USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

APPENDIX C 

INFORMATION ON CHEMICAL MIXTURES AND THEIR COMPONENT CHEMICALS 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Introduction 

This appendix summarizes data on chemicals considered for examples in the present analysis. These 

data can be used in a component-based approach to cumulative risk, particularly in developing an 

understanding of toxicokinetics relevant to determining the likelihood of an interaction. For the data 

contained in this appendix, several points bear emphasis. 

1. No data on either of the two example chemical mixtures was located. Thus, this report contains 

information only on each of 10 components chemicals (i.e., 6 component chemicals for Mixture 1 

and 4 component chemicals for Mixture 2). 

2. This summary of data is by no means an exhaustive review for each of these 10 component 

chemicals. Only brief information is provided for each of the sections for each chemical. We 

placed a great deal of emphasis on not repeating the effort of existing reviews; rather, we provide 

an update on each chemical with current information. 

3. We focused more on pharmacokinetics/pharmacodynamics with special emphasis on 

physiologically-based pharmacokinetic (PBPK) modeling. 

In this report, the toxicological and pharmacokinetic characteristics of ten chemicals are discussed. The 

ten chemicals consist of two groups that can potentially form mixtures in drinking water. The first 

mixture consists of the organophosphorus pesticides parathion, methyl parathion, chlorpyrifos, diazinon, 

fenthion, and fenitrothion. The second mixture consists of the chlorinated hydrocarbons chloroform, 

trichloroethylene, tetrachloroethylene, and 1,1,1 trichloroethane. Each chemical is discussed separately 

in one section of the report. In each section, a brief description of the toxicology of the chemical is 

provided as a background to the selection of appropriate dose metrics for risk assessments that can be 

quantified using PBPK modeling. Subsequently, available data describing the pharmacokinetics (PKs) 

of each chemical in laboratory animals and humans is provided. Finally, available studies regarding 

potential PK interactions between the chemicals are provided. 

The studies incorporated in this review are necessarily limited. The review is based on a detailed search 

of the open literature. However, inevitably there are additional studies to be considered, especially those 

that are not published. 

The principal purpose of this review is to compile data that may be useful in performing a PBPK modelbased cumulative risk assessment (CRA) for the two groups of chemicals in drinking water. 

Literature Cited

Yang, R. S. H. (2004). Final Report and Final Approach. A Report to USEPA under USEPA Contract 

No. 3C-R102-NTEX. May 24, 2004. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Parathion 

1.0 Introduction

Parathion (O, O-diethyl 4-nitrophenyl phosphothioate) is a phosphorothionate insecticide that has no 

registered uses in the U.S. but is widely used elsewhere in agriculture and is present in food and 

environments (Brack et al. 1999; Fenske et al. 2002; Leblanc et al. 2000; Lifshitz et al. 1999; Ripley et 

al. 2000; Simcox et al. 1995). 

1.1 Toxic effects 

Parathion exerts its toxicological effects via inhibition of acetylcholinesterase (Nigg and Knaak 2000; 

Thiermann et al. 1997). Metabolism of parathion exploits CYP450 as a metabolizing enzyme 

(Atterberry et al. 1997; Attia 2000; Attia et al. 1995; Besser et al. 1993; Halpert et al. 1980; Halpert and 

Neal 1981a, b; Howard and Pope 2002; Jett et al. 1994; Katz et al. 1997). Other toxicities such as 

reproductive toxicities, immunotoxicity, cytotoxicity, carcinogenicity, and other effects have also been 

shown (Bustos-Obregon and Diaz 1999; Bustos-Obregon et al. 2001; Cabello et al. 2001; Cao et al.

1999; Carlson and Ehrich 2001; Carlson et al. 2000; Galloway and Handy 2003; Grellner and 

Glenewinkel 1997; Ivens et al. 1998; Levario-Carrillo et al. 2001; Li and Zhang 2001; Liu et al. 1999; 

Melendez Camargo and Lopez Hernandez 1998; Olivier et al. 2001a; Olivier et al. 2001b; Padungtod et 

al. 1999; Padungtod et al. 1998; Padungtod et al. 2000; Rojas et al. 1998; Saleh et al. 2003; Segura et al.

1999; Selgrade et al. 1984; Senel et al. 2001; Tong et al. 1988; Undeger et al. 2000; Van Den Beukel et 

al. 1997; van den Beukel et al. 1998; Wagner et al. 2003; Zaidi et al. 2000). 

1.2 Pharmacokinetics

There are a number of pharmacokinetic studies of parathion and its toxic metabolite, paraoxon, 

conducted both in non-mammalian and mammalian species such as mice, rat, pig and dog via many 

routes of exposure including intravenous, oral and dermal exposure (Braeckman et al. 1983; Brimer et 

al. 1994; Chang et al. 1997; Chang et al. 1994a; Chang and Riviere 1991, 1993; Chang et al. 1994b; 

Denga et al. 1995; Eigenberg et al. 1983; Hurh et al. 2000a; Hurh et al. 2000b, c; Lessire et al. 1996; 

Oneto et al. 1995; Pena-Egido et al. 1988a; Pena-Egido et al. 1988b). 

Parathion at the dose of 3 mg/kg was intravenously administered to a rat. From the pharmacokinetic 

analysis, the terminal half-life and clearance of parathion were 3.4 hr and 93 ml/min/kg respectively 

(Eigenberg et al. 1983). Similar results were obtained from another study in rats, where the terminal 

half-life, AUC and clearance of parathion were 321 min., 52.5 μg-min/mL and 57.1 ml/min/kg 

respectively (Hurh et al. 2000a; Hurh et al. 2000b, c). In these studies, paraoxon levels were lower than 

their detection limits. 

Parathion pharmacokinetics in dogs were somewhat different from that in rats. After 30 mg intravenous 

dosing, plasma clearance and terminal half-life were 21 ml/min and 8.5-11.2 hr respectively. The plasma 

clearance in dogs appeared to be less than one-third of the plasma clearance for the rats. 

1.2.1 Absorption

In a pharmacokinetic study in dogs (Braeckman et al. 1983), parathion was administered at 5 mg/kg 

intravenously and 10 mg/kg orally to determine its absolute bioavialability (F). The fraction absorbed 

was high (57-98%). However, the bioavailability of parathion appeared to have a comparatively large 

variation because of its first pass metabolism and intersubject variation in parathion hepatic extraction 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

ratio (range = 63-97%) (Braeckman et al. 1983). Oral absorption of parathion was also studied in rats 

(Beubler et al. 1985).

Numerous dermal exposure studies have been performed (Antunes-Madeira and Madeira 1984; Bucks et 

al. 1990; Campbell et al. 2000; Carver and Riviere 1989; Carver et al. 1989; Fisher et al. 1985; GyrdHansen et al. 1993; Hawkins and Reifenrath 1986; Knaak et al. 1984; Murphy 1980; Qiao et al. 1996; 

Qiao et al. 1994; Reifenrath et al. 1984; Reifenrath et al. 1991; Riley and Kemppainen 1985; Shah and 

Guthrie 1983; Skinner and Kilgore 1982; Wester et al. 2000; Williams et al. 1990; Williams et al.

1996). After dermal application of 50 mg/kg parathion was performed along the midline of the entire 

back of pigs, the dermal bioavailability (F) was 0.0993. Tissue distribution of parathion in back skin, 

back fat, liver, kidney, muscle, adipose tissue was also determined. It appeared that 25.0-60.8% of the 

administered dose remained at the application site (Brimer et al. 1994). 

1.2.2 Distribution

Protein binding in dog serum and in human serum were 99% and 98% respectively (Braeckman et al.

1983). Tissue distributions were also reported in some species (Brimer et al. 1994). 

1.2.3 Metabolism 

Parathion is metabolized into paraoxon and 4-nitrophenol by desulfuration and dearylation (Fig. 1), 

respectively. 4-Nitrophenol formation is considered as the inactivation pathway, whereas paraoxon 

formation is considered as an activation pathway (Benke GM 1975; Bulusu and Chakravarty 1986, 

1988; Butler and Murray 1993, 1997; Chambers and Forsyth 1989; Chambers et al. 1994; Chaturvedi et 

al. 1991; Contreras et al. 1999; Halpert et al. 1980; Halpert and Neal 1981a, b; Hou et al. 1996; 

Kulkarni and Hodgson 1982; Kuo and Perera 2000; Lapadula et al. 1984; Levi and Hodgson 1985; 

Martinez-Zedillo et al. 1979; Monnet-Tschudi et al. 2000; Morgan et al. 1994; Murray and Butler 1994, 

1995; Mutch et al. 1999; Mutch et al. 2003; Nadin and Murray 1999; Pond et al. 1995; Pond et al. 1998; 

Purshottam and Srivastava 1987; Ramos and Sultatos 1998; Rowland et al. 1991; Soranno and Sultatos 

1992; Sultatos 1986; Sultatos et al. 1984; Sultatos and Gagliardi 1990; Sultatos and Minor 1986; 

Sultatos et al. 1985; Sultatos and Murphy 1983; Tang and Chambers 1999; Vargas Loza et al. 1997; 

Venera et al. 1978; Vitarius et al. 1995; Wallace and Dargan 1987; Watson et al. 1994; Zhang and 

Sultatos 1991; Zhu and Liu 1994). The primary metabolizing organ is the liver by the enzyme 

cytochrome P450 3A4. 

 

In mouse liver microsomes, the apparent Km’s for the formation of paraoxon and p-nitrophenol were 

29.6 and 26.5 μM respectively, and the apparent Vmaxs were 5.8 and 6.7 nmols/100 mg liver/min 

respectively (Sultatos 1986; Sultatos et al. 1984; Sultatos and Gagliardi 1990; Sultatos and Minor 1986; 

Sultatos et al. 1985; Sultatos and Murphy 1983). 

In rat liver microsomes, the kinetic curve for desulfuration of parathion is baphasic with apparent Km’s 

of 0.23 and 71.3 μM and Vmaxs of 3.62 and 4.56 nM/min/mg protein. For the dearylation reaction, 

parathion has an apparent Km and Vmax of 56 μM and 1.49 nM/min/mg protein, respectively (Ma and 

Chambers 1994, 1995). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

 S 

 C2H5 O 

 P – O NO2

 C2H5 O 

 

 Parathion 

 O 

Oxidation C2H5 O 

 P – O NO2

P450 C2H5 O 

Paraoxon 

 S O 

 C2H5 O C2H5 O 

 P – OH HO NO2 P – OH 

 C2H5 O C2H5 O 

 Phosphorothioic Acid Para-Nitrophenol Phosphoric Acid 

 Hydrolysis 

GS – C2H5

 O 

 C2H5 O 

 P – O NO2

 HO 

GSH 

 S 

 C2H5 O 

 P – O NO2

 HO 

GSH 

GS NO2

 GSH 

P450 

Figure 1. Metabolic pathway of parathion (adapted from Benke 1975). 

In a human microsomal study, parathion demonstrated biphasic behaviors in both individual microsomal 

and pooled samples. Its apparent Km1 and Km2 in individual microsomes were 0.30 and 165.5 μM and 

Vmax1 and Vmax2 were 290 and 821 pmol oxon/mg protein/min respectively. In pooled liver 

microsomes, parathion’s apparent Km1 and Km2 were 9.0 and 69.6 μM and Vmax1 and Vmax2 were 

106.6 and 2,478 pmol oxon/mg protein/min, respectively (Buratti et al. 2003; Ma and Chambers 1994, 

1995). Another study in human liver microsomes indicated that CYP3A4 is the major enzyme 

responsible for catalyzing parathion oxidation to paraoxon (Butler and Murray 1993, 1997). 

In addition to the liver, the brain is also capable of metabolizing parathion in various regions such as 

cortex, olfactory bulb/hypothalamus, striatum, cerebellum, midbrain, medulla and pons, and 

hippocampus. However, the total activity appears to be highest in the cortex (Soranno and Sultatos 

1992). 

1.2.4 Pharmacokinetic studies in special population

There have been a number of pharmacokinetic studies in specific populations (Benjaminov et al. 1992; 

Jaramillo and Reyes 1990). Neilsen et al. conducted a pharmacokinetic study in neonatal and young 

pigs. Intravenous parathion (0.5 mg/kg) was administered to newborn, 1 week and 8 weeks old piglets. 

The total body clearance was 7, 35 and 121 ml/min/kg, respectively. Tissue distribution in all groups 

was also presented. Interestingly, the newborn piglets seemed to retain parathion in significant amounts 

in many organs such as the liver, lung, brain, heart and muscle indicating that reduced total body 

clearance in the newborn markedly influenced tissue distribution (Nielsen et al. 1991). 

Pregnancy also affects parathion disposition and its toxicity. Concentrations of parathion were 

significantly higher in blood and brain of pregnant mice at most times after administration (5 mg of 

parathion/kg) of parathion when compared to non-pregnant mice (Weitman et al. 1983, 1986a, b). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

1.3 Pharmacokinetic interaction between parathion and other compounds 

Due to its metabolic pathway via CYP450, there are many possibilities that parathion pharmacokinetics 

may be affected by certain compounds particularly drugs, CYP450 inducers, inhibitors, other 

environmental pollutants and foods (Agyeman and Sultatos 1998; Carr et al. 2002; Chakravarty and 

Sreedhar 1982; Costa and Murphy 1984; Delaunois et al. 1999; Gelal et al. 2001; Graziano et al. 1985; 

Guilhermino et al. 1998a, b; Hurh et al. 2000a; Hurh et al. 2000b, c; Joshi and Thornburg 1986; Karanth

et al. 2001; Miranda et al. 1998; Mourelle et al. 1986; Murphy 1980; O'Shaughnessy and Sultatos 1995; 

Purshottam and Kaveeshwar 1982; Purshottam and Srivastava 1984; Ramos and Sultatos 1998; 

Sawahata and Neal 1982; Siller et al. 1997; Wester et al. 2000). 

Cimetidine, a non-specific inhibitor of CYP450, was capable of antagonizing methyl parathion toxicity 

but failed to decrease parathion-induced toxicity in mice and rats (Weitman et al. 1983). Rats pretreated 

with dexamethasone, a specific inducer of CYP3A23, showed faster clearance of parathion than control 

rats (Hurh et al. 2000a). 

1.4 PBPK modeling of parathion and Monte Carlo simulation

A PBPK model of parathion was developed by Gearhart et al. In brief, the model describes the 

metabolism of parathion to paraoxon by the liver, the inhibition of acetylcholinesterase, 

butyrylcholinesterase, and carboxylesterase by paraoxon in the brain, liver, kidneys, rapidly perfused 

tissues and the arterial and venous blood (Gearhart 1994). Physiologic parameters are available in the 

literature. (Jepson et al. 1994; Kousba and Sultatos 2002). Due to the existence of paraoxonase 

polymorphisms (Costa et al. 2003; Diepgen and Geldmacher-von Mallinckrodt 1986; Eaton 2000; 

Furlong et al. 2000; Haber et al. 2002; Laplaud et al. 1998; Lee et al. 2003; Shih et al. 1998), a 

subsequent study was conducted by using this existing PBPK model with Monte Carlo simulation to 

elaborate the effect of polymorphic paraoxonase (PON1) on its toxicity (Gentry et al. 2002). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

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Methyl Parathion 

2.0 Introduction

Methyl parathion (O,O-dimethyl O-4-nitrophenylphosphorothioate) is a highly toxic organophosphorus 

insecticide approved for specific agricultural crops. Its use is restricted by appropriately trained certified 

pesticide applicators (Garcia et al. 2003). However, it has been used illegally indoors in certain areas of 

the Southern and Midwestern parts of the United States due to its effectiveness and low cost (ATSDR 

2001; Rubin et al. 2002), leading to an increased health risk in non-workers, children and pregnant 

women. 

2.1 Toxic effects

Neurotoxicity is the major toxic effect of methyl parathion (MP) or its metabolite, methyl paraoxon, in 

various species caused by inhibition of acetylcholinesterase (AChE) enzymes, resulting in acetylcholine 

accumulation at postsynaptic receptors and overstimulation of cholinergic systems (Chambers and Carr 

1993; Gupta et al. 2000; Hahn et al. 1991; Ma et al. 2003). The median lethal dose (LD50) of MP in 

mice applied orally and dermally was 14.5 and 1200 mg/kg body weight, respectively, while the dermal 

median effective dose (ED50) that caused 50% reduction in AChE was 550 mg/kg at 24 hours after 

dosing (Haley et al. 1975; Skinner and Kilgore 1982). The developing animals are more sensitive to 

acute toxicity of MP than adults, indicating the age-related differences in sensitivity to MP exposure 

(Liu et al. 1999; Pope and Chakraborti 1992; Pope et al. 1991). 

In humans, manifestations of exposure to MP such as shortness of breath, nose bleeding, vomiting, 

diarrhea, abdominal cramps, headache, eye pain, blurred vision, sweating, confusion, muscle 

contraction, contact burns and erythema multiforme eruption (following dermal exposure) were 

reported. The severe neurotoxic effects include loss of coordination, slurred speech, fatigue and death 

caused by respiratory or cardiac arrest (Azaroff and Neas 1999; Fisher 1986; Karki et al. 2001; Rehner

et al. 2000). Cranial nerve palsies and intermediate syndromes have also been reported in certain 

patients (Karki et al. 2001; Narendra et al. 1989). 

Other effects reported include genotoxic and mutagenic effects (Bartoli et al. 1991; Breau et al. 1985; 

Chen et al. 1981; de Cassia Stocco et al. 1982; Degraeve and Moutschen 1984; Dolara et al. 1993; 

Griffin and Hill 1978; Grover and Malhi 1985; Lodovici et al. 1994; Lodovici et al. 1997; Mathew et al.

1990; Mathew et al. 1992; Nehez et al. 1994; Rashid and Mumma 1984; Rupa et al. 1990; Rupa et al.

1991; Singh et al. 1984; Tripathy et al. 1987; Undeger et al. 2000; Velazquez et al. 1990; 

Vijayaraghavan and Nagarajan 1994; Wiaderkiewicz et al. 1986), effects on calmodulin (Pala et al.

1991), effects on liver and muscle enzymes (Della Morte et al. 1994; Gupta et al. 1994; Jabbar et al.

1990), hematoxicity (Parent-Massin and Thouvenot 1993), immunotoxic effects (Crittenden et al. 1998; 

Institoris et al. 1995; Institoris et al. 1992; Lee et al. 1979; Sunil Kumar et al. 1993; Undeger et al.

2000), hormonal effects (Asmathbanu and Kaliwal 1997; Fatranska et al. 1978; Lukaszewicz-Hussain et 

al. 1985; Sortur and Kaliwal 1999), reproductive and developmental effects (Basha and Nayeemunnisa 

1993; Desi et al. 1998; Dhondup and Kaliwal 1997; Garcia et al. 2003; Gupta et al. 1985; Gupta et al.

