in 

relation to important factors such as dose-response, critical duration of exposure, and 

window(s) of susceptibility. 

• With respect to the key questions being considered by the Agency with this 2012 draft issue 

paper (i.e, whether chlorpyrifos causes long-term effects from fetal or early life exposure and if 

adverse effects can be attributed to doses lower than those which elicit 10% inhibition of AChE), 

when taken together the evidence from 1) the experimental toxicology studies evaluating 

outcomes such as behavior and cognitive function; 2) mechanistic data on possible adverse 

outcome pathways/modes of action; and 3) epidemiologic and biomonitoring studies:

o Qualitatively, the Agency preliminarily concludes that these lines of evidence together 

support a conclusion that exposure to chlorpyrifos results in adverse 

neurodevelopmental outcomes in humans, at least under some (still unclear) 

conditions.

o Quantitatively, the dose –response relationship of AChE inhibition across different life 

stages is established, but other adverse outcome pathways/modes of action are not 

established. The question posed regards the nature and degree of uncertainty around 

points of departure based on 10% AChE inhibition to protect against 

neurodevelopmental outcomes. 

 The database of in vivo animal toxicology neurodevelopmental studies on 

adverse outcomes includes only a small number of studies at or near the 

Agency’s point of departure. Despite this uncertainty, the Agency’s chronic oral 

point of departure is approximately 10-fold lower than doses used in repeated 

dosing studies.

With respect to the mechanistic data, there are some effects which are similarly 

sensitive or more sensitive than AChE inhibition. The fact that there are, 

however, sparse data to support the in vitro to in vivo extrapolation, or the 

extrapolation from biological perturbation to adverse consequence significantly 

limits their quantitative use in risk assessment. 

 As noted above, the lack of an established adverse outcome pathway/mode of 

action makes quantitative use of the epidemiology study in risk assessment 

challenging, particularly with respect to dose-response, critical duration of 

exposure, and window(s) of susceptibility. Despite this uncertainty, the cord 

blood and other measures (meconium) provide qualitative, strong evidence of 

exposure to the fetus during gestation. Moreover, exposure levels in the range 

measured in the epidemiology studies (pg/g) are likely low enough that is 

unlikely to result in AChE inhibition; however, the actual level of such exposure 

during the critical window(s) of susceptibility is not known. 
Page 103 of 124

 There are a variety of dose reconstruction and reverse dosimetry approaches of 

to evaluate the extent that exposures experienced by the mothers in the 

Columbia cohort are potentially high enough to elicit AChE inhibition. These 

vary in their sophistication. One robust approach to inform the scientific 

question of the degree to which neurodevelopmental outcomes may/may not 

occur below AChE inhibiting doses used for regulatory decision making would 

involve reverse dosimetry and dose reconstruction using a PBPK/PD model that 

predicts a variety of dose metrics including AChE inhibition and blood/urinary 

biomarker levels and can simulate different exposure scenarios. Such an 

analysis would also require a more in-depth evaluation of the specific exposure 

conditions likely experienced during the time period of these epidemiology 

studies. Preferably such a PBPK/PD model could accommodate oral, dermal, 

and inhalation routes in addition to multiple lifestages (e.g., gestation, birth, 

childhood, adulthood). This ideal model may not be available for the revised 

risk assessment. In the absence of such a model, the Agency is soliciting 

comments from the Panel on approaches for interpreting biomonitoring data, 

synthesizing data from several lines of evidence to support standard 

assumptions regarding exposure assessment and reconstruction, in addition to 

integrating exposure and hazard information more comprehensively than has 

been performed. In the absence of a sophisticated, life-stage appropriate 

PBPK/PD model, there may be other more practical options to pursue that 

would be fruitful for inclusion in the quantitative chlorpyrifos risk assessment. 

The Agency will solicit comment on these preliminary conclusions, their strengths and uncertainties 

along with the data and analysis that underlie them at the April, 2012 SAP. In addition, the Agency will 

solicit comment on the further analyses that may be conducted to further inform the potential for longterm effects of early life exposure to chlorpyrifos and to evaluate the extent to which these potential 

effects may (or may not) occur at or below 10% AChE inhibition levels previously used for regulatory 

decision making. 
Page 104 of 124

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