# Meeting of the FIFRA Scientific Advisory Panel

Draft Issue Paper: Scientific Issues Concerning Health Effects of Chlorpyrifos

April 10

- 13, 2012Page 2 of 124 

## Table of Contents

**1.0 INTRODUCTION**  **5**

**2.0 SOURCE**

**- TO** 

**- OUTCOME CONSIDERATIO**  **NS 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 INHIB ITION & PLAUSIBLE PA**  **THWAYS**
**LEADING TO NEURODEVE LOPMENTAL OUTCOMES**  **15**

3.1 Adverse Outcome Pathway: AChE Inhibition 15

3.1.1 I nitiating 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 Develop ing 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 Funct ion 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 CHLOR **PYRIFOS EPIDEMIOLOGY REGARDING CHILDREN’S**  **HEALTH 54** 

4.1 Scope and Purpose 54

4.2 Summa ry of Epidemiology Findings 55Page 3 of 124 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 Epidemiolo gy Synthesis and Evaluation 71

4.4 Conclusions 75



### 5.0 EXPOSURE PROFIL **E & BIOMONITORING RE SEARCH**  **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 & PRELI **MINARY CONCLUSIONS 96** 

**7.0 REFERENCES 104** Page 4 of 124 

## Table of Figures

Figure 1 Source to effects pathway (adapted from NRC, 2007). 6 Figure 2 Conceptual framework consi dering 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 Biom onitoring 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 Colum bia 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 Develo pment. 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 92Page 5 of 124 

## 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 issu e 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 p re natal 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 ag reed 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 bi nds 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.
Consis tent with the focus on ChE inhibition

, in the 2011 preliminary risk assessment, EPA evaluated the extensive database of ChE data for multiple lifestages and sel ected points of departure based on consideration of all quality and reliable data.

In 2010, th e 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 out come 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 conc ept 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 experi mental 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 analy sis

.

The draft framework is oriented around the source to outcome pathway ( Figure 1 ) as discussed in the National Research Council’s (NRC) report on t oxicity t esting in the 21 st Century (National Research Council, 2007)

. Since the 2008 SAP on chlorpyrifos, the Agency has re viewed new experimentalPage 6 of 124 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 a nalysis. T his 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 N RC , 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 s cientific 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 a vailable 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 is sues 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 th at such effects may (or may not) occur below doses established from ChE inhibition including in postnatal animals and used for regulatory p urposes.Page 7 of 124 **2.0** 

## Source

**- to** 

**- Outcome Considerations** 

**2.1 Exposure Pathways for Chlorpyrifos**

Chlorpyrifos is an insecticide, acaricide and miticide that control s many insect pest s. Chlorpyrifos was first registered in 1965 for use in a variety of food a nd 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 ris ks 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 De cember 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 adulti cide, 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 p atterns (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 tha t 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 regul atory 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.htmPage 8 of 124 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. , gol f 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 defi ned 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 i s 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 previou sly treated ( e.g. , while harvesting fruit or vegetables). Occupational exposures are evaluated based on the requirements of the FIFRA. In a manner analogous to res idential

exposures described above, occupational exposures are also evaluated using a scena rio 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 a pplication, a series of standardized exposure rates ( i.e. , referred to as unit exposure values 4 ) 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 acti vity individual applicators may perform ( i.e. , known as transfer coefficients 5 ) 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 guida nce used across the Age ncy 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 inpu t. 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 sev eral occasions. 10

, 11

4 http://epa.gov/pesticides/science/handler

- exposure

- data.html

5 http://www.epa.gov/pesticides/science/p ost

- app

- exposure

- data.html

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

7 http://www.epa.gov/risk_assessme nt/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.htmPage 9 of 124 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 e xposure 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 u ses 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 chlo rpyrifos 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
inf ormation 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 h ealth 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/s ap/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.htmPage 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 inhibit ion 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 sensitiv e 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 sugge st 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 pre natal 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/mechanismsPage 11 of 124 of action. I n 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 the ir 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, Cabezas

Delamare, & 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

car boxy lesterase 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 **μ** kat/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. Da ta 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, 1 994; 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 thro ughout 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 weigh t 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 protein

bound chemicals ( e.g. , chlorpyrifos) available for distribution. I ncreases 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 inhibitio n, 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 in hibition, 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 data base, 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 stron g, 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 c hlorpyrifos is the etiological causative agent in the adverse neurodevelopmental outcomes observed. In addition, t here a growing body of literature with laboratory animals (rats and mice) indicating that gestational and/or early postnatal exposure to
chlo rpyrifos 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 adulth ood.

P otential 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 st udies (Atterberry, Burnett, & Chambers, 1997; Benke & Murphy, 1975; Chanda, Harp, Liu, & Pope, 1995; Mortensen, Chanda, Hooper, & Padilla, 1996)

. R at 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) . M orten son et al

. (1996) showed that in the plasma level of CPOase 13

13 CPOase is A

- esterase (PON1) activity specific to chlorpyrifos oxon was 1/11 that of ad u lt 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, 198 5; Maxwell, 1992a, 1992b) . Fetal rats possess very littlePage 13 of 124 carboxylesterase activity (Lassiter, et al., 1998) wit h increasing activity as the postnatal rat matures, reaching adult values after puberty (50 days

- of

- age; Karanth & Pope, 2000; Morgan, Yan, Gree nway, & 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 newb orn 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

m aturation 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

matura t ion 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 childr en, 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, 2 003; 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 & Steph ens, 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 ha ve 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 detoxificat ion 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 toxicantPage 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 activ e 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 Exp**  **osure**

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 bod y 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 e xposure ( 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 p attern 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 productio n 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 chan ging 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 pr ovide 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, b oth a mode of action and an adverse outcome pathway (AOP) include a set of measureable key events that de scribe 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 ass essment. 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 level s 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 A dverse O utcome P athway (Adapted from Ankley et al., 2010)