1984; Kumar and Desiraju 1992; Mahaboob Basha et al. 2001; Mahaboob Basha and Nayeemunnisa 

1993; Nagymajtenyi et al. 1995; Nayeemunnisa and Begum 1992; Sortur and Kaliwal 1999),

embryotoxicity (Tanimura et al. 1967; Uzokwu 1974), cardiac toxicity (Howard and Pope 2002), and 

behavioral effects (George et al. 1992; Liu et al. 1994; Schulz et al. 1990; Zhu et al. 2001). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

2.2 Pharmacokinetics 

2.2.1 Absorption 

Because of its lipid solubility, MP can be absorbed through skin; therefore, the most likely route of 

human exposure is dermal, particularly from agricultural field reentry (Abu-Qare et al. 2000). Oral 

exposure can also occur via contaminated food or water consumption and suicidal attempt (Garcia et al.

2003) while exposure to MP via inhalation during spraying is questionable (Kummer and van Sittert 

1986). 

Oral absorption 

MP is well and rapidly absorbed through the gastrointestinal tract following oral gavage in mice 

(Hollingworth 1967), rats (Garcia-Repetto et al. 1997; Kramer and Ho 2002; Kramer et al. 2002; 

Miyamoto 1963), guinea pigs (Miyamoto 1963), dogs (Braeckman et al. 1983), and humans (Morgan et 

al. 1977). However, the oral bioavailability is very low (5-20%), which can be explained by a significant 

hepatic first-pass effect. The oral absorption rate constant after 1.5-2.5 mg/kg administration of MP in 

rats was 1.2 h-1 (Kramer and Ho 2002; Kramer et al. 2002). 

Dermal absorption 

An in vitro model using human skin in a static diffusion cell system demonstrated that 5.2% of the 

applied dose of MP from a commercial formulation was present after 24 h (Sartorelli et al. 1997). In 

adult female rats and pregnant rats, 20-50% of administered dose was absorbed following a single 

dermal dose of 10-50 mg/kg MP with the absorption rate constant of 0.41 h-1 (Abu-Qare et al. 2000; 

Kramer and Ho 2002; Kramer et al. 2002). 

2.2.2 Distribution

Following oral and dermal administration, MP is extensively bound to plasma protein and rapidly 

distributed to tissues including placenta and fetus. Then it is slowly redistributed to the central 

compartment (Abu-Qare et al. 2000; Garcia-Repetto et al. 1997). The highest level of MP is found in 

adipose tissue. Distribution coefficients of adipose tissue, liver and brain in rats and mice have been 

published (Garcia-Repetto et al. 1995; Sultatos et al. 1990). The terminal half-life varies from 7.2 h to 

15 days, depending on species and gender (Abu-Qare et al. 2000; Braeckman et al. 1980; GarciaRepetto et al. 1997; Kramer and Ho 2002; Kramer et al. 2002). The volume of distribution is relatively 

high (9.6 l/kg in dogs and 10.1 l/kg in female rats) (Braeckman et al. 1980; Kramer and Ho 2002). 

2.2.3 Metabolism

MP is metabolized by hepatic and extrahepatic phase I and phase II enzymes (Figure 2.1) (Abu-Qare et 

al. 2000; Garcia et al. 2003). Phase I metabolism include dearylation of MP, leading to the formation of 

p-nitrophenol and dimethyl thiophosphoric acid, which promotes detoxification. On the other hand, 

desulfuration by cytochrome P450 can activate MP to methyl paraoxon, the neurotoxic metabolite 

(Yamamoto et al. 1983; Zhang and Sultatos 1991). This oxidation process is the major metabolic 

pathway of MP in the liver (Anderson et al. 1992; Sultatos 1987). Methyl paraoxon is also formed in 

the brain. CYP2B has been demonstrated to be responsible for MP activation in rat brain extracts 

(Albores et al. 2001). The activities of dearylation and desulfuration of MP were reduced when a low 

dose of MP was given repeatedly in rats (Yamamoto et al. 1982). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

 S 

 CH3 O 

 P – O NO2

 CH3 O 

Methyl Parathion 

 O 

Oxidation CH3 O 

 P – O NO2

1. CH3 O 

Methyl Paraoxon 

 S O 

 CH3 O CH3 O 

 P – OH HO NO2 P – OH 

 CH3 O CH3 O 

Dimethyl Phosphorothioic Acid Para-Nitrophenol Dimethyl Phosphoric Acid 

 2. Hydrolysis 3. 

GS - CH3

 O 

 CH3 O 

 P – O NO2

 HO 

GSH 6. 

 S 

 CH3 O 

 P – O NO2

 HO 

GSH 4. 

P450 

GS NO2

5. 

 GSH 

P450 

Figure 2.1. Metabolic pathways of MP (Benke and Murphy 1975; Garcia et al. 2003). 

Methyl paraoxon is then hydrolyzed by liver and plasma paraoxonase to form p-nitrophenol and 

dimethyl phosphoric acid (Garcia et al. 2003). The correlation between LD50 of MP in rats of several 

ages and reaction rates of metabolism of methyl paraoxon, both hydrolysis and GSH-dependent 

pathways, was reported, indicating that these pathways contributed to age-related differences in MP 

toxicity (Benke and Murphy 1975). p-Nitrophenol further undergoes glucuronidation and sulfuric 

conjugation. 

MP is also conjugated by glutathione S-aryl transferase to form p-nitrophenyl mercapturic acid (Di Ilio

et al. 1995; Huang and Sultatos 1993; Sultatos and Woods 1988) and by glutathione S-alkyl transferase 

to yield S-methyl glutathione (Radulovic et al. 1987; Radulovic et al. 1986). Furthermore, a study has 

reported another non mixed-function oxidative pathway of MP in brain tissue subfractions that 

transformed MP to its isomer (de Lima et al. 1996). 

 

2.2.4 Excretion

MP is rapidly eliminated after oral and dermal administration. Renal excretion is the major route of MP 

elimination. In rats, 75-90% of administered dose was recovered in urine and less than 10% was found 

in feces (Abu-Qare et al. 2000; Abu-Qare and Abou-Donia 2000; Hollingworth 1967; Miyamoto 1963). 

In humans, the urinary metabolites of MP are p-nitrophenol, dimethylphosphate, and unidentified 

metabolites (Morgan et al. 1977). Therefore, p-nitrophenol has been used as a biomarker of MP 

exposure in humans (Barr et al. 2002; Chang et al. 1997; Esteban et al. 1996; Hryhorczuk et al. 2002; 

Rubin et al. 2002). 

2.3 Interactions

MP has been reported to produce behavioral alterations when given in combination with endosulfan 

(Castillo et al. 2002) or toxaphene (Crowder et al. 1980) and more likely to induce intermediate 

syndrome when combined with parathion (De Bleecker et al. 1992). Conversely, the inhibition of 

cholinesterase enzyme activity was significantly lowered when MP was administered with either 

chlorpyrifos or diazinon, which could be due to competition for cytochrome P-450 enzymes, resulting in 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

inhibition of oxon formation (Abu-Qare et al. 2001; Abu-Qare and Abou-Donia 2001). Moreover, 

cimetidine, chlordecone, mirex and linuron, gentamicin and rifamycin, polychlorinated biphenyls 

(PCBs), and permethrin have also been demonstrated to change the toxicity of MP (Carr et al. 2002; 

Joshi and Thornburg 1986; Ortiz et al. 1995; Tvede et al. 1989; Youssef et al. 1987). No interactions 

with acetaminophen and hexachlorocyclohexane (HCH) have been reported (Costa and Murphy 1984; 

Dikshith et al. 1991). 

2.4 PBPK models

One- to three-compartment classical models has been used to fit the blood concentration data following 

intravenous, oral, and dermal administration (Abu-Qare et al. 2000; Braeckman et al. 1983; Braeckman

et al. 1980; Kramer and Ho 2002; Kramer et al. 2002). No PBPK models for MP have been published. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

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Rashid, K. A., and Mumma, R. O. (1984). Genotoxicity of methyl parathion in short-term bacterial test 

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Rubin, C., Esteban, E., Hill, R. H., Jr., and Pearce, K. (2002). Introduction--the methyl parathion story: a 

chronicle of misuse and preventable human exposure. Environ Health Perspect 110 Suppl 6, 1037-

40. 

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Schulz, H., Desi, I., and Nagymajtenyi, L. (1990). Behavioral effects of subchronic intoxication with 

parathion-methyl in male Wistar rats. Neurotoxicol Teratol 12, 125-7. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Singh, S., Lehmann-Grube, B., and Goedde, H. W. (1984). Cytogenetic effects of paraoxon and methylparathion on cultured human lymphocytes: SCE, clastogenic activity and cell cycle delay. Int Arch 

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Toxicol Sci 7, 35-41. 

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fenitrothion and methylparathion in male rats. Acta Pharmacol Toxicol (Copenh) 53, 96-102. 

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parathion and methyl parathion in male and female rat livers perfused in situ. Drug Metab Dispos

19, 473-7. 

Zhu, H., Rockhold, R. W., Baker, R. C., Kramer, R. E., and Ho, I. K. (2001). Effects of single or 

repeated dermal exposure to methyl parathion on behavior and blood cholinesterase activity in rats. J 

Biomed Sci 8, 467-74. 

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Chlorpyrifos 

3.0 Introduction

Chlorpyrifos (O,O-diethyl O-(3,5,6-trichloro-2-pyridyl) phosphorothiolate) is an OP pesticide with 

restricted uses in the U.S. It is produced in the U.S. and marketed under various trade names, including 

Dursban®, Lorsban®, and other names (ATSDR 1997). Chlorpyrifos (CP) has been used against pests 

in turfgrass, commercial agriculture, and in residential settings, although indoor uses have been 

restricted. CP has been found in drinking water supplies, although often at low levels (ATSDR 1997). 

3.1 Toxic Effects

Numerous systemic effects of CP have been reported, but the critical effect for risk assessment is 

inhibition of acetylcholinesterase by CP or CP-oxon (ATSDR 1997). This can result in headache, 

diaphoresis, nausea, vomiting, diarrhea, epigastric cramping, bradycardia, blurred vision, miosis, 

bronchoconstriction and excess mucous secretions, pulmonary edema, dyspnea, muscle fasciculations, 

salivation, lacrimation, urination, tremors, anxiety, drowsiness, confusion, ataxia, abnormal gait, 

hypotension, and memory impairment (Ballantyne and Marrs 1992). 

Neurodevelopmental effects have also been described (Auman et al. 2000; Campbell et al. 1997; Carr et 

al. 2001; Chakraborti et al. 1993; Chanda and Pope 1996; Crumpton et al. 2000; Dam et al. 1999; Dam

et al. 1998, 2000, 2003; Das and Barone 1999; Garcia et al. 2002; Garcia et al. 2003; Gore 2001, 2002; 

Howard and Pope 2002; Jett et al. 2001; Lassiter et al. 1998; Levin et al. 2002; Levin et al. 2001; 

Olivier et al. 2001; Qiao et al. 2002; Qiao et al. 2001; Qiao et al. 2003; Raines et al. 2001; Richardson 

and Chambers 2003; Roy et al. 1998; Sachana et al. 2001; Slotkin et al. 2001; Slotkin et al. 2002; Song

et al. 1998; Tang et al. 1999; Whitney et al. 1995; Won et al. 2001). Some of these studies suggested 

that developmental neurotoxicity occurred at lower exposure levels than did acetylcholinesterase 

inhibition in adult animals, and developmental neurotoxicity may be worthy of consideration in future 

risk assessments (Abdel-Rahman et al. 2002). 

3.2 Pharmacokinetics

3.2.1 Absorption

CP is well absorbed through the gut after oral exposure to CP in drinking water or food. In humans or 

rodents, 70-90% of the oral dose was absorbed (Ahdaya et al. 1981; Bakke et al. 1976; Nolan et al.

1984; Smith et al. 1967). Only 3% of the dermal dose was absorbed; (Nolan et al. 1984) however, this 

is dependant on the vehicle. When acetone was used as vehicle in another study, 46-99% of the dose 

was absorbed (Shah et al. 1987). Absorption rates through the gut (Cook and Shenoy 2003) and skin 

(Griffin et al. 2000; Sartorelli et al. 1998) have been quantified. 

3.2.2 Distribution

CP rapidly distributes to tissues after absorption (Shah et al. 1987; Shah et al. 1981; Smith et al. 1967). 

3.2.3 Metabolism

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

CP is primarily metabolized in the liver due to the high concentration of cytochrome (CYP) P450s in 

that organ (Ma and Chambers 1994; Sultatos and Murphy 1983b). CP is rapidly bioactivated to CPoxon by multiple isoforms of CYP P450. Both CP and CP-oxon are hydrolyzed by acetylcholinesterase 

to 3,5,6-trichloro-2-pyridinol (TCP), diethyl thiophosphate, and diethyl phosphate (Bakke et al. 1976; 

Nolan et al. 1984; Smith et al. 1967; Sultatos et al. 1985; Sultatos and Murphy 1983a, b). The principal 

metabolic pathways are shown in Figure 3.1 below. 

Figure 3.1 

Principal Metabolic Pathways for Chlorpyrifos 

O N

Cl

Cl

P

O

O

S

Cl

O N

Cl

Cl

P

O

O

O

Cl

P

O

S

P

O

O

(CH3CH2O)2

Chlorpyrifos Chlorpyrifos oxon

CYP450

TCP TCP

(CH3CH2O)2

Diethylthiophosphate (DETP) + Diethylphosphate (DEP) +

CYP450 CYP450

Modified from ATSDR (1997) 

For risk assessment or PBPK model development, studies of metabolism are often used, but further 

incorporation of pharmacodynamic responses such as acetylcholinesterase inhibition would improve the 

risk assessment or PBPK model development. Therefore, either type of study is included in this review. 

Several human pharmacokinetic (PK) or biomonitoring studies have been conducted (Cocker et al.

2002; Drevenkar et al. 1993; Fenske et al. 2002; Sams and Mason 1999). 

A number of animal studies have been conducted addressing PKs, binding, cholinestase inhibition, or 

other endpoints with CP (Abdel-Rahman et al. 2002; Atterberry et al. 1997; Bushnell et al. 1994; 

Bushnell et al. 1993; Carr and Chambers 1996; Carr et al. 2002; Carr et al. 1995; Chanda et al. 1997; 

Chiappa et al. 1995; Cowan et al. 2001; Hunter et al. 1999; Karanth and Pope 2000; Lassiter et al. 1999; 

Li et al. 2000; Li et al. 1995; Liu et al. 1999; Mattsson et al. 2000; Mortensen et al. 1996; Mortensen et 

al. 1998; Moser et al. 1998; Moser and Padilla 1998; Padilla et al. 2000; Padilla et al. 1994; Pond et al.

1995; Pond et al. 1998; Stanton et al. 1994). 

Fetal transfer of CP or metabolites was assessed in several studies, including two recent ones (AbdelRahman et al. 2002; Ashry et al. 2002). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

In vitro PK studies with CP have been conducted. These include studies using cell lines (Barber and 

Ehrich 2001; Ehrich et al. 1997; Monnet-Tschudi et al. 2000), microsomes (Buratti et al. 2003; Katz et 

al. 1997; Ma and Chambers 1994; Poet et al. 2003; Sams et al. 2000; Tang et al. 2001; Usmani et al.

2003), antibodies (Buratti et al. 2003), tissue slices (Liu et al. 2002), or other systems (Amitai et al.

1998). 

Genetic polymorphisms relevant to CP metabolism have been described (Brophy et al. 2000; Costa et al.

1999; Costa et al. 2003; Dai et al. 2001; Furlong et al. 2000a; Furlong et al. 1998; Furlong et al. 2000b). 

3.2.4 Excretion

Most metabolites are found as conjugated metabolites of TCP in the urine. The half-life of elimination 

in rats was 10-16 hours in most tissues and 62 hours for fat (Smith et al. 1967). The half life for 

elimination in humans was estimated at 27 hours (Nolan et al. 1984). 

3.3 Interactions with other OP pesticides

Interactions of CP with other OP pesticides have been studied (Axelrad et al. 2002; Karanth et al. 2001; 

Richardson et al. 2001; Tang et al. 2002; Usmani et al. 2002). 

3.4 PBPK models for chlorpyrifos

One group has developed PBPK models for CP (Timchalk et al. 2002a; Timchalk et al. 2002b). These 

models were based on adult rat and human exposures in gavage, dietary, or dermal studies in a sevencompartment model structure. They included saturable metabolism of CP by CYP and “a-esterases,” 

and binding of CP-oxon to “b-esterases” as a second order process. Regeneration of b-esterases was 

also included. 

A classical PK model for CP was also described (Rigas et al. 2001). 

The Timchalk et al. CP PBPK model could be adapted for use in a PBPK-model based risk assessment. 

Adaptation should include alteration for drinking water scenarios. Also, during the process of 

determining the common mechanism of toxicity for a mixture of OP pesticides, if developmental 

neurotoxicity is an important part of the risk assessment, PBPK models for relevant endpoints in the 

fetus or neonates should be considered. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

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Dam, K., Garcia, S. J., Seidler, F. J., and Slotkin, T. A. (1999). Neonatal chlorpyrifos exposure alters 

synaptic development and neuronal activity in cholinergic and catecholaminergic pathways. Brain 

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Dam, K., Seidler, F. J., and Slotkin, T. A. (1998). Developmental neurotoxicity of chlorpyrifos: delayed 

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Costa, L. G. (2000a). The PON1 gene and detoxication. Neurotoxicology 21, 581-7. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Furlong, C. E., Li, W. F., Costa, L. G., Richter, R. J., Shih, D. M., and Lusis, A. J. (1998). Genetically 

determined susceptibility to organophosphorus insecticides and nerve agents: developing a mouse 

model for the human PON1 polymorphism. Neurotoxicology 19, 645-50. 

Furlong, C. E., Li, W. F., Richter, R. J., Shih, D. M., Lusis, A. J., Alleva, E., and Costa, L. G. (2000b). 

Genetic and temporal determinants of pesticide sensitivity: role of paraoxonase (PON1). 

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effects on glial fibrillary acidic protein. Brain Res Dev Brain Res 133, 151-61. 

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or postnatal chlorpyrifos exposure: effects on neurospecific proteins indicate changing 

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Howard, M. D., and Pope, C. N. (2002). In vitro effects of chlorpyrifos, parathion, methyl parathion and 

their oxons on cardiac muscarinic receptor binding in neonatal and adult rats. Toxicology 170, 1-10. 

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Levin, E. D., Addy, N., Nakajima, A., Christopher, N. C., Seidler, F. J., and Slotkin, T. A. (2001). 

Persistent behavioral consequences of neonatal chlorpyrifos exposure in rats. Brain Res Dev Brain 

Res 130, 83-9. 

Li, W. F., Costa, L. G., Richter, R. J., Hagen, T., Shih, D. M., Tward, A., Lusis, A. J., and Furlong, C. E. 

(2000). Catalytic efficiency determines the in-vivo efficacy of PON1 for detoxifying 

organophosphorus compounds. Pharmacogenetics 10, 767-79. 