### 3.1 **Adverse Outcome Pathway: AChE**  **I nhibition** 

**3.1.1 Initiating**  **E vent & 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 r esidue 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 ofPage 16 of 124 auto nomic 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. I n 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 limit ed 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 R BC 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 s igns 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

- effect

levels ( 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 Protectio n 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

- methy l 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 supporte d 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 ty pically 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 conduc ted (HSRB, 2009).Page 17 of 124 toxicology studies (U.S. Environmental Protection Agency, 2002, 2006) . The Agency has calculated
BMD 10 s and BMDL 10 s where the BMD 10 is the estimated dose where AChE is inhibited by 10% compared to background and the BMDL 10 is the lower confidence bound on the BMD 10 . 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 literatur e 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 t he 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 inh alation 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 hi gh 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 conduct ed 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 wh ere 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 poi nts 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

- nata l 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

- pregn ant) 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 stu dies 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 non

dietary ingestion to young children. For assessment of food, water, and aggregate dietary risk, the Agency typically conducts single day ( i.e. , acute) and chronic assessment s. For assessment of non

dietary 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 cholinesterase 15 (CCA) study (MRID 48139301) which provides brain and RBC data in PND11 pups and 2) (Moser, Simmons, & Gennings, 2 006) 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, O livier, 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 analysis 16 1 . Ta bles 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 est imates 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 BMDL 10 for RBC AChE inh ibition 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) BMDL 10 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 (BMDL 10 s 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 d osing (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 B lood 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 BMD 10 (mg/kg) BMDL 10 (mg/kg)

CCA (MRID 48139301) Female Gavage Brain A ChE 4.11 2.26

CCA (MRID 48139301) Female Gavage RBC A ChE 1.9 1.2

CCA (MRID 48139301) Female 12 hr diet Brain A ChE 4.47 3.30

CCA (MRID 48139301) Female 12 hr diet RBC A ChE 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 BMD 10 (mg/kg) BMDL 10 (mg/kg)

Moser et al., 2006 Male PND 17
G avage Brain AChE 1.89 1.54

Moser et al., 2006 Male PND 17
G avage Whole Blood
A ChE 0.62 0.43

CCA Study (MRID 48139301) Male PND 11
G avage Brain A ChE 2.13 1.53

CCA Study (MRID 48139301) Male PND 11
G avage RBC A ChE 0.82 0.65

CCA Study (MRID 48139301) Female PND 11
G avage Brain A ChE 2.18 1.56

CCA Study (MRID 48139301) Female PND 11
G avage RBC A ChE 0.96 0.75

CCA Study (MRID 48139301) Male PND 11
Milk Brain A ChE 4.4 2.4

CCA Study (MRID 48139301) Male PND 11
Milk RBC A ChE 0.47 0.36

CCA Study (MRID 48139301) Female PND 11 Milk Brain A ChE 1.42 0.91

CCA Study (MRID 48139301) Female PND 11
Milk RBC A ChE 0.5 0.36Page 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 follo wing 11

- days Repeated Oral, Gavage Doses of Chlorpyrifos

Reference Sex & Age/Duration of Dosing Endpoint BMD 10 (mg/kg/day) BMDL 10 (mg/kg/day)

CCA (MRID 48139301) Female PND 11

- 21 Gavage Brain A ChE 0.80 0.69

CCA (MRID 48139301) Female PND 11

- 21 Gavage RBC A ChE 0.17 0.15

CCA (MRID 48139301) Male PND 11

- 21 Gavage Brain A ChE 0.63 0.52

CCA (MRID 48139301) Male PND 11

- 21 Gavage RBC A ChE 0.11 0.09

Table 4

. Results of BMD Modeling of Adult, Female Rat Brain, RBC and Heart C holinesterase Inhibition following Repeat Oral Doses of Chlorpyrifos

Reference Sex, Age/Duration &
Administration method Endpoint BMD 10 (mg/kg/day) BMDL 10 (mg/kg/day)

Dow (MRID 44556901; Maurissen et al., 2000) Dams, GD6

- 20, Gavage Brain A ChE 0.65 0.54

Dow (MRID 44556901; Maurissen et al., 2000) Dams, GD6

- 20, Gavage RBC A ChE 0.06 0.03

Dow (MRID 44648101; Mattsson et al., 2000) Dams, GD6

- 20, Gavage Brain A ChE 1.1 0.8

Dow (MRID 44648101; Mattsson et al., 2000) Dams, GD6

- 20, Gavage RBC A ChE 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

- LD1 0 Gavage Brain A ChE 1.13 0.89

Dow (MRID 44648101; Mattsson et al., 2000) Dams, GD6

- LD1 0 Gavage RBC A ChE 0.050 0.044

Dow (MRID 44648101; Mattsson et al., 2000) Dams, GD6

- LD1 0 Gavage Heart ChE 0.21 0.18

CCA (MRID 48139301) Adult, non

- pregnant 11 days G avage Brain A ChE 1.03 0.95

CCA (MRID 48139301) Adult, non

- pregnant 11 days G avage RBC A ChE 0.45 0.35Page 21 of 124 As shown in Tables 3 and 4 , there is no meaningful difference between BMD 10 s and BMDL 10 s 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 co nsideration. 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 findin g supports the use of the blood measures.

There is good concordance in the estimates across the BMD 10 and BMDL 10 estimates across the adult studies evaluated in this analysis. Notably, brain AChE estimates range from approximately 0.5 to 1 mg/kg/day an d 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 BMD 10 s of 1.48 and 1.50 mg/kg/day and BMDL 10 s 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 a ssessment for chlorpyrifos. For purposes of deriving a point of departure for risk assessment, the Agency is using the BMDL 10 of 0.03 mg/kg/day for RBC AChE inhibition from the DNT study (MRID 445569 01, also called (Maurissen, et al., 2000)

.

**3.1. 3**  **Sum mary of the Dose Response Analysis** 

Acute oral point of departure is BMDL 10 = 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 d ata 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 BMDL 10 = 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 a l.

, 2000) with support from pregnant (GD6

- LD10) rats (MRID 44648101).