Li, W. F., Furlong, C. E., and Costa, L. G. (1995). Paraoxonase protects against chlorpyrifos toxicity in 

mice. Toxicol Lett 76, 219-26. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Liu, J., Chakraborti, T., and Pope, C. (2002). In vitro effects of organophosphorus anticholinesterases on 

muscarinic receptor-mediated inhibition of acetylcholine release in rat striatum. Toxicol Appl 

Pharmacol 178, 102-8. 

Liu, J., Olivier, K., and Pope, C. N. (1999). Comparative neurochemical effects of repeated methyl 

parathion or chlorpyrifos exposures in neonatal and adult rats. Toxicol Appl Pharmacol 158, 186-96. 

Ma, T., and Chambers, J. E. (1994). Kinetic parameters of desulfuration and dearylation of parathion 

and chlorpyrifos by rat liver microsomes. Food Chem Toxicol 32, 763-7. 

Mattsson, J. L., Maurissen, J. P., Nolan, R. J., and Brzak, K. A. (2000). Lack of differential sensitivity to 

cholinesterase inhibition in fetuses and neonates compared to dams treated perinatally with 

chlorpyrifos. Toxicol Sci 53, 438-46. 

Monnet-Tschudi, F., Zurich, M. G., Schilter, B., Costa, L. G., and Honegger, P. (2000). Maturationdependent effects of chlorpyrifos and parathion and their oxygen analogs on acetylcholinesterase and 

neuronal and glial markers in aggregating brain cell cultures. Toxicol Appl Pharmacol 165, 175-83. 

Mortensen, S. R., Chanda, S. M., Hooper, M. J., and Padilla, S. (1996). Maturational differences in 

chlorpyrifos-oxonase activity may contribute to age-related sensitivity to chlorpyrifos. J Biochem 

Toxicol 11, 279-87. 

Mortensen, S. R., Hooper, M. J., and Padilla, S. (1998). Rat brain acetylcholinesterase activity: 

developmental profile and maturational sensitivity to carbamate and organophosphorus inhibitors. 

Toxicology 125, 13-9. 

Moser, V. C., Chanda, S. M., Mortensen, S. R., and Padilla, S. (1998). Age- and gender-related 

differences in sensitivity to chlorpyrifos in the rat reflect developmental profiles of esterase 

activities. Toxicol Sci 46, 211-22. 

Moser, V. C., and Padilla, S. (1998). Age- and gender-related differences in the time course of 

behavioral and biochemical effects produced by oral chlorpyrifos in rats. Toxicol Appl Pharmacol

149, 107-19. 

Nolan, R. J., Rick, D. L., Freshour, N. L., and Saunders, J. H. (1984). Chlorpyrifos: pharmacokinetics in 

human volunteers. Toxicol Appl Pharmacol 73, 8-15. 

Olivier, K., Jr., Liu, J., and Pope, C. (2001). Inhibition of forskolin-stimulated cAMP formation in vitro 

by paraoxon and chlorpyrifos oxon in cortical slices from neonatal, juvenile, and adult rats. J 

Biochem Mol Toxicol 15, 263-9. 

Padilla, S., Buzzard, J., and Moser, V. C. (2000). Comparison of the role of esterases in the differential 

age-related sensitivity to chlorpyrifos and methamidophos. Neurotoxicology 21, 49-56. 

Padilla, S., Wilson, V. Z., and Bushnell, P. J. (1994). Studies on the correlation between blood 

cholinesterase inhibition and 'target tissue' inhibition in pesticide-treated rats. Toxicology 92, 11-25. 

Poet, T. S., Wu, H., Kousba, A. A., and Timchalk, C. (2003). In vitro rat hepatic and intestinal 

metabolism of the organophosphate pesticides chlorpyrifos and diazinon. Toxicol Sci 72, 193-200. 

Pond, A. L., Chambers, H. W., and Chambers, J. E. (1995). Organophosphate detoxication potential of 

various rat tissues via A- esterase and aliesterase activities. Toxicol Lett 78, 245-52. 

Pond, A. L., Chambers, H. W., Coyne, C. P., and Chambers, J. E. (1998). Purification of two rat hepatic 

proteins with A-esterase activity toward chlorpyrifos-oxon and paraoxon. J Pharmacol Exp Ther

286, 1404-11. 

Qiao, D., Seidler, F. J., Padilla, S., and Slotkin, T. A. (2002). Developmental neurotoxicity of 

chlorpyrifos: what is the vulnerable period? Environ Health Perspect 110, 1097-103. 

Qiao, D., Seidler, F. J., and Slotkin, T. A. (2001). Developmental neurotoxicity of chlorpyrifos modeled 

in vitro: comparative effects of metabolites and other cholinesterase inhibitors on DNA synthesis in 

PC12 and C6 cells. Environ Health Perspect 109, 909-13. 

Qiao, D., Seidler, F. J., Tate, C. A., Cousins, M. M., and Slotkin, T. A. (2003). Fetal chlorpyrifos 

exposure: adverse effects on brain cell development and cholinergic biomarkers emerge postnatally 

and continue into adolescence and adulthood. Environ Health Perspect 111, 536-44. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Raines, K. W., Seidler, F. J., and Slotkin, T. A. (2001). Alterations in serotonin transporter expression in 

brain regions of rats exposed neonatally to chlorpyrifos. Brain Res Dev Brain Res 130, 65-72. 

Richardson, J., and Chambers, J. (2003). Effects of gestational exposure to chlorpyrifos on postnatal 

central and peripheral cholinergic neurochemistry. J Toxicol Environ Health A 66, 275-89. 

Richardson, J. R., Chambers, H. W., and Chambers, J. E. (2001). Analysis of the additivity of in vitro 

inhibition of cholinesterase by mixtures of chlorpyrifos-oxon and azinphos-methyl-oxon. Toxicol 

Appl Pharmacol 172, 128-39. 

Rigas, M. L., Okino, M. S., and Quackenboss, J. J. (2001). Use of a pharmacokinetic model to assess 

chlorpyrifos exposure and dose in children, based on urinary biomarker measurements. Toxicol Sci

61, 374-81. 

Roy, T. S., Andrews, J. E., Seidler, F. J., and Slotkin, T. A. (1998). Chlorpyrifos elicits mitotic 

abnormalities and apoptosis in neuroepithelium of cultured rat embryos. Teratology 58, 62-8. 

Sachana, M., Flaskos, J., Alexaki, E., Glynn, P., and Hargreaves, A. J. (2001). The toxicity of 

chlorpyrifos towards differentiating mouse N2a neuroblastoma cells. Toxicol In Vitro 15, 369-72. 

Sams, C., and Mason, H. J. (1999). Detoxification of organophosphates by A-esterases in human serum. 

Hum Exp Toxicol 18, 653-8. 

Sams, C., Mason, H. J., and Rawbone, R. (2000). Evidence for the activation of organophosphate 

pesticides by cytochromes P450 3A4 and 2D6 in human liver microsomes. Toxicol Lett 116, 217-21. 

Sartorelli, P., Aprea, C., Cenni, A., Novelli, M. T., Orsi, D., Palmi, S., and Matteucci, G. (1998). 

Prediction of percutaneous absorption from physicochemical data: a model based on data of in vitro 

experiments. Ann Occup Hyg 42, 267-76. 

Shah, P. V., Fisher, H. L., Sumler, M. R., Monroe, R. J., Chernoff, N., and Hall, L. L. (1987). 

Comparison of the penetration of 14 pesticides through the skin of young and adult rats. J Toxicol 

Environ Health 21, 353-66. 

Shah, P. V., Monroe, R. J., and Guthrie, F. E. (1981). Comparative rates of dermal penetration of 

insecticides in mice. Toxico1 Appl Pharmacol 59, 414-423. 

Slotkin, T. A., Cousins, M. M., Tate, C. A., and Seidler, F. J. (2001). Persistent cholinergic presynaptic 

deficits after neonatal chlorpyrifos exposure. Brain Res 902, 229-43. 

Slotkin, T. A., Tate, C. A., Cousins, M. M., and Seidler, F. J. (2002). Functional alterations in CNS 

catecholamine systems in adolescence and adulthood after neonatal chlorpyrifos exposure. Brain Res 

Dev Brain Res 133, 163-73. 

Smith, G. N., Watson, B. S., and Fischer, F. S. (1967). Investigations on dursban insecticide: 

Metabolism of [36Cl] O,O-diethyl-O- 3,5,6-trichloro-2-pyridyl phosphorothioate in rats. J Agric 

Food Chem 15, 132-8. 

Song, X., Violin, J. D., Seidler, F. J., and Slotkin, T. A. (1998). Modeling the developmental 

neurotoxicity of chlorpyrifos in vitro: macromolecule synthesis in PC12 cells. Toxicol Appl 

Pharmacol 151, 182-91. 

Stanton, M. E., Mundy, W. R., Ward, T., Dulchinos, V., and Barry, C. C. (1994). Time-dependent 

effects of acute chlorpyrifos administration on spatial delayed alternation and cholinergic 

neurochemistry in weanling rats. Neurotoxicology 15, 201-8. 

Sultatos, L. G., Minor, L. D., and Murphy, S. D. (1985). Metabolic activation of phosphorothioate 

pesticides: role of the liver. J Pharmacol Exp Ther 232, 624-8. 

Sultatos, L. G., and Murphy, S. D. (1983a). Hepatic microsomal detoxification of the organophosphates 

paraoxon and chlorpyrifos oxon in the mouse. Drug Metab Dispos 11, 232-8. 

Sultatos, L. G., and Murphy, S. D. (1983b). Kinetic analyses of the microsomal biotransformation of the 

phosphorothioate insecticides chlorpyrifos and parathion. Fundam Appl Toxicol 3, 16-21. 

Tang, J., Cao, Y., Rose, R. L., Brimfield, A. A., Dai, D., Goldstein, J. A., and Hodgson, E. (2001). 

Metabolism of chlorpyrifos by human cytochrome P450 isoforms and human, mouse, and rat liver 

microsomes. Drug Metab Dispos 29, 1201-4. 

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Tang, J., Cao, Y., Rose, R. L., and Hodgson, E. (2002). In vitro metabolism of carbaryl by human 

cytochrome P450 and its inhibition by chlorpyrifos. Chem Biol Interact 141, 229-41. 

Tang, J., Carr, R. L., and Chambers, J. E. (1999). Changes in rat brain cholinesterase activity and 

muscarinic receptor density during and after repeated oral exposure to chlorpyrifos in early postnatal 

development. Toxicol Sci 51, 265-72. 

Timchalk, C., Kousba, A., and Poet, T. S. (2002a). Monte Carlo analysis of the human chlorpyrifosoxonase (PON1) polymorphism using a physiologically based pharmacokinetic and 

pharmacodynamic (PBPK/PD) model. Toxicol Lett 135, 51-9. 

Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002b). 

A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the 

organophosphate insecticide chlorpyrifos in rats and humans. Toxicol Sci 66, 34-53. 

Usmani, K. A., Rose, R. L., Goldstein, J. A., Taylor, W. G., Brimfield, A. A., and Hodgson, E. (2002). 

In vitro human metabolism and interactions of repellent N,N-diethyl-m- toluamide. Drug Metab 

Dispos 30, 289-94. 

Usmani, K. A., Rose, R. L., and Hodgson, E. (2003). Inhibition and activation of the human liver 

microsomal and human cytochrome P450 3A4 metabolism of testosterone by deployment-related 

chemicals. Drug Metab Dispos 31, 384-91. 

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chlorpyrifos on acetylcholine release in rat brain. Neurotoxicology 22, 39-48. 

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Diazinon 

4.0 Introduction

Diazinon (O,O-diethyl O-2 isopropyl-6-methylpyrimidinyl phosphothiolate) is an organophosphate 

insecticide that is used in agriculture and as a topically applied pesticide in animal use (ATSDR 1996; 

Garfitt et al. 2002; USEPA 2003). It is a colorless liquid and was available as granules, emulsifiable 

concentrate, dust, and wettable powder. It is soluble in most organic solvents and is stable in neutral 

media but is slowly hydrolyzed in alkaline media and more rapidly in acidic media (HSDB 2003). 

4.1 Toxic effects

Many effects have been determined to occur in chronic bioassays in laboratory animals. Some of the 

target organs affected include the respiratory, kidney, cardiovascular, gastrointestinal, hematological, 

hepatic, endocrine, lymphatic, reproductive (ATSDR 1996), immune (Galloway and Handy 2003), and 

nervous systems (Gordon and Mack 2003). 

Acute oral LD50s in rats between 76 and 408 mg/kg have been reported (ATSDR 1996). Dermal 

LD50’s in rats ranged between 455 and 1100 mg/kg (ATSDR 1996). 

Diazinon has been reported to cause genotoxicity in a number of assays, including in the S.

typhimurium, mouse lymphoma cell forward mutation assay, and Chinese hamster cell chromosomal 

aberration assay, but was negative in several other assays (ATSDR 1996; Hatjian et al. 2000). 

Risk assessment for exposure to diazinon, however, has generally been based on inhibition of brain 

acetylcholinesterase (AChE) as the critical endpoint of toxicity (ATSDR 1996). As with other OP 

pesticides, the mode of action of diazinon is inhibition of AChE in the central and peripheral nervous 

system. Diazinon is a weak inhibitor of AChE while the oxon analog is much more potent. Therefore, 

activation by mixed function oxygenases, primarily in the liver, is an important bioactivating step. 

Other metabolic pathways (see Figure 4.1 below) are generally detoxifying. Symptoms of acute toxic 

exposure include vomiting, unconsciousness, giddiness, sweating, diarrhea, tachycardia, muscle 

fasciculations, abdominal pain, and bronchospasm (ATSDR 1996). 

ATSDR has published an oral Minimal Risk Level for intermediate term exposure to diazinon of 0.0002 

mg/kg/day (ATSDR 1996). USEPA established a chronic reference dose (RfD) of 0.0002 mg/kg/day in 

the diet (USEPA 2000). 

4.2 Pharmacokinetics

Diazinon pharmacokinetics are qualitatively similar to other organophosphate pesticides described in 

this report. 

4.2.1 Absorption

Several oral absorption studies have been performed. 85% of the single oral dose of 4.0 mg/kg diazinon 

was absorbed by Beagle dogs in one study (Iverson et al. 1975). Other oral absorption studies were 

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conducted in rats, goats, sheep and cows (Abdelsalam and Ford 1986; Janes et al. 1973; Machin et al.

1974; Machin et al. 1971; Mount 1984; Wu et al. 1996a). In a dermal study, human volunteers absorbed 

34% of the dose applied to the abdomen or forearm for 24 hours (Wester et al. 1993). 

4.2.2 Distribution

Diazinon is found widely distributed in all tissues examined after oral absorption (Abdelsalam and Ford 

1986; de Blaquiere et al. 2000; Janes et al. 1973; Machin et al. 1974; Machin et al. 1971; Mucke et al.

1970; Tomokuni and Hasegawa 1985; Tomokuni et al. 1985). No studies of distribution after inhalation 

or dermal exposures are available. After an i.v. dose of 0.2 mg/kg in ethanol, the terminal halflife was 

1.5 hours (Iverson et al. 1975). Distribution coefficients for diazinon were reported (Garcia-Repetto et 

al. 1995). 

4.2.3 Metabolism

The principal metabolic pathways of diazinon are shown in Figure 4.1. 

N

N

CHCH3

CH3

CH3

O

P

CH3CH2

S

CH3CH2

N

N

CHCH3

CH3

CH3

O

P

CH3CH2

O

CH3CH2

N

N

CHCH3

CH3

CH3

HO

P

CH3CH2 OH

O

CH3CH2

P

CH3CH2 OH

S

CH3CH2

P450

Diazinon Diazoxon

2-isopropyl-4-methyl-6-hydroxypyrimidine

+ +

plus diethylphosphate or diethylthiophosphate

Figure 4.1 Principal metabolic pathways for diazinon. Adapted from Poet, 2003. 

P450 

Diazinon is subject to oxidative desulfurization and hydrolysis of the ester. Hydrolysis of the ester can 

occur before or after desulfurization, i.e., either diazinon or diazoxon can be hydrolyzed to yield 2-

isopropyl-4-methyl-6-hydroxypyrimidine (IMHP) and either diethylphosphate or diethylthiophosphate 

(Iverson et al. 1975; Machin et al. 1975; Mucke et al. 1970); however, the P450 catalyzed oxidative 

cleavage of phosphorothioate (i.e., diazinon) triester bond is much more efficient (Yang et al. 1971). 

Desulfurization is mediated by cytochrome P450 isoenzymes while oxidative cleavage or hydrolysis is 

mediated by cytochrome P450 or various esterases, respectively (Poet et al. 2003b; Walker and 

Mackness 1987). Diazinon and its metabolites may also be oxidized at alkyl carbons (Aizawa 1989; 

Yang et al. 1971). 
USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

In a human self-poisoning case, diazinon was found in serum and monoethyl phosphate, diethyl 

phosphate, and diethyl phosphorothioate were detected in the urine (Klemmer et al. 1978). In rat liver 

microsomes, diazoxon hydrolysis occurred without NADPH (Yang et al. 1971). 

Rates of metabolism in rat liver and intestinal microsomes of diazinon and diazoxon were recently 

reported by Poet and coworkers. The authors measured CYP450 mediated oxidative desulfurization (to 

form diazoxon) or hydrolysis (to form IMHP and diethylthiophosphate) as well as hydrolysis by esterase 

(PON1) in microsomes from both tissues. Based on the measured rates, the authors conclude that 

intestinal metabolism may be important, especially for low level oral doses (Poet et al. 2003b). Also in 

microsomes, Vittozzi et al. measured the activity of expressed cytochrome P450s for desulfurization and 

hydrolysis (Vittozzi et al. 2001). Significant activity was obtained for all nine cytochrome P450s tested 

that varied over about one order of magnitude (CYP2C19, 3A4, 2B6, 1A2, 1A1, 2C8, 2C9, 2D6, and 

2A6). A separate study by the same group indicated that CYP2C11, 3A2, and 2B1/2 were involved but 

that 2E1 and 1A1 were not (Fabrizi et al. 1999). Kappers et al. indicated that CYP2C19 was the major 

isoform involved in diazinon metabolism, but that others such as CYP1A2 and CYP3A4 may also be 

showing some activity (Kappers et al. 2001). However, using immunoinhibition and other techniques, 

Buratti and coworkers found that the principal isoforms involved in diazinon metabolism were CYP3A4, 

1A2, and 2B6 (Buratti et al. 2003), while Sams et al. felt that CYP 2D6 and 3A4 were the most 

important isozymes (Sams et al. 2000). 

Toxicity and acetylcholinesterase inhibition was studied in PON1 knockout mice (Li et al. 2000). 

Rates of inhibition of acetylcholinesterase were measured in some studies (Kamal and Al-Jafari 2000). 

4.2.4 Excretion

Most diazinon is excreted as metabolites in urine, while smaller amounts are excreted in feces or, after 

extensive metabolism, as CO2 in expired air. Approximately 75% of a 4.0 mg/kg oral dose to rats was 

excreted as urinary metabolites, 20% in the feces, and about 6% as CO2 (Mucke et al. 1970). 

Approximately 85% of total label was recovered in a 24 hour urine sample from dogs receiving a single 

oral dose of diazinon. After an i.v. dose, the dogs excreted approximately 58% of the label in urine 

(Iverson et al. 1975). In human volunteers, urinary excretion of diethyl phosphate and diethyl 

thiophosphate was reported after oral and dermal dosing (Cocker et al. 2002; Garfitt et al. 2002). 

Diazinon has been found in hair (of rabbits) as a potential biomarker of exposure (Tutudaki et al. 2003). 

Blood cholinesterase inhibition has also been used as a biomarker for diazinon exposure (Nigg and 

Knaak 2000). 

4.2.5 Special populations and variability

Several studies have suggested high variability in human metabolism of diazinon (Buratti et al. 2003; 

Kappers et al. 2001). Polymorphisms in PON1 have been described (Brophy et al. 2000; Cherry et al.

2002; Costa et al. 2003; Davies et al. 1996; Mackness et al. 2003). 

4.3 Interactions with other chemicals

The toxicity was increased and pharmacokinetics of diazinon were affected by cimetidine (Wu et al.

1996b). Interactions between diazinon and methyl parathion were reported in the blood and brain of 

pregnant rats and the fetus after a single dermal dose (Abu-Qare and Abou-Donia 2001). Neurite 

outgrowth was assessed for mixtures of diazinon and chlorpyrifos (Axelrad et al. 2002). 

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4.4 Diazinon PBPK models

One PBPK model for diazinon has been published in abstract form (Poet et al. 2003a). The model 

incorporated an oral exposure route, desulfurization and “hydrolysis” (an oxidation reaction actually) by 

a CYP450 enzyme, hydrolysis of the oxon by PON1 in liver and blood, and second order binding and 

inhibition and regeneration of B-esterases in the liver, blood, diaphragm, and brain. 

4.5 Literature Cited

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(PON 1) as a biomarker of susceptibility for organophosphate toxicity. Biomarkers 8, 1-12. 

Davies, H. G., Richter, R. J., Keifer, M., Broomfield, C. A., Sowalla, J., and Furlong, C. E. (1996). The 

effect of the human serum paraoxonase polymorphism is reversed with diazoxon, soman and sarin. 

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data from a human volunteer study with oral and dermal doses. Toxicol Lett 134, 105-13. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Gordon, C. J., and Mack, C. M. (2003). Influence of gender on thermoregulation and cholinesterase 

inhibition in the long-evans rat exposed to diazinon. J Toxicol Environ Health A 66, 291-304. 

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response without changes in peripheral cholinesterase enzymes following exposure to a sheep dip 

containing diazinon in vivo and in vitro. Mutat Res 472, 85-92. 

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metabolites of diazinon in sheep. J Agric Food Chem 21, 121-4. 

Kamal, M. A., and Al-Jafari, A. A. (2000). Dual substrate model for novel approach towards a kinetic 

study of acetylcholinesterase inhibition by diazinon. J Enzyme Inhib 15, 201-13. 

Kappers, W. A., Edwards, R. J., Murray, S., and Boobis, A. R. (2001). Diazinon is activated by 

CYP2C19 in human liver. Toxicol Appl Pharmacol 177, 68-76. 

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percent diazinon with treatment and recovery. Clin Toxicol 12, 435-44. 

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(2000). Catalytic efficiency determines the in-vivo efficacy of PON1 for detoxifying 

organophosphorus compounds. Pharmacogenetics 10, 767-79. 

Machin, A. F., Anderson, P. H., and Hebert, C. N. (1974). Residue levels and cholinesterase activities in 

sheep poinsoned experimentally with diazinon. Pest Sci 5, 49-56. 

Machin, A. F., Quick, M. P., Rogers, H., and Anderson, P. H. (1971). The conversion of diazinon to 

hydroxydiazinon in the guinea-pig and sheep. Bull Environ Contam Toxicol 6, 26-7. 

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Hepatic metabolism in the sheep, cow, pig, guinea-pig, rat, turkey, chicken and duck. Pest Sci 6, 

461-73. 

Mackness, B., Durrington, P., Povey, A., Thomson, S., Dippnall, M., Mackness, M., Smith, T., and 

Cherry, N. (2003). Paraoxonase and susceptibility to organophosphorus poisoning in farmers dipping 

sheep. Pharmacogenetics 13, 81-8. 

Mount, M. E. (1984). Diagnostic value of urinary dialkyl phosphate measurement in goats exposed to 

diazinon. Am J Vet Res 45, 817-24. 

Mucke, W., Alt, K. O., and Esser, H. O. (1970). Degradation of 14 C-labeled Diazinon in the rat. J Agric 

Food Chem 18, 208-12. 

Nigg, H. N., and Knaak, J. B. (2000). Blood cholinesterases as human biomarkers of organophosphorus 

pesticide exposure. Rev Environ Contam Toxicol 163, 29-111. 

Poet, T. S., Kousba, A. A., Wu, H., Dennison, S. L., and Timchalk, C. (2003a). Development of a 

physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the 

organophosphate pesticide, diazinon. The Toxicologist. Abstract # 1485.

Poet, T. S., Wu, H., Kousba, A. A., and Timchalk, C. (2003b). In vitro rat hepatic and intestinal 

metabolism of the organophosphate pesticides chlorpyrifos and diazinon. Toxicol Sci 72, 193-200. 

Sams, C., Mason, H. J., and Rawbone, R. (2000). Evidence for the activation of organophosphate 

pesticides by cytochromes P450 3A4 and 2D6 in human liver microsomes. Toxicol Lett 116, 217-21. 

Tomokuni, K., and Hasegawa, T. (1985). Diazinon concentrations and blood cholinesterase activities in 

rats exposed to diazinon. Toxicol Lett 25, 7-10. 

Tomokuni, K., Hasegawa, T., Hirai, Y., and Koga, N. (1985). The tissue distribution of diazinon and the 

inhibition of blood cholinesterase activities in rats and mice receiving a single intraperitoneal dose of 

diazinon. Toxicology 37, 91-8. 

Tutudaki, M., Tsakalof, A. K., and Tsatsakis, A. M. (2003). Hair analysis used to assess chronic 

exposure to the organophosphate diazinon: a model study with rabbits. Hum Exp Toxicol 22, 159-64. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

USEPA (2000). Refined Anticipated Residues/Acute and Chronic Dietary Risk Assessment (Including 

Beef Fat). U.S. Environmental Protection Agency. November 14, 2000. 

USEPA (2003). Restricted Use Pesticides Report. U.S. Environmental Protection Agency. Accessed 

9/28/03. http://www.epa.gov/opprd001/rup/rupjun03.htm. 

Vittozzi, L., Fabrizi, L., Di Consiglio, E., and Testai, E. (2001). Mechanistic aspects of 

organophosphorothionate toxicity in fish and humans. Environ Int 26, 125-9. 

Walker, C. H., and Mackness, M. I. (1987). "A" esterases and their role in regulating the toxicity of 

organophosphates. Arch Toxicol 60, 30-3. 

Wester, R. C., Sedik, L., Melendres, J., Logan, F., Maibach, H. I., and Russell, I. (1993). Percutaneous 

absorption of diazinon in humans. Food Chem Toxicol 31, 569-72. 

Wu, H. X., Evreux-Gros, C., and Descotes, J. (1996a). Diazinon toxicokinetics, tissue distribution and 

anticholinesterase activity in the rat. Biomed Environ Sci 9, 359-69. 

Wu, H. X., Evreux-Gros, C., and Descotes, J. (1996b). Influence of cimetidine on the toxicity and 

toxicokinetics of diazinon in the rat. Hum Exp Toxicol 15, 391-5. 

Yang, R. S., Hodgson, E., and Dauterman, W. C. (1971). Metabolism in vitro of diazinon and diazoxon 

in rat liver. J Agric Food Chem 19, 10-3. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Fenthion 

5.0 Introduction

Fenthion [O,O-dimethyl O-(4-(methylthio)-m-tolyl) phosphorothioate, (DMTP), Figure 1] is an 

organophosphorus insecticide used against mosquitoes, pests and bugs (EXTOXNET 1996). It is 

available in dust, emulsifiable or liquid concentrate, and granular and wettable powder formulations. 

Fenthion is soluble in organic solvents such as DMSO, acetone, methanol and ether, but not in water 

(NTP 2003). 

SCH3

CH3

P O

S

O

O

H3C

H3C

Figure 5.1: Chemical structure of fenthion 

5.1 Toxic effects

Fenthion is moderately toxic to laboratory animals (mice, rats, guinea pigs and rabbits). The acute oral 

and intraperitoneal LD50’s varied from approximately 125 to >1000 mg/kg body weight (DuBois and 

Kinoshita 1964; IPCS 1971; Ma 1995). Mortality of rats orally treated with fenthion in subchronic 

studies (30 mg/kg for 13 weeks, or 5.0 mg/kg for 3 months) was reported (NIH 1979). 

Fenthion predominately causes cholinergic toxicity in animals and humans. Its oxon inhibits plasma, 

erythrocyte and brain cholinesterase activity (Bai et al. 1990; De Bleecker et al. 1994; Dellinger and 

Mostrom 1988; Ma 1995; Misra et al. 1985; Misra et al. 1994; Tsatsakis et al. 2002; Tsatsakis et al.

1998). An acute no-observed-adverse –effect-level (NOAEL) of 0.07 mg/kg/day was determined in a 2-

year oral monkey study (USEPA 1999a, 2001). Other effects unrelated to cholinergic mechanisms, 

however, were also reported (Bagchi et al. 1995; Bagchi et al. 1996; Cova et al. 1995; Kojima et al.

1992). 

Fenthion did not show mutagenic effect in the bacterial reverse mutation test or the in vitro chromosome 

aberration test in Chinese hamster ovary cells, but did in unscheduled DNA synthesis study and mouse 

micronucleus assays (USEPA 1999a). In a 103-week chronic feeding study, no elevated incidence of 

tumor was observed in both sexes of F344 rats and female B6C3F1 mice; a slightly increased incidence 

of sarcomas, fibrosarcomas, and rhabdomyosarcomas of the integumentary system in male B6C3F1 

mice was observed (NIH 1979). Fenthion is not considered a carcinogen (Ma 1995; USEPA 1999a). 

5.2 Pharmacokinetics

5.2.1 Absorption

Depending on application, fenthion may be absorbed from the gastrointestinal tract, skin and respiratory 

tract. The former two pathways, however, have been more intensively studied. Generally, fenthion is 

readily absorbed from GI tract. Blood levels peak a few hours after oral dosing in rats (Ma 1995), rabbits 

(Emteres et al. 1985) and lactating goat (Ma 1995). Absorption was almost complete (96-100% at 72 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

hours) and not dose-dependent (at 10 or 100 mg/kg) in Wistar rats fasted for 16-24 hour before gavage 

(Ma 1995). 

Dermal absorption of fenthion is slow and incomplete. Eighteen hours after a single dermal dose, 

prepared as an application formulation, in pigs or lactating cows, the tissue residue levels were generally 

low, whereas at the application site the levels were much higher (Ma 1995). USEPA set a dermal 

absorption factor as 20% in 1996 and re-set it as 3% in 1999 based on the LOAEL’s (lowest observed 

adverse effect level) of cholinesterase inhibition from an oral development toxicity study and a 21-day 

dermal toxicity study in rabbits (USEPA 1999b). 

5.2.2 Distribution

From the limited information on its distribution in body, fenthion and its metabolites had relatively high 

concentration in fat, liver and kidney (EXTOXNET 1996). In milk from fenthion-treated dairy cows, 

the fenthion level was 50 times higher in the “fat” fraction than that in the “non-fat” fraction (O'Keeffe

et al. 1983). 

5.2.3 Metabolism

Principal metabolic pathways are shown in Figure 5.2. Fenthion has several possible oxidative 

metabolites in body such as sulfoxide (SO), sulfone (SO2), oxygen analogue (P=O), oxygen analogue 

sulfoxide (P=O, SO) and sulfone (P=O, SO2) (Wright and Riner 1979). Of them, the oxygen analogues 

are bioactivated forms with higher anti-cholinesterase activity (IPCS 1971, 1976). 

Incubated with rat liver microsomes, fenthion was oxidized to oxygen analogue and fenthion sulfoxide. 

Fenthion sulfone, however, was not detected. The main enzymes involved were cytochrome P450s 

(especially CYP1A1) and flavin-containing monooxygenase (Kitamura et al. 2003; Venkatesh et al.

1991). In liver cytosol of rats, fenthion sulfoxide was reduced to fenthion catalyzed by aldehyde oxidase 

(Kitamura et al. 2003). 

14C-Fenthion was extensively metabolized in rats (Ma 1995). No unchanged parent compound was 

detected in the urine and very little (< 2%) in the feces. The major group of metabolites (about 60% of 

the total label) was composed of the three phenols (phenol fenthion, phenol sulfoxide and phenol 

sulfone) and their glucuronide and sulfate conjugates. Four demethyl metabolites accounted for about 

30% of the label, whereas the oxygen analogue sulfoxide constituted only 1-4%. The metabolite profiles 

were not affected by dosing route, dose, sex or pre-treatment with fenthion. 

In pigs, fenthion was oxidized to fenthion sulfoxide, fenthion sulfone, oxygen analogue and oxygen 

analogue sulfoxide and sulfone. These metabolites were further hydrolyzed and excreted via urine in 

conjugated forms (Ma 1995). 

No data from human studies is available. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

SCH3

CH3

P O

S

CH3O

CH3O

SCH3

CH3

P O

S

CH3O

CH3O O

SCH3

CH3

P O

S

CH3O

CH3O O

O

SCH3

CH3

P O

O

CH3O

CH3O

SCH3

CH3

P O

O

CH3O

CH3O O

SCH3

CH3

P O

O

CH3O

CH3O O

O

SCH3

CH3

P O

S

HO

CH3O

SCH3

CH3

HO

SCH3

CH3

HO

O

SCH3

CH3

HO

O

O

Fenthion

Fenthion sulfoxide

Fenthion sulfone

Fenthion oxygen analogue

Fenthion oxygen analogue sulfoxide

Fenthion oxygen analogue sulfone

Demethylated fenthion

Phenol fenthion

Phenol fenthion sulfoxide

Phenol fenthion sulfone

Figure 5.2: Proposed metabolism pathways of fenthion in the rats. 

Adapted from Ma, 1995. 

5.2.4 Excretion

Fenthion was rapidly eliminated after a single dose in Wistar rats, over 90% of the administered 

radiolabel being excreted within 48 hours, and less than 1% retained 72 hours after treatment (Ma 1995). 

In pigs and dairy cows rapid elimination and low bioaccumulation were also observed (Ma 1995). In 

New Zealand white rabbits, the halflife of a single dose (20mg/kg) was about 11-12 hours regardless the 

route of administration (Emteres et al. 1985). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

The polar metabolites of fenthion are mainly excreted via urine in rats, pigs and dairy cows (Ma 1995). 

Milk is a significant pathway for the elimination of the parent compound from lactating dairy cows 

(IPCS 1971; O'Keeffe et al. 1983; Wright and Riner 1979). 

5.3 Interactions of fenthion with other OP pesticides

Fenthion potentiated the acute intraperitoneal toxicity of malathion, dioxathion, and coumaphos in rats, 

but intraperitoneal administration of 13 other organophosphate or carbamate insecticides to rats in 

combination with fenthion did not result in greater than additive toxic effects (Ma 1995). Dietary 

combination of equitoxic doses (2 mg/kg) of fenthion with coumaphos, neither of which alone affected 

cholinesterase activity when fed to dogs for six weeks, was found to potentiate the anticholinesterase 

activity in serum and erythrocytes by 75 and 30%, respectively. The potentiation was less evident with 

malathion, and no potentiation was noted with dioxathion (Ma 1995). Pretreatment with fenthion 

significantly potentiated the acute toxicity of 2-sec-butylphenyl N-methycarbamate (BPMC) in mice and 

dogs, which may be a result of the inhibited detoxification of the carbamate (Ma 1995; Miyaoka et al.

1984; Miyaoka et al. 1987). 

5.4 PBPK models

No PBPK models on fenthion have been reported yet. 

5.5 Literature Cited

Bagchi, D., Bagchi, M., Hassoun, E. A., and Stohs, S. J. (1995). In vitro and in vivo generation of 

reactive oxygen species, DNA damage and lactate dehydrogenase leakage by selected pesticides. 

Toxicology 104, 129-40. 

Bagchi, D., Bhattacharya, G., and Stohs, S. J. (1996). In vitro and in vivo induction of heat shock 

(stress) protein (Hsp) gene expression by selected pesticides. Toxicology 112, 57-68. 

Bai, C. L., Qiao, C. B., Zhang, W. D., Chen, Y. L., and Qu, S. X. (1990). A study of the pesticide 

fenthion: toxicity, mutagenicity, and influence on tissue enzymes. Biomed Environ Sci 3, 262-75. 

Cova, D., Perego, R., Nebuloni, C., Fontana, G., and Molinari, G. P. (1995). In vitro cytotoxicity of 

fenthion and related metabolites in human neuroblastoma cell lines. Chemosphere 30, 1709-15. 

De Bleecker, J., Lison, D., Van Den Abeele, K., Willems, J., and De Reuck, J. (1994). Acute and 

subacute organophosphate poisoning in the rat. Neurotoxicology 15, 341-8. 

Dellinger, J., and Mostrom, M. (1988). Effects of topical fenthion on blood cholinesterase and vagal 

tone in dogs. Vet Hum Toxicol 30, 229-34. 

DuBois, K. P., and Kinoshita, F. (1964). Acute toxicity and anti-cholinesterase action of O,O-dimethyl 

O-4-(methylthio)-m-tolyl phosphorothioate (DMTP; Baytex) and related compounds. Toxicol Appl 

Pharmacol 6, 86-95. 

Emteres, R., Abdelghani, A., and Anderson, A. C. (1985). Determination of the half life of fenthion in 

New Zealand White rabbits using three routes of administration. J Environ Sci Health B 20, 577-91. 

EXTOXNET (1996). Pesticide Information Profiles: Fenthion. Oregon State University. 

http://ace.ace.orst.edu/info/extoxnet/pips/fenthion.htm. 

IPCS (1971). WHO Pesticide Residues Series, No. 1. World Health Organization, Geneva. 

IPCS (1976). Data Sheets on Pesticides No. 23: Fenthion. World Health Organization, Geneva. 