### 3. **2 Adverse Outcome Pathway: Neurodevelopme**  **ntal 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 chl orpyrifos 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 chlo rpyrifos, other pesticides, nerve gases) during the developmental period affects key events in nervous system development (see below). These studiesPage 22 of 124 conducted to date generally report a correlation between OP exposure and a tested effect without
considerat ion 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 infor mation 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 neuro transmission. 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, su ggest 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

neurotransm ission 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 throughou t 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 ea rly 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 s equencing 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 compo nents 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 cel ls (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 cultur es, 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 huma n 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.Page 23 of 124 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 inc luding 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 ch emical 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 t o 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, Aldun ate, & 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 c holinesterase 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 o f 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 expo sure (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 measur ed 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 differen tiation 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 chlorp yrifos, chlorpyrifos oxon, and TCPy on both neurite outgr owth 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

concentra tions (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 concentrat ions 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 axon s (Yang, et al., 2008) . Similar results were observed, with both chlorpyrifos and chlorpyrifos oxon decreasing axonalPage 24 of 124 outgrowth at concentrations below those that inhibited AChE activity. To establish whether the target of chlorpyrifos and chlorpyrifos oxo n 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 req uired 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 conversi on 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 inferr ed 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 neuromus cular 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 zeb rafish 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 f rom 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 tha t 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 chl orpyrifos on gestational days 17

- 20. In a previous study from the same laboratory this exposure paradigm resulted in a non

significant 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 d uring 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 ACh E is a plausible adverse outcome pathway for chlorpyrifos, a number of questions remain. There is substantial evidence for aPage 25 of 124 morphogenic role of AChE in nervous system development distinct from its role as an esterase to hydrolyze acetylcholine. In vitro e vidence 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 di rect 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 i s likely involved in brain morphogenesis (Hohmann & Berger

Sweeney, 1998) . While cholinesterase inhibitors can affect cholinergic signaling by inhibition of the catalytic activity of AChE and subsequent increase in acetylcholine, some, including chlorpyrif os 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 in hibition

.

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 sit es. 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 a n 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)

. Subs equent 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

- D onia, 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

corre lation 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 bel ow concentrations that inhibited AChE (Huff, Corcoran, Anderson, & Abou

- Donia, 1994) ; specifically, c hlorpyrifos 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.Page 26 of 124 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 G

protein (G i ) and activation inhibits adenylate cyclase and decreases cAMP formation. In c ontrast, m1, m3, and m5 receptors increase PI hydrolysis via the stimulatory G

- protein G p (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 IC 50 for inhibition of cAMP of 155 nM. U nlike 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 agoni st 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 pri marily 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 embr yonic 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 i n 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, Guth rie, & 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 chlorpy rifos 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)

. T his binding , however, occurs at relatively high concentrati ons (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 rece ptor 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 re ceptors. 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 inPage 27 of 124 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 syste m 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. Chl orpyrifos 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 d evelopment, 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 Campolo ngo, 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 impaire d 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 r eceptor 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; Se gall, 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 severa l 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 ha s not been adequately studied. Thus, prolongation of the ligand residence time, plus blockade of ligandPage 28 of 124 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 syndrome s ( 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 inhibit ory 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 als o 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 correlat e 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, trea tment 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 r esearch 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 influe nce 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.
Somew hat 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 a l. also measured chlorpyrifos oxon

induced 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% inhi bition 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.Page 29 of 124 In summary, most of the research i n 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 a re 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 develo pment 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
actio ns 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 potent ial) 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 cellula r 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 c ells 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 fo r 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 (Q iao, 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 signifi cant 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 rese archers 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 wa s 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 significan t 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 chlorpyrif os oxon was found to be active in cerebellar granule cells but not in PC12 cells. Inhibition of AChE was not determined in these studies.Page 30 of 124 A limited number of studies have examined the oxidative stress response in rat brain after in vivo exposure to chlorp yrifos. 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 fo llow

- 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 chlorpyrif os 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 e xposure 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 exami ned 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 evidenc e (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 neurodevelop ment 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, aggres sion, and respiration. While the majority of the receptors are either negatively or positively coupled to adenylate cyclase, the 5

- HT 3 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. Var iousPage 31 of 124 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 & S tanwood, 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 i nvolved 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 labo ratory 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, 200 9) . 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 devel oping 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 syst em 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 p roduced 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 somatose nsory 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, 200 5; 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 lo ng

- 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 (Kling er 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; Slotk in & Seidler, 2007b, 2007c) , serotonin turnover (Aldridge, Meyer, et al.,Page 32 of 124 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, 200 7b) 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 administr ation 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

- HT 1A and 5

- HT 2 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 br ain. 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
cholines terase 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 ner vous 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 pro duce 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 developm ent of the serotonergic nervous system (Aldridge, Meyer, et al., 2005a; Aldridge, et al., 2004)

.Page 33 of 124 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, a s 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 i ts 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 ex tension/retraction, as well as growth cone advancement. In addition to these integral roles in brain structure and growth, micr otubules and the microtubule associated motor proteins kinesin (Hirokawa & Noda, 2008) and dynein (Vallee, Williams, Varma, & Ba rnhart, 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 s ystem, 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, 198 9; 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; Fukushi ma, 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 anterog rade and retrograde axonal transport in the sciatic nerve 6 days after cessation of treatment 2003 (Terry et al., 2003) . This samePage 34 of 124 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, Bu ccafusco, 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
ki nesin

- 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 in teresting 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, th ereby precluding inhibition assessment. A recently published in vitro paper (Middlemore

Risher, Adam, Lambert, & Terry, 2011) assessed the effect of chlorpyrifos or chlorpyrifos oxon transport of mitochondria in rat cortical neurons. They found dose

- relate d 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 mov ement. 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 acti vity 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 s eparated 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 chlorpy rifos oxon

- treated brains were missing 6 microtubule associated proteins. This observation was supported by the nanoimaging showing that microtubules from chlorpyrifos oxon

treated 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 do ses 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,Page 35 of 124 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 ox on, 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 dynamic s, 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 polymeri zation 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 afte r 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 extensi ve 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; Middlemore

Risher, 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) t hat 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, di fferentiation, 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 a ction 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 construc tion 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 surpr ising 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 be tween exposure and neurological disorders in humans for which a definitive mode ofPage 36 of 124 action has not been established: for example lead, methyl mercury, and ethanol (Alfonso

- Loeches & Guerri, 2011; Castoldi et al., 2008; Farina, Rocha, & Aschner, 2011; Johans son 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 mult itude 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 (Watanab e 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 Outc omes 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 Agen cy 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 o f 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 f rom, 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 deve lopment.