Kitamura, S., Suzuki, T., Kadota, T., Yoshida, M., Ohashi, K., and Ohta, S. (2003). In vitro metabolism 

of fenthion and fenthion sulfoxide by liver preparations of sea bream, goldfish, and rats. Drug Metab 

Dispos 31, 179-86. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Kojima, T., Tsuda, S., and Shirasu, Y. (1992). Non-cholinergic mechanisms underlying the acute lethal 

effects of P = S type organophosphorus insecticides in rats. J Vet Med Sci 54, 529-33. 

Ma, S. (1995). Fenthion. Pesticide Residues in Food: 1995 evaluations Part II Toxicological & 

Environmental. Health Canada, Ottawa, Canada. 

Misra, U. K., Nag, D., Bhushan, V., and Ray, P. K. (1985). Clinical and biochemical changes in 

chronically exposed organophosphate workers. Toxicol Lett 24, 187-93. 

Misra, U. K., Prasad, M., and Pandey, C. M. (1994). A study of cognitive functions and event related 

potentials following organophosphate exposure. Electromyogr Clin Neurophysiol 34, 197-203. 

Miyaoka, T., Takahashi, H., Tsuda, S., and Shirasu, Y. (1984). Potentiation of acute toxicity of 2-secbutylphenyl N-methylcarbamate (BPMC) by fenthion in mice. Fundam Appl Toxicol 4, 802-7. 

Miyaoka, T., Tsuda, S., and Shirasu, Y. (1987). Effect of O, O-dimethyl O-(3-methyl-4-

methylthiophenyl) phosphorothioate (fenthion) pretreatment on acute toxicity of 2-sec-butylphenyl 

N-methylcarbamate (BPMC) in dogs. Nippon Juigaku Zasshi 49, 173-5. 

NIH (1979). Bioassay of fenthion for possible carcinogenicity. U.S. Department of Health, Education 

and Welfare, Bethesda, Maryland. DHEW Pub No. (NIH) 79-1353. 

NTP. NTP Chemical Repository: Baytex. http://ntp-server.niehs.nih.gov/htdocs/ 

CHEM_H&S/NTP_Chem5/Radian55-38-9.html 

O'Keeffe, M., Eades, J. F., and Strickland, K. L. (1983). Fenthion residues in milk and milk products 

following treatment of dairy cows for warble-fly. J Sci Food Agric 34, 192-197. 

Tsatsakis, A. M., Bertsias, G. K., Liakou, V., Mammas, I. N., Stiakakis, I., and Tzanakakis, G. N. 

(2002). Severe fenthion intoxications due to ingestion and inhalation with survival outcome. Hum 

Exp Toxicol 21, 49-54. 

Tsatsakis, A. M., Manousakis, A., Anastasaki, M., Tzatzarakis, M., Katsanoulas, K., Delaki, C., and 

Agouridakis, P. (1998). Clinical and toxicological data in fenthion and omethoate acute poisoning. J 

Environ Sci Health B 33, 657-70. 

USEPA (1999a). Human Health Risk Assessment: Fenthion. U.S. Environmental Protection Agency, 

Washington, D.C., October 13, 1999. 

USEPA (1999b). Fenthion: Re-evaluation of the dermal absorption factor. U.S. Environmental 

Protection Agency, Washington, D.C. HED DOC. No.013746. 

USEPA (2001). Interim Reregistration Eligibility Decision for Fenthion. U.S. Environmental Protection 

Agency, Washington, D.C. Report # EPA738-R-00-013. 

Venkatesh, K., Levi, P. E., and Hodgson, E. (1991). The effect of detergents on the purified flavincontaining monooxygenase of mouse liver, kidney and lungs. Gen Pharmacol 22, 549-52. 

Wright, F. C., and Riner, J. C. (1979). Biotransformation and deposition of residues of fenthion and 

oxidative metabolites in the fat of cattle. J Agric Food Chem 27, 576-7.

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Fenitrothion 

6.0 Introduction

Fenitrothion (O,O-dimethyl O-(3-methyl-4-nitrophenyl) phosphorodithioate), also commonly called 

SumithionTM, is an organophosphorus insecticide and was registered for use in ant and roach bait. There 

are no approved domestic food or feed uses for fenitrothion, and exposure to fenitrothion in the U.S. is 

minimal. However, fenitrothion is used in countries to control pests on crops, stored grains, and cotton. 

Fenitrothion is also used elsewhere in forest spraying and public health campaigns. As a result, the 

human health effects associated with exposure to fenitrothion remain a concern, especially among 

pesticide workers and applicators whose acute exposure to organophosphorus pesticides can sometimes 

occur at levels high enough to inhibit blood acetylcholinesterase activity (Nigg and Knaak 2000; OhayoMitoko et al. 2000; Satoh and Hosokawa 2000). 

Degradation rates reported for fenitrothion are as follows: 

• Soil and groundwater → T1/2 < one week (Meister 1994; U.S.EPA 1987) 

• Surface water → T1/2 = 1.5-2 days (Novathion 1987) 

• Surface water in dark → T1/2 = 21.6 - 49.5 days (Novathion 1987) 

• Plants → T1/2 = 1-2 days (Möllhoff 1968) 

• Plants (fenitrooxon) →T1/2 = few hours (Möllhoff 1968) 

6.1 Toxic effects

Like other organophosphorus compounds, fenitrothion acts in the organism as a cholinesterase inhibitor, 

after conversion to fenitrooxon. Some evidence indicates that acetylcholinesterase inhibition in brain 

depends more on the rate of penetration than on the rate of oxidation and decomposition of fenitrothion 

(JMPR 1988; Miyamoto 1969). Fenitrothion appears to affect cytochrome P450 enzyme activity in the 

liver and testes of rats (Clos et al. 1994; Gradowska-Olszewska et al. 1984). 

Fenitrothion has anticholinesterase activity and moderate acute toxicity with oral LD50 values in rats and 

mice ranging from 330 to 1,416 mg/kg body weight (Miyamoto et al. 1963b). Acute dermal toxicity in 

rodents is reported to range from 890 to more than 2,500 mg/kg body weight (WHO 1992). The LC50

value in rats exposed for 8 h is estimated to be more than 186 mg/m3

 (WHO 1992). In short-term studies 

on rats and dogs, no-observed-adverse-effect levels (NOAELs) based on brain cholinesterase activity 

were 10 mg/kg diet and 50 mg/kg diet, respectively. Long-term studies on rats and mice indicated a 

NOAEL of 10 mg/kg diet (WHO 1992). An acceptable daily intake (ADI) of 0.003 mg/kg body weight 

was established in 1984 (WHO 1986), but no occupational exposure limits (OEL) have been published. 

No carcinogenic effects were found in any of the long-term fenitrothion studies (WHO 1992). 

Fenitrothion was not found to be mutagenic in in vitro and in vivo studies or teratogenic at doses of up to 

30 mg/kg body weight in rabbits and up to 25 mg/kg body weight in rats (Benes et al. 1975; WHO 

1992). Other toxicity studies have been conducted (Chevalier et al. 1982; Groszek et al. 1995; Khan et 

al. 1990; Misu et al. 1966; Myatt et al. 1975; Trottier et al. 1980; Yoshida et al. 1987). Fenitrothion has 

also been shown to be neurotoxic, immunosuppressive, a pulmonary toxicant, and cause disturbances of 

prenatal development (Berlinska and Sitarek 1997; Kunimatsu et al. 1996; Khan et al. 1990; Lehotzky 

and Ungvary 1976). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

In a field spraying operation in Nigeria and Kenya, humans exposed to fenitrothion exhibited depressed 

plasma cholinesterase activity (Ohayo-Mitoko et al. 2000; Vandekar 1965; Wilford et al. 1965). A study 

of fenitrothion on 24 human subjects was also conducted and showed that both plasma and 

cholinesterase activity was not depressed in all but one case (Nosal and Hladka 1968). Fenitrothion has 

been reported as causing “intermediate syndrome” due to acute poisoning (Groszek et al. 1995). 

The presence of chemicals in the environment that have antiandrogenic activity and thus the ability to 

disrupt the endocrine system is a source of concern. In several studies fenitrothion has been shown to be 

a competitive androgen receptor antagonist both in vivo and in vitro (Curtis 2001; Sohoni et al. 2001; 

Tamura et al. 2001; Turner et al. 2002). One study, however, exhibited fenitrothion not having 

significant androgenic or antiandrogenic activity in vivo (Sunami et al. 2000). Fenitrothion might also 

alter estradiol metabolism by inhibition of certain P450 enzymes and produce changes in adrenal 

function (Berger and Sultatos 1997; Yamamoto et al. 1982b). 

6.2 Pharmacokinetics

Various studies in mouse, rat, guinea pig, and humans have dealt with the pharmacokinetic and 

biochemical aspects of fenitrothion and its metabolites (Aprea et al. 1999; Douch et al. 1968; Hladka 

and Nosal 1967; Hollingworth et al. 1967; Meaklim et al. 2003; Miyamoto 1964a; Miyamoto 1964b; 

Miyamoto and Sato 1969; Miyamoto 1969; Miyamoto et al. 1963a; Nishizawa et al. 1961; Vandanis and 

Crawford 1964). 

6.2.1 Absorption

Fenitrothion is presumably rapidly absorbed from the mammalian intestinal tract when given orally. 

Additionally, it can also be absorbed by the intact skin and by inhalation. (Kohli et al. 1974; Moody and 

Franklin 1987; Moody et al. 1987). 

6.2.2 Distribution

The presence of the oxygen analogue was demonstrated in all tissues examined (brain, heart, lung, liver, 

kidney, spleen, and muscle), and it was detectable in blood one min after intravenous injection of 

fenitrothion (Muller 2000). 

6.2.3 Metabolism

The oxygen analogue is the most important metabolite with respect to toxicity. It is formed in the 

microsomal fraction of the cell, the main organs responsible for the transformation being the liver and 

kidney. The major excretion product found is 3-methyl-4-nitrophenol which can be oxidized further to 

3-carboxyl-4-nitrophenol. Other metabolites are the dimethyl derivatives, which, with increasing dose, 

are excreted in increasing amounts. Nine metabolites have been isolated, most of which have also been 

identified. In vitro, formation of the oxygen analogue depended on the availability of reduced nicotine 

adenine dinucleotide phosphate (NADPH2) and oxygen (Miyamoto et al. 1963a; Miyamoto 1969). Liver 

slices incubated with fenitrothion did not produce measurable amounts of fenitrooxon, while liver 

homogenates and the supernatant fraction of such homogenates appreciably activated added fenitrothion 

(Miyamoto et al. 1963a; Miyamoto 1969). No correlation between the toxicity and rate of formation of 

fenitrooxon could, however, be demonstrated (JMPR 1988; Miyamoto et al. 1963a; Miyamoto 1969). 

No observations were made in these studies on the distribution into fatty tissues, but studies of residues 

in milk, meat, and fat from cattle indicated the presence of approximately 0.001 mg/kg in these samples 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

(JMPR 1988; Miyamoto and Sato 1969). Other studies involving the metabolism of fenitrothion have 

been described (Anjum and Qadri 1986; Kasagami et al. 2002; Sultatos 1991; Yamamoto et al. 1983; 

Yamamoto et al. 1982; Yoshida et al. 1975). 

Figure 6.1: Metabolic pathway of fenitrothion in vivo (Kumar et al. 1993). 

Esterase 

CYP450 

CH3

O NO2 P

S

CH3O

O-CH3

CH3

P O

O

CH3O

O-CH3

NO2

CH3

P OH NO2

S

CH3O

O-CH3

OH

Fenitrothion

Fenitroxon

P

O

CH3O

OH

O-CH3 CH3

OH NO2

+

+

CYP450 

6.2.4 Excretion

Fenitrothion and its metabolites are excreted mainly in the urine (90-95%) (Aprea et al. 1999; 

Hollingworth et al. 1967). Up to 10% was recovered in feces (Hollingworth et al. 1967). Within three 

days nearly complete recovery of an orally administered dose (15 mg/kg) could be obtained 

(Hollingworth et al. 1967). The ratios between the amounts of metabolites was dependent upon the dose 

given (Hollingworth et al. 1967). Other urinary excretion studies have been described (Aprea et al.

1999; Hladka and Nosal 1967; Kojima et al. 1989; Nosal and Hladka 1968). 

6.3 Interactions of fenitrothion with other chemicals

Interactions of fenitrothion with other compounds such as malathion (Hladka et al. 1974), diethyl 

maleate (Sultatos et al. 1991), 2-sec-butylphenol methylcarbamate (BPMC) (Takahashi et al. 1984), and 

N,N-diethyl-m-toluamide (DEET) (Moody et al. 1987) have also been studied. 

6.4 PBPK models

To date there are no published PBPK models for fenitrothion; however, there are numerous 

pharmacokinetic data that could be used in model development (Aprea et al. 1999; Douch et al. 1968; 

Hladka and Nosal 1967; Hollingworth et al. 1967; Kojima et al. 1989; Meaklim et al. 2003; Meaklim 

and McNeil 1999; Miyamoto 1964a; Miyamoto 1964b; Miyamoto and Sato 1969; Miyamoto 1969; 

Miyamoto et al. 1963a; Muller 2000; Nishizawa et al. 1961; Nosal and Hladka 1968; Vandanis and 

Crawford 1964). 

6.5 Literature Cited

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Anjum, F., and Qadri, S. S. (1986). In vivo metabolism of fenitrothion (0,0-dimethyl-0-(4-nitro-m-tolyl) 

phosphorothioate) in fresh water teleost (Tilapia mossambica). Bull Environ Contam Toxicol 36, 

140-145. 

Aprea, C., Sciarra, G., Sartorelli, P., Ceccarelli, F., and Centi, L. (1999). Multiroute exposure 

assessment and excretion of urinary metabolites of fenitrothion during manual operations on treated 

ornamental plants in greenhouses. Arch Environ Contam Toxicol 36, 490-497. 

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214. 

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methyl parathion. Influence of structure on anticholinesterase activity. Metabolism in the white 

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insecticides in the case of non-AChE inhibitory oxons. Pest Manag Sci 58, 1107-1117. 

Khan, M. F., Abidi, P., Anwer, J., Ray, P. K., and Anand, M. (1990). Pulmonary biochemical 

assessment of fenitrothion toxicity in rats. Bull Environ Contam Toxicol 45, 598-603. 

Kohli, J. D., Hasan, M. Z., and Gupta, B. N. (1974). Dermal absorption of fenitrothion in rat. Bull 

Environ Contam Toxicol 11, 285-290. 

Kojima, T., Yashiki, M., Miyazaki, T., Chikasue, F., and Ohtani, M. (1989). Detection of Smethylfenitrothion, aminofenitrothion, aminofenitroxon and acetylaminofenitroxon in the urine of a 

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and brain of rat. Biomed Chromatogr 7, 301-305. 

Kunimatsu, T., Kamita, Y., Isobe, N., and Kawasaki, H. (1996). Immunotoxicological insignificance of 

fenitrothion in mice and rats. Fundam Appl Toxicol 33, 246-253. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Lehotzky, K., and Ungvary, G. (1976). Experimental data on the neurotoxicity of fenitrothion. Acta 

Pharmacol Toxicol (Copenh) 39, 374-382. 

Meaklim, J., Yang, J., Drummer, O. H., Killalea, S., Staikos, V., Horomidis, S., Rutherford, D., 

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toxicokinetics and toxicologic evaluation in human volunteers. Environ Health Perspect 111, 305-

308. 

Meaklim, J. F., and McNeil, J. J. (1999). Fenitrothion ingestion in humans: Subacute effects. 

Unpublished report no. HT-0539 from the department of Epidemiology and Preventative Medicine. 

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Misu, Y., Segawa, T., Kuruma, I., Kojima, M., and Takagi, H. (1966). Subacute toxicity of O, odimethyl O-(3-methyl-4-nitrophenyl) phosphorothioate (Sumithion) in the rat. Toxicol Appl 

Pharmacol 9, 17-26. 

Miyamoto, J. (1964a). Studies on the mode of action of organophosphorus compounds. Part III. 

Activation and degradation of Sumithion and Methylparathion in vivo. Agric Biol Chem 28, 411-

421. 

Miyamoto, J. (1964b). Studies on the mode of action of organophosphorus compounds. Part IV. 

Pentration of Sumition, Methylparathion and their oxygen analogs into guinea pig brain and 

inhibition of cholinesterase in vivo. Agric Biol Chem 28, 422-430. 

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264. 

Miyamoto, J., and Sato, Y. (1969). Determination of insecticide residue in animal and plant tissues. VI. 

Determination of Sumithion residue in cattle tissues. Botyu Kagaku 34, 3-6. 

Miyamoto, J., Sato, Y., Kadota, T., Fujinami, A., and Endo, M. (1963a). Studies on the mode of action 

of organophosphorus compounds. Part I. Metabolic fate of P32 labelled Sumition and methyl 

parathion in the guinea-pig and white rat. Agric Biol Chem 27, 381-389. 

Miyamoto, J., Sato, Y., Kadota, T., Fujinami, A., and Endo, M. (1963b). Studies on the mode of action 

of organophosphorus compounds. Part II. Inhibition of mammalian cholinesterase in vivo following 

administration of Sumithion and methylparathion. Agric Biol Chem 27, 669-676. 

Möllhoff, E. (1968). Beit zur Frage der Rückstände und ihrer Bestimmung in Pflanzen nach Anwendung 

von Präparaten der E 605, pp. 331-58, Pflanzenschutz-Nachrichten Bayer. 

Moody, R. P., and Franklin, C. A. (1987). Percutaneous absorption of the insecticides fenitrothion and 

aminocarb in rats and monkeys. J Toxicol Environ Health 20, 209-218. 

Moody, R. P., Riedel, D., Ritter, L., and Franklin, C. A. (1987). The effect of DEET (N,N-diethyl-mtoluamide) on dermal persistence and absorption of the insecticide fenitrothion in rats and monkeys. 

J Toxicol Environ Health 22, 471-479. 

Muller, U. (2000). Pesticide residues in food 2000: Fenitrothion. Chemicals and Non-Presciption 

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Myatt, G. L., Ecobichon, D. J., and Greenhalgh, R. (1975). Fenitrooxon and S-methyl fenitrothion: acute 

toxicity and hydrolysis in mammals. Environ Res 10, 407-414. 

Nigg, H. N., and Knaak, J. B. (2000). Blood cholinesterases as human biomarkers of organophosphorus 

pesticide exposure. Rev Environ Contam Toxicol 163, 29-111. 

Nishizawa, Y., Fujii, K., Kadota, T., Miyamoto, J., and Sakamoto, H. (1961). Studies of the 

organophosphorus insecticides. Part VII. Chemical and biological preperties of new low toxic 

organophosphorus insecticide. O,O-Dimethyl-O-(3-methyl-4-nitrophenyl) thiophosphorothioate. 

Agric Biol Chem 25, 605-610. 

Nosal, M., and Hladka, A. (1968). Determination of the exposure to fenitrothion (0,0-dimethyl-0-3-

methyl-4-nitrophenyl-thiophosphate) on the basis of the excretion of p-nitro-m-cresol by the urine of 

the persons tested. Int Arch Arbeitsmed 25, 28-38. 