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 e xamined at some time after weaning. That is, papers reporting evaluations shortly after birth or during the pre

- weaning period do not reflect long

term consequences and may also be confounded by AChE/ChE inhibition during concurrent or recent exposure; th us, 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 cholinerg ic 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; Ca rr 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,
20 09; 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,Page 37 of 124 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 appar atuses. 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
consid eration, 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 labo ratories, 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 a nd direct postnatal exposures in the same subjects, andPage 38 of 124 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 s uch

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; Abou

Donia 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 Li é ge, 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 vari ous 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 significa nce, 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 di rect 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 ra ther 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 attribut e 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 foll ows 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. The se 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 bee n 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 loca ted 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 seri es 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 alter ed 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 ful ly 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

- 2 0), 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 in creased in males (worse performance) but decreased in females (better performance) (Aldridge, Levin, et al., 2005; Levin, et al., 2001) . OnlyPage 40 of 124 the late postnatal exposure (PND11

- 14, 5 mg/kg/d) was without effect on this measure (Levin, et al., 2001)

.

Se veral 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/k g/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 o r 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 high

dose rats sho wed 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 c omparisons 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 (pos itional 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 learn ing (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 o f changes, with somewhat greater magnitude of differences, indicating dose

response. 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 W ater M aze

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 ex posure (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 addi tional 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 findi ng across several studies, albeit in the same laboratory.Page 41 of 124 These recent findings, plus another study available only in Chinese 17 , 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 descr ibed 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 o n 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 rec ently 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 cognit ive 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 ot her 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, dose

response 6 Cognitive deficit

- rat, F not M, no dose

- response 4 Cognitive deficit in M, improved function in F

- rat 3 Cognitive deficit in M, improved function in F

- rat 7 No effect

– rat
, M&F 3 Cognitive deficit in M, improved function in F

- rat, dose

- response 9

Morris Water Maze Cognitive deficit

- mouse, M&F, no dose

- response 8 Cognitive deficit

– mouse, M&F 11 Cognitive deficit

- rat, M&F, dose

response 2

Foraging Maze Cognitive deficit in F, improved function in M

– mouse, dose

response in F not
M 10

T

- maze Spontaneous Alternation No effect

– rat, M&F 6 No effect

– rat, M&F 4 No effect

– rat
, M&F 3 No effect

– rat
, M&F 3

Delayed Spatial Alternation No effect

– rat, M&F 1

Passive Avoidance No effect

– mouse, only M tested 5 No effect

– mouse, only M tested 5

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., 2011Page 43 of 124 3.2.2.2.2 Anxiety/ E motion

Anxiety and despa ir/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 assess ed 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 me asures. 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 beha viors 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

g roup 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 e ntries, supporting a conclusion of increased anxiety, but only at the middle dose. The lack of dose

response 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 differe nce 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 a s 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 p ups 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 als o 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 n ot 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 cho colate 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,

ove rall there is evidence for long

- term changes in anxiety behavior following chlorpyrifos exposure, as shown in Table 7

.Page 45 of 124 Table 7 Summar y of Anxiety/Emotion O utcomes 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&F 1 Decreased anxiety

- mouse, M not F, no dose

- response 5 Decreased anxiety

- rat,
M not F 2 Decreased anxiety

- mouse, F not M, dose

response 5 Increased anxiety

- mouse, only F tested, no dose

- response 3

Light

- Dark box (anxiety) Increased anxiety

- mouse, F not M 7 Decreased anxiety

- mouse, dams 6 Increased anxiety

- mouse, only F tested, no dose

- response 3

Despair No effect

– mouse, M&F 7

Novelty/preference Decreased anxiety

- rat, M&F 2
No effect

– mouse, M&F 4 No effect

– mouse, M&F 4

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., 2010Page 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 fo llowing 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:f emale 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) agonisti c 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 beha vior, 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 we re 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 agg ressive 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 Be havior

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 (Riccer i, et al., 2003; Ricceri, et al., 2006) . None of the newer studies addressed male behaviors. These findings are summarized below ( Table 8 ).

Summary

Overall, decr eased 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 convinc ingly 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, dose

response 3 No effect 1 Increased solicitation, dose

- response 1
No effect 3
No effect 4

Male:male social Increased aggressive postures, dose

response 2 Increased agonistic behavior, differs
across time

- course, dose

- response 1 Increased solicitation, dose

- response 1 Increased agonistic behavior, over time, no dose

- response 1 Increased attack behaviors, dose

response 2
No effect social 4,a

Induced maternal No effect 2 Increased maternalism, dose

response 2

Natural maternal Decreased maternalism 4

Natural maternal:male Decreased aggression 5 Decreased defensiveness 4

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., 2010Page 48 of 12 4 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 late ncies 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, dec reased, 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. , explor ation, 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 be en 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 wit h 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 i nterpreted 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 gest ationally 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 laborator ies. 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 postnat al, 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 thePage 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&F 6 Increased activity