Novathion (1987). Data Manual. Cheminova Agro A/S, Lemvig, Denmark. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Ohayo-Mitoko, G. J., Kromhout, H., Simwa, J. M., Boleij, J. S., and Heederik, D. (2000). Self reported 

symptoms and inhibition of acetylcholinesterase activity among Kenyan agricultural workers. Occup 

Environ Med 57, 195-200. 

Satoh, T., and Hosokawa, M. (2000). Organophosphates and their impact on the global environment. 

Neurotoxicology 21, 223-227. 

Sohoni, P., Lefevre, P. A., Ashby, J., and Sumpter, J. P. (2001). Possible androgenic/anti-androgenic 

activity of the insecticide fenitrothion. J Appl Toxicol 21, 173-178. 

Sultatos, L. G. (1991). Metabolic activation of the organophosphorus insecticides chlorpyrifos and 

fenitrothion by perfused rat liver. Toxicology 68, 1-9. 

Sultatos, L. G., Huang, G. J., Jackson, O., Reed, K., and Soranno, T. M. (1991). The effect of 

glutathione monoethyl ester on the potentiation of the acute toxicity of methyl parathion, methyl 

paraoxon or fenitrothion by diethyl maleate in the mouse. Toxicol Lett 55, 77-83. 

Sunami, O., Kunimatsu, T., Yamada, T., Yabushita, S., Sukata, T., Miyata, K., Kamita, Y., Okuno, Y., 

Seki, T., Nakatsuka, I., and Matsuo, M. (2000). Evaluation of a 5-day Hershberger assay using 

young mature male rats: methyltestosterone and p,p'-DDE, but not fenitrothion, exhibited androgenic 

or antiandrogenic activity in vivo. J Toxicol Sci 25, 403-415. 

Takahashi, H., Miyaoka, T., Tsuda, S., and Shirasu, Y. (1984). Potentiated toxicity of 2-sec-butylphenyl 

methylcarbamate (BPMC) by O,O-dimethyl O-(3-methyl-4-nitrophenyl)phosphorothioate 

(fenitrothion) in mice; relationship between acute toxicity and metabolism of BPMC. Fundam Appl 

Toxicol 4, 718-723. 

Tamura, H., Maness, S. C., Reischmann, K., Dorman, D. C., Gray, L. E., and Gaido, K. W. (2001). 

Androgen receptor antagonism by the organophosphate insecticide fenitrothion. Toxicol Sci 60, 56-

62. 

Trottier, B., Fraser, A. R., Planet, G., and Ecobichon, D. J. (1980). Subacute toxicity of technical 

fenitrothion in male rats. Toxicology 17, 29-38. 

Turner, K. J., Barlow, N. J., Struve, M. F., Wallace, D. G., Gaido, K. W., Dorman, D. C., and Foster, P. 

M. (2002). Effects of in utero exposure to the organophosphate insecticide fenitrothion on androgendependent reproductive development in the Crl:CD(SD)BR rat. Toxicol Sci 68, 174-183. 

U.S.EPA (1987). Pesticide Fact Sheet Number 142. US EPA, Offices of Pesticide Programs, 

Registration Division, Washington, DC. 

Vandanis, A., and Crawford, L. G. (1964). Comparative metabolism of O,O-dimethyl-O-p-nitrophenyl 

phosphorothioate (methylparathion and O,O-dimethyl-O-(3-methyl-1-nitrophenyl) phosphorothioate 

(Sumithion). J Econ Entomol 57, 136-139. 

Vandekar, M. (1965). Observations on the toxicity of carbaryl, folithion and 3-isopropylphenyl nmethylcarbamate in a village-scale trial in Southern Nigeria. Bull World Health Organ 33, 107-115. 

WHO (1986). Environmental health criteria 63, Geneva. 

WHO (1992). Environmental health criteria 133, Geneva. 

Wilford, K., Lietaert, P. E. A., and Foll, C. V. (1965). Toxicological observations during large scale 

field trial of OMS-43 in Northern Nigeria (preliminary report), Geneva, 18-24 February 1965. 

Yamamoto, T., Egashira, T., Yoshida, T., and Kuroiwa, Y. (1982a). Comparison of the effect of an 

equimolar and low dose of fenitrothion and methylparathion on their own metabolism in rat liver. J 

Toxicol Sci 7, 35-41. 

Yamamoto, T., Egashira, T., Yoshida, T., and Kuroiwa, Y. (1982b). Increase of adrenal weight in rats by 

the repeated administration of fenitrothion. Toxicol Lett 11, 187-191. 

Yamamoto, T., Egashira, T., Yoshida, T., and Kuroiwa, Y. (1983). Comparative metabolism of 

fenitrothion and methylparathion in male rats. Acta Pharmacol Toxicol (Copenh) 53, 96-102. 

Yoshida, M., Shimada, E., Yamanaka, S., Aoyama, H., Yamamura, Y., and Owada, S. (1987). A case of 

acute poisoning with fenitrothion (Sumithion). Hum Toxicol 6, 403-406. 

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Yoshida, T., Homma, K., Suzuki, Y., and Uchiyama, M. (1975). Effect of fenitrothion on hepatic 

microsomal components of drug metabolizing system in mice. Chem Pharm Bull (Tokyo) 23, 2155-

2157. 

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Chloroform 

7.0 Introduction

Chloroform (trichloromethane, CHCl3) is a dense liquid that is insoluble in water and volatile under 

environmental conditions (McCulloch 2003). The major domestic use for chloroform is in the 

manufacture of refrigerant HCFC-22 (Chemical Marketing Reporter 1995). Chloroform is also used as 

a laboratory reagent and extraction solvent for pharmaceuticals. A significant amount of chloroform has 

been released to the environment as a by-product of the treatment of drinking and waste waters and 

through reactions of chlorine with organic chemicals (Meek et al. 2002). 

7.1 Mechanisms of toxicity

 

Oral and inhalation exposures to chloroform cause toxicity to the liver, kidney, and nasal epithelium 

(USEPA 2001). Chloroform can also cause reproductive or developmental toxicity, although most of 

the effects are secondary to maternal toxicity (USEPA 2001). Increased incidences of liver and kidney 

tumors have been observed in several animal species after exposures to chloroform via several routes, 

although there is no adequate human data for carcinogenicity (USEPA 2001). The mode of toxicity of 

chloroform is probably through oxidative metabolism to form phosgene (Pohl et al. 1977), which can 

react to form covalent bonds with microsomal proteins (Corley et al. 1990; Rosenthal 1987). 

7.2 Pharmacokinetics

7.2.1 Absorption

 

Chloroform is generally absorbed rapidly in humans and animals. It is easily absorbed into the blood 

from the lungs after inhalation exposures. Human inhalation studies include exposures via surgical 

anesthesia (Smith et al. 1973), indoor swimming pools (Aggazzotti et al. 1993; Cammann and Hubner 

1995; Levesque et al. 1994; Levesque et al. 2000), and shower air (Jo et al. 1990; Levesque et al. 2002). 

Chloroform can also be absorbed through the skin easily. Dermal exposures were considered in 

conjunction with inhalation exposures in some of the indoor swimming pool studies (Cammann and 

Hubner 1995; Levesque et al. 1994; Levesque et al. 2000) and shower air studies (Jo et al. 1990; 

Levesque et al. 2002). Other human studies of dermal-only exposures include showering with facemask 

(Corley et al. 2000; Gordon et al. 1998) and topical administration of chloroform to volunteers (Dick et 

al. 1995). Dermal absorption in animals was studied in guinea pigs (Jakobson et al. 1982) and hairless 

rats (Islam et al. 1995). Gastrointestinal absorption of chloroform is also fast and extensive (USEPA 

2001). Oral exposure studies of humans were done in volunteers using 13C-labeled chloroform (Fry et 

al. 1972) and additional information is available from an accidental chloroform poisoning case (Rao et 

al. 1993). Animal studies via oral exposure were reported in mice and rats by Withey et al. (Withey et 

al. 1983) and Pereira (Pereira 1994), respectively. 

7.2.2 Distribution

Chloroform is widely distributed throughout the body after being absorbed. Radiolabeled chloroform in 

mice was reported to distribute to the liver, kidney, lungs, spleen, body fat, muscle, and nervous tissue 

(Bergman 1979; Cohen and Hood 1969). The highest levels of chloroform detected in human 

postmortem samples are in the body fat (5–68 μg/kg) and lower levels (1–10 μg/kg) were detected in the 

kidney, liver, and brain (McConnell et al. 1975). In a study with 14C-chloroform injected in male mice 

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intraperitoneally, the maximum radioactivity levels were observed in the liver, kidney, and blood within 

10 minutes of dosing (Gemma et al. 1996). It was also found that the presence of testosterone affected 

chloroform accumulations in mouse kidney (Ilett et al. 1973; Pohl et al. 1984; Smith et al. 1973) and 

resulted in higher nephrotoxicity in male mice. 

7.2.3 Metabolism

The major metabolic pathway of chloroform in humans and animals (Figure 7.1) is oxidative 

metabolism that produces reactive phosgene and the minor pathway is reductive metabolism that forms 

dichloromethyl free radical (USEPA 2001). In the presence of oxygen, chloroform is converted to 

trichloromethanol, which spontaneously dehydrochlorinates to produce phosgene (Pohl et al. 1981; 

Stevens and Anders 1981). These reactions are catalyzed by cytochrome P450 in liver and kidneys (Ade

et al. 1994; Branchflower et al. 1984; Smith and Hook 1984). Phosgene can in turn react with 

nucleophilic groups in cellular macromolecules and form covalent adducts (Noort et al. 2000; Pereira 

and Chang 1981; Pereira et al. 1984; Pohl et al. 1977; Pohl et al. 1981; Pohl et al. 1980). Phosgene can 

also undergo hydrolysis to form carbon dioxide and hydrochloric acid, or react with glutathione to form 

diglutathionyl dithiocarbonate, gluathione disulfide, and carbon monoxide (Pohl et al. 1981; USEPA 

2001). In the absence of oxygen, chloroform is converted to dichloromethyl free radical, which can 

form covalent adducts with microsomal enzymes and can also cause lipid peroxidation (USEPA 2001). 

Metabolic pathways of chloroform overlapping with the other three volatile organics in Mixture 2 are 

shown in Figure 8.1 under trichloroethylene. 

Figure 7.1 Major metabolic pathways of chloroform (adapted from USEPA (2001)) 

CHCl3

P450 

HOCCl3

O2 H2O

HCl 

O = CCl2 2HCl + CO2

Phosgene

H2O

Cysteine 

H2

C CH

S NH

C

O

CO2H

2-oxothiazolidine4-carboxylic acid 

C Cl

O

GS

GSH 

HCl 

C SG

O

GS

S-chloro-carbonyl glutathione 

diglutathionyl dithiocarbonate 

CO + GSSG 

Adducts 
USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

7.2.4 Excretion

Chloroform is excreted through the lungs either unchanged or as carbon dioxide, with small amounts 

detected in urine after inhalation (Corley et al. 1990; Fry et al. 1972; Gordon et al. 1988), oral (Fry et al.

1972), and dermal (Dick et al. 1995) exposures. 

7.3 Physiologically based pharmacokinetic (PBPK) models

The first PBPK model for chloroform was developed by Corley and colleagues (Corley et al. 1990) to 

describe the fate of chloroform in several species via numerous exposure routes. Several subsequent 

PBPK models (Chinery and Gleason 1993; Corley et al. 2000; Gearhart et al. 1993; Levesque et al.

2000; McKone 1993; Roy et al. 1996) were developed based on the Corley model to include a variety of 

exposure scenarios. A schematic representation of a general PBPK model for chloroform is shown in 

Figure 7.2. 

Richly Perfused Tissues 

Lung Blood 

Slowly Perfused Tissues 

Alveolar Air 

Fat 

Kidney 

Skin 

Liver 

Intestine Stomach

Dermal Exposure 

Oral Exposure 

Intraperioneal 

Injection 

Exhaled Air 

Intravenous 

Infusion 

Inhalation 

Exposure 

Metabolites 

Figure 7.2 PBPK model for chloroform in rats, mice, and human (adapted from (Corley et al. 2000)) 

In the Corley model (Corley et al. 1990), the exposure routes include oral, inhalation, and 

intraperitoneal. Liver and kidney were both sites of metabolism for chloroform. The amount of 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

metabolite binding to cellular macromolecules was used as the indicator for chloroform toxicity. Due 

to the lower rates of metabolism, ventilation, and cardiac output in larger species than in smaller species, 

the relative potency of chloroform was predicted as mice > rats > humans in the Corley model (Corley et 

al. 1990). 

Reitz and colleagues (Reitz et al. 1990) extended the Corley model to include pharmacodynamic 

endpoints for cancer risk assessment. Two dose metrics were used for the liver compartment, while the 

kidney compartment was not considered. The first type of dose metric used was the average daily 

macromolecular binding. The other type of dose surrogate was cytotoxicity due to the formation of 

reactive chloroform metabolite, phosgene. It was concluded that cytotoxicity is the dose metric best 

reflecting carcinogenicity (Reitz et al. 1990). These two dose metrics were later analyzed for 

interindividual variability and parameter uncertainty by Allen and colleagues (Allen et al. 1996). The 

cytotoxicity dose metric was much more sensitive to interindividual variability than the average daily 

macromolecular binding was. 

Gearhart and colleagues (Gearhart et al. 1993) adjusted partition coefficients, rate of metabolism, 

cardiac output, and minute ventilation according to body temperature. These adjustments strengthened 

the Corley model (Corley et al. 1990) according to the fitting of gas uptake data of mice by loosening 

the assumption of enzyme loss and resysthesis. 

Chinery and Gleason (Chinery and Gleason 1993) further included the skin compartment to describe the 

fate of chloroform after adsorption through dermal exposure. The skin compartment was further divided 

into three subcompartments: the aqueous solution, stratum corneum, and viable skin. The model was 

able to predict the concentration of chloroform in the exhaled air from humans exposed while showering 

through inhalation only and the combination of dermal and inhalation routes. 

In a PBPK model similar to Chinery and Gleason’s, McKone (McKone 1993) assumed skin to be only 

one compartment. It was demonstrated that chloroform metabolism by the liver was not linear with 

respect to higher exposure concentrations (60-100 mg/L). 

Based on the Corley model (Corley et al. 1990), Levesque and colleagues (Levesque et al. 2000) used a 

PBPK model to predict the fate of chloroform for individuals exposed while swimming through dermal 

and inhalation routes. Dermal exposure was described using an overall skin permeability constant. The 

levels of macromolecular binding in swimmers calculated from the model are much lower than the 

smallest no-observed-effect level for liver tumors in animals. 

Corley and colleagues (Corley et al. 2000) added a single skin compartment to the original Corley 

model (Corley et al. 1990) and described the kinetics of human dermal exposure to chloroform while 

bathing. With the adjustment of model parameters according to temperature, the model can predict the 

relationship between water temperature (30-40°C) and exhaled chloroform observed from experiments 

(Gordon et al. 1998). 

Constan and colleagues (Constan et al. 2002) used cytolethality and regerative cell proliferation as 

pharmacodynamic endpoints to perform a chloroform inhalation cancer risk assessment. The NOAEL 

for chloroform-induced hepatotoxicity in humans was estimated to be 110 ppm using experimental data 

from B6C3F1 mouse and PBPK-PD model calculations. 

Meek and colleagues (Meek et al. 2002) recently performed an assessment of exposure-response 

analyses and risk characterization using PBPK models. Inhalation, oral, and dermal exposures were 

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considered from ten-minute shower, discrete periods of water and food consumption, as well as 

inhalation of chloroform at a variety of concentrations. Dose metrics used for carcinogenicity were the 

maximum rate of metabolism per unit kidney cortex volume and mean rate of metabolism per unit 

kidney cortex volume during each dose interval. For non-neoplastic effects, the dose metrics used were 

the mean rate of metabolism per unit centrilobular region of the liver and the average concentration of 

chloroform in the non-metabolizing centrilobular region of the liver. 

7.4 Literature Cited

Ade, P., Guastadisegni, C., Testai, E., and Vittozzi, L. (1994). Multiple activation of chloroform in 

kidney microsomes from male and female DBA/2J mice. Journal of Biochem Pharmacol 9, 289-295. 

Aggazzotti, G., Fantuzzi, G., Righi, E., Tartoni, P., Cassinadri, T., and Predieri, G. (1993). Chloroform 

in alveolar air of individuals attending indoor swimming pools. Arch Environ Health 48, 250-254. 

Allen, B. C., Covington, T. R., and Clewell, H. J. (1996). Investigation of the impact of pharmacokinetic 

variability and uncertainty on risks predicted with a pharmacokinetic model for chloroform. 

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Bergman, K. (1979). Whole-body autoradiography and allied tracer techniques in distribution and 

elimination studies of some organic solvents: benzene, toluene, xylene, styrene, methylene chloride, 

chloroform, carbon tetrachloride and trichloroethylene. Scand J Work Environ Health 5 Suppl 1, 1-

263. 

Branchflower, R. V., Nunn, D. S., Highet, R. J., Smith, J. H., Hook, J. B., and Pohl, L. R. (1984). 

Nephrotoxicity of chloroform: metabolism to phosgene by the mouse kidney. Toxicol Appl 

Pharmacol 72, 159-168. 

Cammann, K., and Hubner, K. (1995). Trihalomethane concentrations in swimmers' and bath attendants' 

blood and urine after swimming or working in indoor swimming pools. Arch Environ Health 50, 61-

65. 

Chemical Marketing Reporter. (1995). Chemical profile: Chloroform. Schnell Publishing, New York, 

NY. 

Chinery, R. L., and Gleason, A. K. (1993). A compartmental model for the prediction of breath 

concentration and absorbed dose of chloroform after exposure while showering. Risk Anal 13, 51-62. 

Cohen, E. N., and Hood, N. (1969). Application of low-temperature autoradiography to studies of the 

uptake and metabolism of volatile anesthetics in the mouse. I. Chloroform. Anesthesiology 30, 306-

314. 

Constan, A. A., Wong, B. A., Everitt, J. I., and Butterworth, B. E. (2002). Chloroform inhalation 

exposure conditions necessary to initiate liver toxicity in female B6C3F1 mice. Toxicol Sci 66, 201-

208. 

Corley, R. A., Gordon, S. M., and Wallace, L. A. (2000). Physiologically based pharmacokinetic 

modeling of the temperature-dependent dermal absorption of chloroform by humans following bath 

water exposures. Toxicol Sci 53, 13-23. 

Corley, R. A., Mendrala, A. L., Smith, F. A., Staats, D. A., Gargas, M. L., Conolly, R. B., Andersen, M. 

E., and Reitz, R. H. (1990). Development of a physiologically based pharmacokinetic model for 

chloroform. Toxicol Appl Pharmacol 103, 512-527. 