- mouse, only M tested, dose

response 13
No effect

- mouse 8
No effect

– rat, M&F 10 No effect

– rat, M&F 11
No effect

– mouse, only F tested 2 No effect

- mouse, M&F 12 Decreased activity

- rat, M not F 4
No effect

– rat, M&F 9 Increased activity

- mouse, only M tested, dose

response 13 Increased activity

- mouse, M&F, dose

response 12

Increased activity

- rat, M&F 4
No effect

– rat, M&F 9 Decreased activity

- rat, M&F, dose

response 3

Activity

measures Increased activity

- rat, M&F, dose

response 6 No effect

- mouse, only F tested 14
No effect

- mouse, M&F 5 Increased activity

- rat, M&F, DR 10
No effect

- mouse, M&F 13 No effect

– mouse, only F tested 2 Increased activity

- rat, M not F 1 Increased activity

- mouse, M&F 12 Decreased activity

- rat, M not F 9 Decreased activity

- mouse, M&F 14 Decreased activity

- rat, M not F 9 Increased activity

- mouse, M&F 12
No effect

– mouse, M&F 13 No effect

– rat, M&F 7

Habituation

rate Increased habituation

- rat, M&F, dose

response 6 Decreased habituation

- rat, F not M, dose

response 10
No effect

– mouse 8 No effect

– rat, M&F 11 No effect

– rat, M&F 9 Decreased habituation

- rat, M&F 9

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., 2008Page 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. F ew of the se papers assess AChE inhibition at all, much less at time points potentially critical to the ultimate outcomes, e.g. , within hours after dos ing 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. Env ironmental 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 ea rlier 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 her e 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 t he 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 mat uration. 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 do sing 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 subse quent 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 asympt otic 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 studiesPage 52 of 124 r eporting 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 s ignificant 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 ab errant 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), t hese 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 ar e 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 assess ment. 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. Dose

response is not always evident, since many studies only use one dose, and of those using two or more doses, there is not always a monot onic 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 vehic le, 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 wi th 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, som e 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 labo ratory 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 pr oduce some degree, however minimal, of AChE inhibition. Thus it is not possible to know whether effects would be present at lowerPage 53 of 124 doses, since they have not been adequately studied; thus far, only one study (Braquenier, et al., 2010)
has tested a dose low er 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 trac k possible mechanisms from early initiating events to the final neurological outcome. Such studies represent longer term research efforts by the different laboratories.Page 54 of 124 

## 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 thr ee 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 Develo pment 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 meeti ng, 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 bir th 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 (pr ospective 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, ana lyzed, 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 environmen tal 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 c ohort 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 New born 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 sugge sted 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 tha t random

18 See Meeting Minutes at: http://www.epa.gov/scipoly/sap/meetings/2008/september/sap0908report.pdf.Page 55 of 124 or sys tematic 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 m ode 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 expos ure on children’s environmental health in con junction 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 exte nd 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 investi gations 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 Develop ment 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 measure s in young ch ildren as modified by paraoxonase

- 1 ( PON1 ) genotype and phenotype (Eskenazi et al., 2010; Harley et al., 2011)

.

Re searchers have also performed epidemiologic methods research which in many ways reduces uncertainties related to key measures within these s tudies. 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
resid ential 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 co mbination 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 ep idemiologic 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 analy sis 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 s ection 4.3 reflects EPA’s synthesis and evaluation of the current chlorpyrifos epidemiology database. In accordance with the draf t “ 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#materialsPage 56 of 124 **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; E ngel 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 populat ions 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

- containi ng pesticide products on December 31, 2000 , during the time period of these cohort investigations . In the agri cultural setting, chlorpyrifos wa s 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 ex perienced 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 recr uited 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 c ompound 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 measuresPage 57 of 124 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 mater nal 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 basi c 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 pr ivate 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 t wo 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 us age, 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 measur ement 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 PON 1 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 c oncentration 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 bi phenyls (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, CHAM**  **COS (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.cdp r.ca.gov/docs/pur/pur01rep/01_pur.htm) ), chlorpyrifos, however, was notPage 58 of 124 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 wit h 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 gestatio n, 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.

Da ta 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, participant s 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 ma ternal urin e 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 remaind er 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 epidemiol ogic 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, a utho rs 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 i nfant 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 ch aracteristics, 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 o f prenatal chlorpyrifos exposure assessment likely played a role in this observation (Needham, 2005)

.Page 59 of 124 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 IndexPage 60 of 124 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% C I:

– 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 inve rsely 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 chlorp yrifos exposure explains th e 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 ris k 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 lev el. 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 significan tly 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

asso ciated 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 pestic ide exposure (Eskenazi et al., 2004) . On the contrary, investigators reported positive associations between birth length and head circumference associated with non

specific organophosphate exposure measures (DAPs). Researchers observed decreases in gestat ional

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 chlorpy rifos. Authors did not observe an association using TCPy as a measure of chlorpyrifos exposure; however they did report an increased riskPage 61 of 124 of preterm delivery with decreasing cholinesterase concentrations, a third biomarker of organophosphate exposure. To i dentify “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)

u sing 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 modify ing 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 as sociation 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 PON 1

- 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 i nteraction =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 neonata l 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 differ ent 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 condi tions. 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 eval uation, 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 fo r 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. Notab ly, 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 evi dence 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 refl exes were significantly related to DAP and DMP levels; the risk estimate for DEP and the number of abnormal reflexes within the lowestPage 62 of 124 PON1 level was not significant. In the CHAMACOS cohort, Young et al. (2005) observed a statistically significant associat ion 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 uri nary 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 ad dition, 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 t oddlerhood (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 ei ther 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 significant ly 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 exp osed 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 chlorpyri fos 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

prena tal 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 author s concluded that their findings arePage 63 of 124 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 throug h 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

- e xistent 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 matern al 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

- fo ld 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
d isorder (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 t he 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 yea rs. 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 assoc iated with scores on the K

CPT ADHD Confidence Index > 70th percentile (OR = 5.1; 95% CI: 1.7

– 15.7 and with a composite ADHD indicator of the various measures (OR = 3.5; 95% CI: 1.1

– 10.7). Some outcomes exhibited evidence of effect modification by sex, wit h 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 leve ls, and attentional difficulties at age 5 years using three different measures of this neurodevelopmental outcome.Page 64 of 124 **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. T hese 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

- speakin g 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 interv iew 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 chlorpy rifos 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 co gnitive/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. Am ong 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 associat ion 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 low est 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), polychlo rinated 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 yea rs 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.Page 65 of 124 **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 cons ider. 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 asce rtainment 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 socio

economic 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 prec ision 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 ro le 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 pop ulations 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 E PA and National Institute of
E nvironmental Health Science (NIEHS) Children’s Environmental Health and Disease Prevention Research
Centers 20

. Cohort studies were comprised of several hundred mother

- infant pairs (range: n=102 to n=488) and w ere likely sufficiently statistically powered t o 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 sensiti vity 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 estimat ing exposure during the critical window of d evelopment.

Selection criteria were clearly defined and appropriate across the three cohorts. Investigators limited participants to indi viduals 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, autho rs 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. T he 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/Page 66 of 124 However, as detailed in Section 5.0, comparison with biomonitoring levels in the U.S. general popul ation, 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 mor e, 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 measurement s 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 (unmeasur ed, future) health

outcome.

Researchers utilized similar tools to measure adverse health outcomes across the studies. The use of similar outcome measures across studies aid s comparability of studies and assessment of the consistency of results. Neurodeve lopmental 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, neuro development al 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 adju dicators, 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 be havior 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 f or 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 a nd 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, w hile other organophosphates are metabolized to DMP. For risk assessment purposes, i t 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 i nconsistencies across study findings , as noted by others (Needham,Page 67 of 124 2005)

. The three birth cohorts recruited participants who likely have chlorpyrifos exposures that are high er 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 childre n’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 met abolite (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 reporte d 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

- sam ples 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 repres entative 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 n ot 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 accur ately 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 th at 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 cancell ation. 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

- subjec t 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 TCP y (p = 0.001) in this study. However, these researchers alsoPage 68 of 124 demonstrated the presence of TCPy in meconium; meconium is fecal matter accumulated in intestines of developing fetus from week 1 3 of gestation and released immediately prior to or within a few d ays 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 collect ed 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 pe riod 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 meconiu m 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)

. Add itionally, the correlation between the analytic measure of exposure and a measure reflective of exposure during a large proportion of the gestational period suggest s 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

- differ ential 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 m easured confoundi ng 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 neurodevelopm ent ( e.g. , lead exposure, mercury exposure, air pollution, tobacco smoke exposure, maternal alcohol intake, an d exposure to other pesticides). The selection of cohorts that are homogenous with respect to many demographic and social factors significantly re duces 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 c onfound 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 Newbo rn 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 techniqu es 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 adjust ed. 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 confoundingPage 69 of 124 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 collect ed 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, author s 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 (Ba yley 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 publis hed 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 P DI, 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 varia ble 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. Non e 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 potent ially 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 i t 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 obt ained 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 co hort 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 dat a, 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 i t is difficult toPage 70 of 124 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 confound ing 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 actu al 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, th e 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, t he three children’s health cohorts considered herein have several strengths as well as limitations to con sider 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

- e stimation (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 deliber ations 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 exposur e 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 o f 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 ha ve 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 eff ect 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
associati on. F actor that may lead to false positive associations include unmeasured or poorly meas ured 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 associ ations 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 exp osure 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 whic h 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 neurodevelo pmental 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 st rong design, conduct and analyse s utilized in these investigations. While factors are present across these studies which may have le d 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 over

estimation of effects, i.e.

, po sitive confounding bias as a result of poorly measured factors related to
socio

- economic status and other environmental chemical exposure s.



### 4.3 Chlorpyrifos Epidemiology Synthesis and Evaluation

In this chapter, EPA reviewed the results of epidemiologi cal 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 acti on which may explain a causal role for chlorpyrifos in adversePage 72 of 124 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 effe ct, 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 neuro developmental outcomes in the three cohorts. Within this database, moderately strong associations have been observed, and in many instances measured with precision.
T he 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 intel ligence 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, EP A 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 chlorpyri fos and attentional problems, pervasive developmental disorder, and ADHD

like 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 di fferent 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 newbor n 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 MD I 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 d ecrease 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 g rowth. 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 statu s seemed to enhance these positive associations, such that in the non

- susceptible groups ( i.e.

, PON1

- 108 CC, PON1 192 RR, and high arylesterase activity), increasing DAP concentrations were significantly associated with increased birth weight and head circu mference. 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 co ncentrations 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 P ON1 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 in creasingly 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 cha nges 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 ar ticles. 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 ha ve sufficient power to detect departures from linearity in the shape exposure

outcome 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 evaluatePage 74 of 124 the shape of dose

- response curves, test the linearity assumption, and investigate potentia l 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 af ter 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 chlorpy rifos in maternal air samples were significantly lower after the voluntary cancellation (4.9 ng/m 3 ), relative to those taken before 2001 (8 ng/m 3 ; 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 significan t (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 experime ntal regimen

, ” as advocated by AB Hill in the assessment of causality in observational studies

. Changes in exposure were re flected in decreasing chlorpyrifos levels, and appear to have been associated with changes in developmental outcomes. However, as not ed 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 generaliza bility 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
urinar y 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 finding s 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 wit h a substantially different exposure

range. In the studies reviewed, the exploration of gene

- by

- environment, and phenotype

- by

environment 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.Page 75 of 124 

### 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 method ological 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 consideration s, 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 w ould 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, e xposure measurement error has been shown to mo re 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 co horts 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 iss ues 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 chlorpyri fos 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

neur ological 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 t his paper.

In summary, while the strengths and limitations of the studies would be more likely to lead to an under

estimation 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#materialsPage 76 of 124 The temporal association, stren gth 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 domain s , 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.Page 77 of 124 

## 5.0 **Exposure Profile & Biomonitoring Research**

In the previous Section , the Agency performed a targeted evaluat ion 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 a nd 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 involvin g 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 betw een 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 C**  **onsiderations**

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 m edia.