Dick, D., Ng, K. M., Sauder, D. N., and Chu, I. (1995). In vitro and in vivo percutaneous absorption of 

14C-chloroform in humans. Hum Exp Toxicol 14, 260-265. 

Fry, B. J., Taylor, T., and Hathway, D. E. (1972). Pulmonary elimination of chloroform and its 

metabolite in man. Arch Int Pharmacodyn Ther 196, 98-111. 

Gearhart, J. M., Seckel, C., and Vinegar, A. (1993). In vivo metabolism of chloroform in B6C3F1 mice 

determined by the method of gas uptake: the effects of body temperature on tissue partition 

coefficients and metabolism. Toxicol Appl Pharmacol 119, 258-266. 

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Gemma, S., Faccioli, S., Chieco, P., Sbraccia, M., Testai, E., and Vittozzi, L. (1996). In vivo CHCl3

bioactivation, toxicokinetics, toxicity, and induced compensatory cell proliferation in B6C3F1 male 

mice. Toxicol Appl Pharmacol 141, 394-402. 

Gordon, S. M., Wallace, L. A., Callahan, P. J., Kenny, D. V., and Brinkman, M. C. (1998). Effect of 

water temperature on dermal exposure to chloroform. Environ Health Perspect 106, 337-345. 

Gordon, S. M., Wallace, L. A., Pellizzari, E. D., and Oneill, H. J. (1988). Human breath measurements 

in a clean-air chamber to determine half-lives for volativle organic-compounds. Atmos Environ 22, 

2165-2170. 

Ilett, K. F., Reid, W. D., Sipes, I. G., and Krishna, G. (1973). Chloroform toxicity in mice: correlation of 

renal and hepatic necrosis with covalent binding of metabolites to tissue macromolecules. Exp Mol 

Pathol 19, 215-229. 

Islam, M. S., Zhao, L., McDougal, J. N., and Flynn, G. L. (1995). Uptake of chloroform by skin during 

short exposures to contaminated water. Risk Anal 15, 343-352. 

Jakobson, I., Wahlberg, J. E., Holmberg, B., and Johansson, G. (1982). Uptake via the blood and 

elimination of 10 organic solvents following epicutaneous exposure of anesthetized guinea pigs. 

Toxicol Appl Pharmacol 63, 181-187. 

Jo, W. K., Weisel, C. P., and Lioy, P. J. (1990). Routes of chloroform exposure and body burden from 

showering with chlorinated tap water. Risk Anal 10, 575-580. 

Levesque, B., Ayotte, P., LeBlanc, A., Dewailly, E., Prud'Homme, D., Lavoie, R., Allaire, S., and 

Levallois, P. (1994). Evaluation of dermal and respiratory chloroform exposure in humans. Environ 

Health Perspect 102, 1082-1087. 

Levesque, B., Ayotte, P., Tardif, R., Charest-Tardif, G., Dewailly, E., Prud'Homme, D., Gingras, G., 

Allaire, S., and Lavoie, R. (2000). Evaluation of the health risk associated with exposure to 

chloroform in indoor swimming pools. J Toxicol Environ Health A 61, 225-243. 

Levesque, B., Ayotte, P., Tardif, R., Ferron, L., Gingras, S., Schlouch, E., Gingras, G., Levallois, P., and 

Dewailly, E. (2002). Cancer risk associated with household exposure to chloroform. J Toxicol 

Environ Health A 65, 489-502. 

McConnell, G., Ferguson, D. M., and Pearson, C. R. (1975). Chlorinated hydrocarbons and the 

environment. Endeavour 34, 13-18. 

McCulloch, A. (2003). Chloroform in the environment: occurrence, sources, sinks and effects. 

Chemosphere 50, 1291-1308. 

McKone, T. E. (1993). Linking a PBPK model for chloroform with measured breath concentrations in 

showers: implications for dermal exposure models. J Expo Anal Environ Epidemiol 3, 339-365. 

Meek, M. E., Beauchamp, R., Long, G., Moir, D., Turner, L., and Walker, M. (2002). Chloroform: 

exposure estimation, hazard characterization, and exposure-response analysis. J Toxicol Environ 

Health B Crit Rev 5, 283-334. 

Noort, D., Hulst, A. G., Fidder, A., van Gurp, R. A., de Jong, L. P., and Benschop, H. P. (2000). In vitro 

adduct formation of phosgene with albumin and hemoglobin in human blood. Chem Res Toxicol 13, 

719-726. 

Pereira, M. A. (1994). Route of administration determines whether chloroform enhances or inhibits cell 

proliferation in the liver of B6C3F1 mice. Fundam Appl Toxicoly 23, 87-92. 

Pereira, M. A., and Chang, L. W. (1981). Binding of chemical carcinogens and mutagens to rat 

hemoglobin. Chem Biol Interact 33, 301-305. 

Pereira, M. A., Chang, L. W., Ferguson, J. L., and Couri, D. (1984). Binding of chloroform to the 

cysteine of hemoglobin. Chem Biol Interact 51, 115-124. 

Pohl, L. R., Bhooshan, B., Whittaker, N. F., and Krishna, G. (1977). Phosgene: a metabolite of 

chloroform. Biochem Biophys Res Commun 79, 684-691. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Pohl, L. R., Branchflower, R. V., Highet, R. J., Martin, J. L., Nunn, D. S., Monks, T. J., George, J. W., 

and Hinson, J. A. (1981). The formation of diglutathionyl dithiocarbonate as a metabolite of 

chloroform, bromotrichloromethane, and carbon tetrachloride. Drug Metab Dispos 9, 334-339. 

Pohl, L. R., George, J. W., and Satoh, H. (1984). Strain and sex differences in chloroform-induced 

nephrotoxicity. Different rates of metabolism of chloroform to phosgene by the mouse kidney. Drug 

Metab Dispos 12, 304-308. 

Pohl, L. R., Martin, J. L., and George, J. W. (1980). Mechanism of metabolic activation of chloroform 

by rat liver microsomes. Biochem Pharmacol 29, 3271-3276. 

Rao, K. N., Virji, M. A., Moraca, M. A., Diven, W. F., Martin, T. G., and Schneider, S. M. (1993). Role 

of serum markers for liver function and liver regeneration in the management of chloroform 

poisoning. J Anal Toxicol 17, 99-102. 

Reitz, R. H., Mendrala, A. L., Corley, R. A., Quast, J. F., Gargas, M. L., Andersen, M. E., Staats, D. A., 

and Conolly, R. B. (1990). Estimating the risk of liver cancer associated with human exposures to 

chloroform using physiologically based pharmacokinetic modeling. Toxicol Appl Pharmacol 105, 

443-459. 

Rosenthal, S. L. (1987). A review of the mutagenicity of chloroform. Environ Mol Mutagen 10, 211-

226. 

Roy, A., Weisel, C. P., Lioy, P. J., and Georgopoulos, P. G. (1996). A distributed parameter 

physiologically-based pharmacokinetic model for dermal and inhalation exposure to volatile organic 

compounds. Risk Anal 16, 147-160. 

Smith, A. A., Volpitto, P. P., Gramling, Z. W., DeVore, M. B., and Glassman, A. B. (1973). 

Chloroform, halothane, and regional anesthesia: a comparative study. Anesth Analg 52, 1-11. 

Smith, J. H., and Hook, J. B. (1984). Mechanism of chloroform nephrotoxicity. III. Renal and hepatic 

microsomal metabolism of chloroform in mice. Toxicol Appl Pharmacol 73, 511-524. 

Stevens, J. L., and Anders, M. W. (1981). Effect of cysteine, diethyl maleate, and phenobarbital 

treatments on the hepatotoxicity of [1H]chloroform. Chem Biol Interact 37, 207-217. 

USEPA (2001). Toxicological Review of Chloroform. U.S. Environmental Protection Agency, 

Washington, DC. 

Withey, J. R., Collins, B. T., and Collins, P. G. (1983). Effect of vehicle on the pharmacokinetics and 

uptake of four halogenated hydrocarbons from the gastrointestinal tract of the rat. J Appl Toxicol 3, 

249-253. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Trichloroethylene 

8.0 Introduction

Trichloroethylene (TCE) is one of the most important industrial chemicals of our time. It is an organic 

solvent that has been used widely in dry cleaning, metal degreasing, and as a solvent for oils and resins. 

Because of the large production volume and its wide applications, TCE is one of the top, if not the top, 

environmental pollutants in ground water. Recently, an entire supplemental volume (Volume 108, 

Supplement 2, May 2000) of Environmental Health Perspective was devoted to Trichloroethylene 

Health Risks. This volume contains many excellent review articles which cover the areas of toxicology 

and risk assessment extensively. The USEPA, in its re-assessment of TCE health risks, devoted a great 

deal of effort to publishing a document on Trichloroethylene Health Risk Assessment: Synthesis and 

Characterization (EPA/600/P-01/002A; quoted in this write-up as USEPA, 2001). This document and 

its related Science Advisory Board review (www.epa.gov/science1/pdf/ehc03002.pdf) also provided an 

excellent source of information on TCE. Thus, the summary below represents a brief update of the 

current information. 

8.1 Toxic effects

 

8.1.1 Noncancer effects 

Neuro- or neuro-behavioral (Boyes et al., 2000; Ohta et al., 2001; USEPA, 2001; Waseem et al., 2001; 

Kilburn, 2002; Moser et al., 2003), male reproductive (Kumar et al., 2000; 2001; Forkert et al., 2002; 

2003), developmental (Boyer et al., 2000; Rodenbeck et al., 2000; USEPA, 2001; Johnson et al., 2003), 

liver (USEPA, 2001), renal (USEPA, 2001; Mensing et al., 2002), immuno- (Griffin et al., 2000a,b,c; 

Kaneko et al., 2000; USEPA, 2001) toxicities in experimental animals and/or humans are reported or 

implicated. In vitro studies using isolated cell cultures have demonstrated and reconfirmed many of the 

species-, sex-, and tissue-dependent differences in hepato- and renal- toxicities observed in vivo

(Cummings et al., 2000a,b; Lash et al., 2001). 

Halogenated hydrocarbons such as TCE are among the most common water supply contaminants in the 

U.S. and elsewhere. Epidemiological studies have found an association, but not a cause-and-effect 

relation, between halogenated hydrocarbon contamination and increased incidence of congenital cardiac 

malformations or other defective birth outcomes. However, some animal studies in birds and rats as 

well as in tissue cultures had demonstrated statistically significant increased incidence of congenital 

cardiac malformations or other defective birth outcomes (Boyer et al., 2000; Johnson et al., 2003) while 

others turned out negative (Fisher et al., 2001). The most recent study (Johnson et al., 2003) reported 

that maternal rats exposed to more than 250 ppb TCE, a very low dose study, showed an associated 

increased incidence of cardiac malformations in their developing fetuses. 

8.1.2 Cancer effects 

TCE causes liver, lung tumors and lymphomas in mice and kidney and testicular tumors in rats (Bull, 

2000; Green, 2000; Lash et al., 2000a; USEPA, 2001). In humans, TCE was implicated to be a 

carcinogen (Wartenberg et al., 2000; USEPA, 2001). It is well established that two metabolites of TCE, 

dichloroacetic acid (DCA) and trichloroacetic acid (TCA) are important contributors to carcinogenicity 

of TCE (Bull, 2000; Tao et al., 2000; Bull et al., 2002). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Regarding renal cancer in humans, German epidemiological studies of prevalence of renal cancer 

following high exposure of TCE in workers have been the subject of considerable scientific debate, reevaluation, and repeated studies (Brauch et al., 1999; Bruning et al., 1999; Green and Lash, 1999; 

Schraml et al., 1999; Brauch et al., 2000; Bruning and Bolt, 2000; Green, 2000; Lash et al., 2000a; 

Wartenberg et al., 2000; USEPA, 2001; Bruning et al., 2003). 

8.2 Pharmacokinetics

The pharmacokinetics of TCE has been reviewed thoroughly (ATSDR, 1997; Fisher, 2000). More 

recent updates are provided below. 

8.2.1 Absorption

Dose-dependent gastrointestinal absorption of TCE and its kinetics in male Sprague-Dawley rats over a 

wide range of oral bolus doses were characterized by Lee et al. (2000b). Dietary incorporation of guar 

gum, a thickener and stabilizer in foods and pharmaceuticals, was found to decrease TCE accumulation 

in the body by reducing absorption and fat tissue mass (Nakashima and Ikegami, 2001). 

8.2.2 Distribution

PBPK models for the systemic transport of TCE to various tissues and organs with a special emphasis to 

fat tissues were established by Albanese et al. (2002). 

8.2.3 Metabolism

Human and animal studies 

The principal metabolic pathways for TCE and metabolic steps where interactions with chloroform 

(CHL), tetrachloroethylene (PERC), and/or 1,1,1 trichloroethane (MC) may occur are denoted in Figure 

8.1. The metabolism of TCE has been reviewed thoroughly (Lash et al., 2000b; USEPA, 2001); more 

recent updates are provided below. 

Lash et al. (1999a) reported direct, in vivo, evidence of GSH conjugation of TCE in human volunteers 

exposed to 100 ppm TCE and demonstrated markedly higher amounts of S-(1,2-dichlorovinyl) 

glutathione (DCVG) in males than females. However, Bloemen et al. (2001) studied urinary 

concentrations of metabolites from GST-dependent pathway in human volunteers exposed to 50 and 100 

ppm TCE for 15 min or occupationally exposed (0.4 to 21 ppm TWA) workers. They found little or no 

such metabolites and suggested the glutathione-mediated metabolism is of minor importance in humans. 

There were evidences suggesting that TCE is metabolized in the reproductive tract of the mouse and 

monkey; the fact that TCE and its metabolites accumulated in seminal fluid in human diagnosed with 

clinical infertility also suggested associations between production of TCE metabolites, reproductive 

toxicity, and impaired fertility (Forkert et al., 2003). 

In vitro studies 

Extensive in vitro biotransformation studies have been published on a variety of enzyme preparations 

and cell culture systems ranging from cell free tissue preparations (Lipscomb et al., 1997, 1998; 

Lipscomb and Garrett, 1998; Lash et al., 1999b; Cai and Guengerich, 2000; Snawder and Lipscomb, 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

2000; Cummings et al., 2001; Lipscomb et al., 2003a) to highly purified human enzymes (Cai and 

Guengerich, 2001), to primary and other cell cultures (Lash et al., 1999b; Cummings and Lash, 2000; 

Cummings et al., 2000a,b; Walgren et al., 2000; Cummings et al., 2001) including collagen gel 

sandwich cultures of rat hepatocytes (De Smet et al., 2000). 

H

Cl Cl

Cl

SG

Cl Cl

H S

Cl Cl

H NH3+

H

COOS

Cl Cl

H NHCOCH3

H

COOH

CCl3 H

OH

OH

Fe

O

Cl

H C+

Cl

Cl

gluc O

CCl3

OH

Cl2CH

O

OH

Cl3C

O

SG

Cl H

Cl S

Cl H

Cl NH3+

H

COOS

Cl H

Cl NHCOCH3

H

COOH

C C S

Cl

H SH

Cl Cl

H

H

Cl2HC

S

OH CCl3

Thioketene

S-1,2-DCVC

CYP450

CYP450

TCOH TCOG

TCA DCA

CHO

S-1,1-DCVC

GSH

GSH

β-Lyase

β-Lyase

Cytotoxicity

Mutagenicity

Cytotoxicity

Mutagenicity

1,2-DVT

Excretion

in Urine

Excretion

in Urine

1

2

2

1

1

1

3

3

Potential interaction between PCE and TCE

Potential interaction between MC and TCE

Potential interaction between CHL and TCE

1

1

3

3

3

3

Figure 8.1

Figure 8.1 Superimposed on a metabolic 

schematic for TCE (Dobrev, et al., 2001), 

metabolic steps where pharmacokinetic 

interactions could occur are denoted by 

symbols. Identification of additional 

potential interactions should occur during 

CRA Steps 3 and 6.

Cai and Guengerich (2001) demonstrated that the direction reaction of TCE oxide with either human 

P450 2E1, P450 2B1, or NADPH-P450 reductase was shown to lead to enzyme inactivation, and no 

recovery of either enzyme occurred. 

8.2.4 Elimination and Excretion

Presystemic elimination of TCE has been shown by Lee et al. (1996) to be inversely related to dose. 

When relatively high doses were administered to rats via the portal vein, first-pass hepatic extraction 

became negligible. This phenomenon could result not only from metabolic saturation, but from suicidal 

destruction of cytochrome P450 and hepatocellular injury as well (Lee et al., 2000a). Subsequent 

pharmacokinetic analysis by Lee et al. (2000b) indicated that TCE was eliminated by capacity-limited 

hepatic metabolism, no evidence for P450 2E1 destruction, with incursion into nonlinear kinetics with 

bolus doses greater or equal to 8 to 16 mg/kg. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

8.3 Interactions of TCE with other chemicals 

Interaction studies reported in the recent literature covered diverse subject areas. Each of the relevant 

papers is discussed briefly below. 

Dobrev et al. (2001; 2002) studied “Interaction Thresholds” in rats and humans using interactive PBPK 

modeling of a ternary mixture of TCE, tetrachloroethylene, and 1,1,1-trichloroethane. Because of 

competitive inhibition of the primary metabolic system, P450 2E1, an alternative pathway, the GST 

conjugation system, becomes important. It was demonstrated that at or below the current threshold limit 

values (TLVs) for these three chemicals, the coexposure to these chemicals would result in significant 

interactions. 

Very high doses (2000 to 5000 mg/kg, ip) of TCE induced anticonvulsive effect of a number of drugs 

(Shih et al., 2001); it was suggested that this effect might be predominantly mediated by GABA 

receptors. 

A full-factorial design for neurobehavioral evaluations of mixtures of TCE, heptachlor, and di (2-

ethylhexyl) phthalate in F344 rats was carried out by Moser et al. (2003). In general, significant overall 

interactions that deviated from response additivity were detected for most endpoints (11 of 14). Most of 

the interactions are antagonistic in nature. 

Pretreatment of TCE in Sprague-Dawley rats altered drug kinetics of theophylline, quinidine, and 

pentobarbital (Kukongviriyapan et al., 2001). 

Dietary incorporation of guar gum, a thickener and stabilizer in foods and pharmaceuticals, was found to 

decrease TCE accumulation in the body by reducing absorption and fat tissue mass (Nakashima and 

Ikegami, 2001). 