Table 11 Common C hlorpyrifos B iomarkers of E xposure. Biomarker/Analyte Biologic Matrix Populatio n 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 varia bility; 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 chlorp yrifos exposure in exposure studies and epidemiologic research. Several review articlesPage 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 detai led 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 obser vation 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. D EP and DETP are considered non

- specific biomarkers because they are metabolites of multiple of OP pesticides, including c hlorpyrifos

- methyl , coumaphos, diazinon, disulfoton, ethion, parathion, phorate, sultotepp, and terbufos (CDC, 2009). As such, it is m ore 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, chlorpyrifo s

- 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, studie s 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

measure ments 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 ex tent 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 in terpreting 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 singl e 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 p esticide 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 invest igators 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).Page 79 of 124 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
wh y 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 qua ntities 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 suggest ed 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 Columbi a 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 gene ral 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/o r 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 con centrations 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 adult s and African

- Americans generally have more muscle m ass 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 ( US EPA, 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. (199 9) 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.

Wh ile 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 observa tional 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 Pesticid es 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 Expo sure 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, 2 006) 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 resident ial 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 ran ge 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 take

home 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 co nventional 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 Un iversity 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 st udy populations are provided in Figure 4 and Figure 5 , respectively. These figures provide information on the range of observed TCPy levels ba sed 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 b ox length indi cates 25 th and 75 th percentiles, central line indicates median, and whiskers indicate 5 th

and 95 th percentiles . However, several studies only reported central tendency values and the range of observed values. In these cases, study summary data are rep resented 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 comparis on of these survey periods, 50 th percentile TCPy levels were 2.8 and 3.1 μg/L, respectively. Likewise, 95 th 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 previousl y 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 NHEXAS

AZ. These median levels were roughly 2

- 4 times higher than the median TC Py 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 resul ts 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) we re 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 replaceme nt 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 re

introduction of conven tional 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

- 2 002, respectively (Hill et al., 1995; CDC, 2009). Based on comparison of these three different survey periods, 50 th percentile TCPy levels were 3, 1.5, and 1.9 μg/L, respectively. While this suggests that 50 th percentile concentration may have declined r elative to 1988

- 1992, the 95 th 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 95 th percentiles in all three NHANES survey periods. On the other hand, TCPy levels in the registrant

- submitted worke r 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 th e 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 l evels 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 l evels TCPy levels of pregnant women in NHANES 2001

- 2002. Berkowitz et al. (2003) of Mount Sinai also observed median TCPy level s that were higher than NHANES. In particular, Berkowitz et al. observed the median TCPy level of their cohort was 7.6 μg/L, wh ich was higher than the median TCPy levels observed
in NHA NES 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

gen eral 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 c ollected, 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 C omparison of TCPy B iomonitoring S tudies I nvolving C hild S tudy P opulations. a

a Boxplots are presented in which the box length indicates 25th and 75th percentiles, central line indicates median, and whiske rs indicate 5th and 95th percentiles. However, several studies only reported central tendency values and th e 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 1000NHANES 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=51Farm 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 GMPage 84 of 124 Figure 5 Summary of C omparison of TCPy B iomonitoring S tudies I nvolving A dult S tudy P opulations. a

a Boxplots are presented in which the box length indicates 25th and 75th percentiles, central line indicates median, and whiske rs indicate 5th and 95th percentiles. However, several studies only reported central tendency values and the range of observed va lues. 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 1000NHANES 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=34Farm families, Spouses pre-applFarm 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, 2010TCPy Urinary Concentration (μg/L) Reference Range Maternal Cohorts

GM GM GM GM

GM GM GMPage 85 of 124 Table 12 Summ ary R esults 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 95 th 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

— post

appl 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

— post

appl 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

— post

appl 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, Spring

Summer Spot

- Longitudinal 45 89 0.5

- 3.3

- - - - - - -

Farmers, not sprayed 2001, Spring

Summer Spot

- Longitudinal 45 89 0.5

- 3.5

- - - - - - -

Farmers, self

- sprayed 2001, Spring

Summer 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 95 th Max

Curwin et al.,
2007 Father, nonfarm 2001, Spring

Summer Spot 23 94 3.32

- 13 3.8

- - - - - 47

Father, farm 2001, Spring

Summer Spot 24 100 3.32

- 17 6.5

- - - - - 58

Mother, nonfarm 2001, Spring

Summer Spot 24 95 3.32

- 11 1.8

- - - - - 35

Mother, farm 2001, Spring

Summer Spot 24 100 3.32

- 14 5.6

- - - - - 52

Child , nonfarm 2001, Spring

Summer Spot 51 100 3.32

- 15 5.4

- - - - - 54

Child , farm 2001, Spring

Summer 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 Postnat al 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

E skenazi et al.,
2007 CHAMACOS 1 st Preg. 2000

- 2003 Spot 442 71

- - - - - - 3.76

- - - CHAMACOS 2 nd Preg. 2000

- 2003 Spot 419 82

- - - - - - 4.6

- - -Page 87 of 124 

### 5.3 **Chlorpyrifos Biomonitoring and AChE Inhibition**

M onitoring of ChE has been performed in both research and state

- based occupational surveillance to

m easure 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. I nhibition 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, plas ma 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 ACh E/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 a nd 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 an d 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 i ntentional 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
als o 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 eval uation, 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 betweenPage 88 of 124 urin ary 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 dur ing 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 comp arison 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 U rinary TCPy levels (μg /g creatinine) reported by
Galabrant et al. (2009) and Farahat et al. (2011) with the General U.S. P opulation.

Population Group Time Period n TCPy (μg /g creatinine)

Average Median (95% CI) 95 th 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 repor ted 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 thei r 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 activit y 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 t hat 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

represen ting 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 p oint 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 bet ween 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. h ad 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 d id 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 betwe en absolute AChE activity and TCPy. Limitations were also noted by the i nvestigators. Galabrant et al. , for example, reported that there study population only included workers and did not include potentially sensitive populations, such as children or el derly. 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 a ppropriate 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**

T here is a need in environmental exposure assessment to qua ntitatively 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 uri ne. 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 A gency will be soliciting comments on approaches to best int erpret 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 ca n 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 approa ch. 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 th e 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

- Compar ative: 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 ef fects (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 pharmacoki netic 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 excr eted into the urine as TCPy. This methodology is more fully described in the example biomonitoring study report (Honeycutt and Day 1994).Page 91 of 124 **•** 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 chlorpyrifos

oxon as intermediate) and also happens to be an environmental degradate to which human direct exposures take place. Therefore, caution should be exercised in the
i nterpretation 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 resulti ng 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 measure ments 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 othe r approaches that do not use creatinine normalization

, in the analysis of biomonitoring data associated with triclosa n ( RED 2008 ). A major limitation of the Mage et al. ( 2004 ) approach is that the human dose estimate does not account for route

- specific b ioavailability 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 A pproach: 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 (in cluding 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 PBP K model available for the AChE inhibiting mode of action of chlorpyrifos that provides the critical internal dose metrics associated with AChEPage 92 of 124 inhibition. The model was developed based on adult non

- pregnant rats and was then scaled and calibrated for the h uman 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 lifestage

spe cific 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 lifestage

dependent 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 route

specific information

- Assumes that analytes are derived from a single parent c hemical

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) - Doe s not address all susceptible lifestages

- Does not incorporate all relevant routes of exposure

Ch lorpyrifos 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 conside ring approaches to best evaluate the current biomonitoring information and is soliciting input from the SAP. T here is no ideal reverse dosimetry analysis approach. Each approach needs to be carefully considered for its strengths and limitations . Table 14 p rovides a brief syn p osis of several approaches discussed above. The following section 5.4.2 provides a more extensive background of the PBPK

- PD model available for chlorpyrifos.Page 93 of 124 **5.4.2 Physiologically Based Pharmacokinetic** 

**- Pharmacodynamic**  **Model for Chlor**  **pyrifos : 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 avai lable (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 mechani stic mode of action information that chlorpyrifos exerts its cholinergic toxicological effects through metabolic activation to chlorpyrifos

- oxon and detoxi fication 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 inh ibition 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 detoxif ication 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 ref inement 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 robus t 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)

. A lthough 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 th e 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

- ap propriate 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

resu lting from peri

- natal exposures to chlorpyr ifos . Epidemiological studies have also supported associations between pre natal exposures to OP s including chlorpyrifos, and cognitive delays in children

.Page 94 of 124 A t 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 pr egnancy

- 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 (Timcha lk 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 i n 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

- to

mouth)(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 e ffort is further behi nd 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 adul t 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 prod uce a robust PBPK model that can be used for human health risk assessment applications (as opposed to for AChE

- inhibition mode of action )

. M odeling future effort s aimed PBPK modeling for chlorpyrifos needs to include in utero exposures along with dermal a nd/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).Page 95 of 124 **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 als o 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 contra st, when chlorpyrifos was administered to adult non

pregnant 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 i s 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 r outes 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 sup ported 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 cha nges 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 limita tions 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 othe r 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 physiologica l and chemical

specific 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). T his 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 curre ntly available for chlorpyrifos.Page 96 of 124 

## 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 a nimal

, 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/adve rse 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, his torically, 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 rese arch 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 determi ne 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 scientif ic 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 i nvolved 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 f rom 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 chlorpyrif os 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 con fidence 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 facto rs 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 oftenPage 97 of 124 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 move s 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 SA P 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 o f 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 p ending 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 new est 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 syst ems, PND11 rat pups provide the most sensitive cholinesterase inhibition data for deriving the acute oral POD (BMDL 10 0.36 mg/kg). As the metabolic systems mature at or near adult levels, this age

- dependant sensitivity disappears. In repeated dosing stu dies, 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 (BMDL 10 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 chlorpyrifo s 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 includ es 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 neurol ogical 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 (Br aquenier 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 poi nts 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 dosePage 98 of 124 level produced significant effec ts. Comparing a NOEL of 0.2 mg/kg/d to a repeated dosing

BMDL 10 of 0.03 mg/kg/d suggests that A ChE inhibition is a sensitive and protective endpoint; however, ideally studies that thoroughly evaluate neurodevelopmental effects at lower, non

inhibitory do ses c ould 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). T he 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 pu rsued 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 establis hing 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 b inding of chl orpyrifos/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 r eport on Toxicity Testing in the 21 st 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 i n 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 req uires 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 chlorpyrifo s/c hlorpyrifos oxon binding to AChE, the inhibition of ne urite 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 s ystem 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

- r esponse 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.Page 99 of 124 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 prelim inary 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 wh ich 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 coho rt 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 explanation s 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 chlorpyrifo s 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 re lation 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 thro ugh 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 measur ement of both neurodevelopmental outcomes and

exposure most likely occurred in these studies. While the magnitude of this error is difficult toPage 100 of 124 quantify, this source of error would unlikely be related to the either the outcomes or exposures measured in the se 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 presen t 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). Considerin g 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 r eviewed 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

ex posure, 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; h owever, 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 literat ure 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 ac tual 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 l ess 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 C olumbia 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 ca ncellation 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. OnePage 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 so phisticated 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 inh ibition 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 i nterpretations

. 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 informat ion 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 BMDL 10 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 postu lated 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 chl orpyrifos 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 ot her (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 wo uld 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 e stablished for neurodevelopmental outcomes. T his growing body of literature does demonstrate, however

, that chlorpyrifos and/or its oxon are biologically active on a number of processes that a ffect the developing brain. Although these studies are not app ropriate 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

outc ome pathway complicate s 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 b y 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 . T he question posed regards the nature and d egree of uncertainty around points of departure base d 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 s ome effects which are similarly sensitive or more sensitive than AC hE 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 o utcome p athway/mode of action makes quantitative use of the epidemiology study in risk assessment challenging, particularly with respect to dose

- response, cr itical d urat ion of exposure, and window(s) of s usceptibility. 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 e valuate 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 ne urodevelopmental 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 . Pre ferably 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 additio n 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 w ill solicit comment on the further analyses that may be conducted to further inform the potential for long

term 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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