8.4 PBPK models

TCE is undoubtedly one of the chemicals, which were most extensively studied using PBPK modeling 

technique. The initial development of PBPK models was reported by Andersen et al. (1987). This 

initial PBPK model for TCE was followed by a number of variations by others for different goals 

(Fisher et al., 1989; 1990; Koizumi, 1989; Dallas et al., 1991). As the science advances, more and more 

sophistication were incorporated into the later PBPK models. Thus, PBPK models with incorporation of 

TCE metabolites, as well as reproductive physiology and toxicology (Fisher et al., 1989; 1990; 1991; 

Abbas et al., 1996; Abbas and Fisher, 1997; Fisher et al., 1998; Greenberg et al., 1999), and 

pharmacokinetic and pharmacodynamic interactions (Elmasri et al., 1996; Byczkowski et al., 1999) were 

seen in the literature. Furthermore, the application of PBPK modeling in risk assessment received 

progressively more emphasis (Allen and Fisher, 1993; Fisher and Allen, 1993; Gearhart et al., 1993; 

Clewell et al., 1995; Cronin et al., 1995; Bogen and Gold, 1997; Simon, 1997). The 2000 Monograph in 

EHP and more recent PBPK modeling efforts included its application in risk assessment (Clewell et al., 

2000; Fisher, 2000), statistical analyses for variability and uncertainty (Bois, 2000a,b), further 

toxicological interaction studies to define “Interaction Thresholds” (Dobrev et al., 2001; 2002). PBPK 

modeling studies for other specific purposes or toxic endpoints are also seen. Thus, Poet et al. (2000) 

utilized PBPK modeling for assessing percutaneous absorption of TCE in rats and humans. PBPK 

models for the transport of TCE in adipose tissues were reported by Albanese et al. (2002) and PBPK 

modeling for male Long-Evans rats to aid in evaluation of neurotoxicity data was published by Simmons 

et al. (2002). 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

8.5 Risk assessment related 

Because TCE is a very important industrial chemical and a prevalent environmental pollutant, the risk 

assessment, particularly cancer risk assessment became an area of much scientific debate. 

Consequently, quite a number of publications, review articles, and documents are available specifically 

dealing with mechanisms of toxicity of TCE and PBPK modeling in relation to risk assessment, as well 

as the process of risk assessment of TCE (Allen and Fisher, 1993; Fisher and Allen, 1993; Gearhart et 

al., 1993; Clewell et al., 1995; Cronin et al., 1995; Bogen and Gold, 1997; Simon, 1997; Brauch et al., 

1999; Bruning et al., 1999; Green and Lash, 1999; Motohashi et al., 1999a,b; Schraml et al., 1999; 

Barton and Clewell, 2000; Bois, 2000a,b; Bruning and Bolt, 2000; Bull, 2000; Chen, 2000; Clewell et 

al., 2000; Fisher, 2000; Green, 2000; Lash et al., 2000a,b; Moore and Harington-Brock, 2000; Pastino et 

al., 2000; Rhomberg, 2000; Wartenberg et al., 2000; Ruden, 2001a,b; Stewart, 2001; USEPA, 2001; 

Ruden 2001a,b; 2002a,b; Bruning et al., 2003; Lipscomb et al., 2002; 2003b; Ruden, 2003). 

8.6 Literature Cited

Abbas, R., and Fisher, J. W. (1997). A physiologically based pharmacokinetic model for 

trichloroethylene and its metabolites, chloral hydrate, trichloroacetate, dichloroacetate, 

trichloroethanol, and trichloroethanol glucuronide in B6C3F1 mice. Toxicol Appl Pharmacol 147, 

15-30. 

Abbas, R. R., Seckel, C. S., Kidney, J. K., and Fisher, J. W. (1996). Pharmacokinetic analysis of chloral 

hydrate and its metabolism in B6C3F1 mice.[erratum appears in Drug Metab Dispos 1997 

Dec;25(12):1449]. Drug Metab Dispos 24, 1340-6. 

Albanese, R. A., Banks, H. T., Evans, M. V., and Potter, L. K. (2002). Physiologically based 

pharmacokinetic models for the transport of trichloroethylene in adipose tissue. Bull Math Biol 64, 

97-131. 

Allen, B. C., and Fisher, J. W. (1993). Pharmacokinetic modeling of trichloroethylene and 

trichloroacetic acid in humans. Risk Anal 13, 71-86. 

Andersen, M. E., Gargas, M. L., Clewell, H. J., 3rd, and Severyn, K. M. (1987). Quantitative evaluation 

of the metabolic interactions between trichloroethylene and 1,1-dichloroethylene in vivo using gas 

uptake methods. Toxicol Appl Pharmacol 89, 149-57. 

ATSDR. 1997. Toxicological Profile for Trichloroethylene. Agency for Toxic Substances and Disease 

Registry. 335 pp. 

Barton, H. A., and Clewell, H. J., 3rd (2000). Evaluating noncancer effects of trichloroethylene: 

dosimetry, mode of action, and risk assessment. Environ Health Perspect 108 Suppl 2, 323-34. 

Bloemen, L. J., Monster, A. C., Kezic, S., Commandeur, J. N., Veulemans, H., Vermeulen, N. P., and 

Wilmer, J. W. (2001). Study on the cytochrome P-450- and glutathione-dependent biotransformation 

of trichloroethylene in humans. Int Arch Occup Environ Health 74, 102-8. 

Bogen, K. T., and Gold, L. S. (1997). Trichloroethylene cancer risk: simplified calculation of PBPKbased MCLs for cytotoxic end points. Reg Toxicol Pharmacol 25, 26-42. 

Bois, F. Y. (2000a). Statistical analysis of Fisher et al. PBPK model of trichloroethylene kinetics. 

Environ Health Perspect 108 Suppl 2, 275-82. 

Bois, F. Y. (2000b). Statistical analysis of Clewell et al. PBPK model of trichloroethylene kinetics. 

Environ Health Perspect 108 Suppl 2, 307-16. 

Boyer, A. S., Finch, W. T., and Runyan, R. B. (2000). Trichloroethylene inhibits development of 

embryonic heart valve precursors in vitro.[comment]. Toxicol Sci 53, 109-17. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Boyes, W. K., Bushnell, P. J., Crofton, K. M., Evans, M., and Simmons, J. E. (2000). Neurotoxic and 

pharmacokinetic responses to trichloroethylene as a function of exposure scenario. Environ Health 

Perspect 108 Suppl 2, 317-22. 

Brauch, H., Weirich, G., Brieger, J., Glavac, D., Rodl, H., Eichinger, M., Feurer, M., Weidt, E., 

Puranakanitstha, C., Neuhaus, C., Pomer, S., Brenner, W., Schirmacher, P., Storkel, S., Rotter, M., 

Masera, A., Gugeler, N., and Decker, H. J. (2000). VHL alterations in human clear cell renal cell 

carcinoma: association with advanced tumor stage and a novel hot spot mutation. Cancer Res 60, 

1942-8. 

Brauch, H., Weirich, G., Hornauer, M. A., Storkel, S., Wohl, T., and Bruning, T. (1999). 

Trichloroethylene exposure and specific somatic mutations in patients with renal cell carcinoma. J 

Natl Cancer Inst 91, 854-61. 

Bruning, T., and Bolt, H. M. (2000). Renal toxicity and carcinogenicity of trichloroethylene: key results, 

mechanisms, and controversies. Crit Rev Toxicol 30, 253-85. 

Bruning, T., Mann, H., Melzer, H., Sundberg, A. G., and Bolt, H. M. (1999). Pathological excretion 

patterns of urinary proteins in renal cell cancer patients exposed to trichloroethylene. Occup Med 

(Lond) 49, 299-305. 

Bruning, T., Pesch, B., Wiesenhutter, B., Rabstein, S., Lammert, M., Baumuller, A., and Bolt, H. M. 

(2003). Renal cell cancer risk and occupational exposure to trichloroethylene: results of a 

consecutive case-control study in Arnsberg, Germany. Am J Ind Med 43, 274-85. 

Bull, R. J. (2000). Mode of action of liver tumor induction by trichloroethylene and its metabolites, 

trichloroacetate and dichloroacetate. Environ Health Perspect 108 Suppl 2, 241-59. 

Bull, R. J., Orner, G. A., Cheng, R. S., Stillwell, L., Stauber, A. J., Sasser, L. B., Lingohr, M. K., and 

Thrall, B. D. (2002). Contribution of dichloroacetate and trichloroacetate to liver tumor induction in 

mice by trichloroethylene. Toxicol Appl Pharmacol 182, 55-65. 

Byczkowski, J. Z., Channel, S. R., and Miller, C. R. (1999). A biologically based pharmacodynamic 

model for lipid peroxidation stimulated by trichloroethylene in vitro. J Biochem Mol Toxicol 13, 

205-14. 

Cai, H., and Guengerich, F. P. (2000). Acylation of protein lysines by trichloroethylene oxide. Chem Res 

Toxicol 13, 327-35. 

Cai, H., and Guengerich, F. P. (2001). Reaction of trichloroethylene and trichloroethylene oxide with 

cytochrome P450 enzymes: inactivation and sites of modification. Chem Res Toxicol 14, 451-8. 

Chen, C. W. (2000). Biologically based dose-response model for liver tumors induced by 

trichloroethylene. Environ Health Perspect 108 Suppl 2, 335-42. 

Clewell, H. J., 3rd, Gentry, P. R., Covington, T. R., and Gearhart, J. M. (2000). Development of a 

physiologically based pharmacokinetic model of trichloroethylene and its metabolites for use in risk 

assessment. Environ Health Perspect 108 Suppl 2, 283-305. 

Clewell, H. J., Gentry, P. R., Gearhart, J. M., Allen, B. C., and Andersen, M. E. (1995). Considering 

pharmacokinetic and mechanistic information in cancer risk assessments for environmental 

contaminants: examples with vinyl chloride and trichloroethylene. Chemosphere 31, 2561-78. 

Cronin, W. J. t., Oswald, E. J., Shelley, M. L., Fisher, J. W., and Flemming, C. D. (1995). A 

trichloroethylene risk assessment using a Monte Carlo analysis of parameter uncertainty in 

conjunction with physiologically-based pharmacokinetic modeling. Risk Anal 15, 555-65. 

Cummings, B. S., and Lash, L. H. (2000). Metabolism and toxicity of trichloroethylene and S-(1,2-

dichlorovinyl)-L-cysteine in freshly isolated human proximal tubular cells. Toxicol Sci 53, 458-66. 

Cummings, B. S., Parker, J. C., and Lash, L. H. (2000b). Role of cytochrome P450 and glutathione Stransferase alpha in the metabolism and cytotoxicity of trichloroethylene in rat kidney. Biochem 

Pharmacol 59, 531-43. 

Cummings, B. S., Parker, J. C., and Lash, L. H. (2001). Cytochrome p450-dependent metabolism of 

trichloroethylene in rat kidney. Toxicol Sci 60, 11-9. 

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USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Cummings, B. S., Zangar, R. C., Novak, R. F., and Lash, L. H. (2000a). Cytotoxicity of 

trichloroethylene and S-(1, 2-dichlorovinyl)-L-cysteine in primary cultures of rat renal proximal 

tubular and distal tubular cells. Toxicology 150, 83-98. 

Dallas, C. E., Gallo, J. M., Ramanathan, R., Muralidhara, S., and Bruckner, J. V. (1991). Physiological 

pharmacokinetic modeling of inhaled trichloroethylene in rats. Toxicol Appl Pharmacol 110, 303-14. 

De Smet, K., Bruning, T., Blaszkewicz, M., Bolt, H. M., Vercruysse, A., and Rogiers, V. (2000). 

Biotransformation of trichloroethylene in collagen gel sandwich cultures of rat hepatocytes. Arch 

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Dobrev, I. D., Andersen, M. E., and Yang, R. S. H. (2001). Assessing interaction thresholds for 

trichloroethylene in combination with tetrachloroethylene and 1,1,1-trichloroethane using gas uptake 

studies and PBPK modeling. Arch Toxicol 75, 134-44. 

Dobrev, I. D., Andersen, M. E., and Yang, R. S. H. (2002). In silico toxicology: simulating interaction 

thresholds for human exposure to mixtures of trichloroethylene, tetrachloroethylene, and 1,1,1-

trichloroethane. Environ Health Perspect 110, 1031-9. 

El-Masri, H. A., Constan, A. A., Ramsdell, H. S., and Yang, R. S. H. (1996). Physiologically based 

pharmacodynamic modeling of an interaction threshold between trichloroethylene and 1,1-

dichloroethylene in Fischer 344 rats. Toxicol Appl Pharmacol 141, 124-132. 

Fisher, J. W. (2000). Physiologically based pharmacokinetic models for trichloroethylene and its 

oxidative metabolites. Environ Health Perspect 108 Suppl 2, 265-73. 

Fisher, J. W., and Allen, B. C. (1993). Evaluating the risk of liver cancer in humans exposed to 

trichloroethylene using physiological models. Risk Anal 13, 87-95. 

Fisher, J. W., Channel, S. R., Eggers, J. S., Johnson, P. D., MacMahon, K. L., Goodyear, C. D., 

Sudberry, G. L., Warren, D. A., Latendresse, J. R., and Graeter, L. J. (2001). Trichloroethylene, 

trichloroacetic acid, and dichloroacetic acid: do they affect fetal rat heart development? Int J Toxicol

20, 257-67. 

Fisher, J. W., Gargas, M. L., Allen, B. C., and Andersen, M. E. (1991). Physiologically based 

pharmacokinetic modeling with trichloroethylene and its metabolite, trichloroacetic acid, in the rat 

and mouse. Toxicol Appl Pharmacol 109, 183-95. 

Fisher, J. W., Mahle, D., and Abbas, R. (1998). A human physiologically based pharmacokinetic model 

for trichloroethylene and its metabolites, trichloroacetic acid and free trichloroethanol. Toxicol Appl 

Pharmacol 152, 339-59. 

Fisher, J. W., Whittaker, T. A., Taylor, D. H., Clewell, H. J., 3rd, and Andersen, M. E. (1989). 

Physiologically based pharmacokinetic modeling of the pregnant rat: a multiroute exposure model 

for trichloroethylene and its metabolite, trichloroacetic acid. Toxicol Appl Pharmacol 99, 395-414. 

Fisher, J. W., Whittaker, T. A., Taylor, D. H., Clewell, H. J., 3rd, and Andersen, M. E. (1990). 

Physiologically based pharmacokinetic modeling of the lactating rat and nursing pup: a multiroute 

exposure model for trichloroethylene and its metabolite, trichloroacetic acid. Toxicol Appl 

Pharmacol 102, 497-513. 

Forkert, P. G., Lash, L. H., Nadeau, V., Tardif, R., and Simmonds, A. (2002). Metabolism and toxicity 

of trichloroethylene in epididymis and testis. Toxicol Appl Pharmacol 182, 244-254. 

Forkert, P. G., Lash, L., Tardif, R., Tanphaichitr, N., Vandevoort, C., and Moussa, M. (2003). 

Identification of trichloroethylene and its metabolites in human seminal fluid of workers exposed to 

trichloroethylene. Drug Metab Dispos 31, 306-11. 

Gearhart, J. M., Mahle, D. A., Greene, R. J., Seckel, C. S., Flemming, C. D., Fisher, J. W., and Clewell, 

H. J., 3rd (1993). Variability of physiologically based pharmacokinetic (PBPK) model parameters 

and their effects on PBPK model predictions in a risk assessment for perchloroethylene (PCE). 

Toxicol Lett 68, 131-44. 

Green, L. C., and Lash, T. L. (1999). Re: "Renal cell cancer correlated with occupational exposure to 

trichloroethylene". J Cancer Res Clin Oncol 125, 430-2. 

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Green, T. (2000). Pulmonary toxicity and carcinogenicity of trichloroethylene: species differences and 

modes of action. Environ Health Perspect 108 Suppl 2, 261-4. 

Greenberg, M. S., Burton, G. A., and Fisher, J. W. (1999). Physiologically based pharmacokinetic 

modeling of inhaled trichloroethylene and its oxidative metabolites in B6C3F1 mice. Toxicol Appl 

Pharmacol 154, 264-78. 

Griffin, J. M., Blossom, S. J., Jackson, S. K., Gilbert, K. M., and Pumford, N. R. (2000a). 

Trichloroethylene accelerates an autoimmune response by Th1 T cell activation in MRL +/+ mice. 

Immunopharmacology 46, 123-37. 

Griffin, J. M., Gilbert, K. M., Lamps, L. W., and Pumford, N. R. (2000b). CD4(+) T-cell activation and 

induction of autoimmune hepatitis following trichloroethylene treatment in MRL+/+ mice. Toxicol 

Sci 57, 345-52. 

Griffin, J. M., Gilbert, K. M., and Pumford, N. R. (2000c). Inhibition of CYP2E1 reverses CD4+ T-cell 

alterations in trichloroethylene-treated MRL+/+ mice. Toxicol Sci 54, 384-9. 

Johnson, P. D., Goldberg, S. J., Mays, M. Z., and Dawson, B. V. (2003). Threshold of trichloroethylene 

contamination in maternal drinking waters affecting fetal heart development in the rat. Environ 

Health Perspect 111, 289-92. 

Kaneko, T., Saegusa, M., Tasaka, K., and Sato, A. (2000). Immunotoxicity of trichloroethylene: a study 

with MRL-lpr/lpr mice. J Appl Toxicol 20, 471-5. 

Kilburn, K. H. (2002). Is neurotoxicity associated with environmental trichloroethylene (TCE)? Arch 

Environ Health 57, 113-20. 

Koizumi, A. (1989). Potential of physiologically based pharmacokinetics to amalgamate kinetic data of 

trichloroethylene and tetrachloroethylene obtained in rats and man. Br J Ind Med 46, 239-49. 

Kukongviriyapan, V., Simajareuk, S., Kukongviriyapan, U., Cha-on, U., and Airarat, W. (2001). 

Alteration of drug kinetics in rats following exposure to trichloroethylene. Pharmacology 63, 90-4. 

Kumar, P., Prasad, A. K., and Dutta, K. K. (2000). Steroidogenic alterations in testes and sera of rats 

exposed to trichloroethylene (TCE) by inhalation. Hum Exp Toxicol 19, 117-21. 

Kumar, P., Prasad, A. K., Mani, U., Maji, B. K., and Dutta, K. K. (2001). Trichloroethylene induced 

testicular toxicity in rats exposed by inhalation. Hum Exp Toxicol 20, 585-9. 

Lash, L. H., Fisher, J. W., Lipscomb, J. C., and Parker, J. C. (2000b). Metabolism of trichloroethylene. 

Environ Health Perspect 108 Suppl 2, 177-200. 

Lash, L. H., Lipscomb, J. C., Putt, D. A., and Parker, J. C. (1999b). Glutathione conjugation of 

trichloroethylene in human liver and kidney: kinetics and individual variation. Drug Metab Dispos

27, 351-9. 

Lash, L. H., Parker, J. C., and Scott, C. S. (2000a). Modes of action of trichloroethylene for kidney 

tumorigenesis. Environ Health Perspect 108 Suppl 2, 225-40. 

Lash, L. H., Putt, D. A., Brashear, W. T., Abbas, R., Parker, J. C., and Fisher, J. W. (1999a). 

Identification of S-(1,2-dichlorovinyl)glutathione in the blood of human volunteers exposed to 

trichloroethylene. J Toxicol Environ Health A 56, 1-21. 

Lash, L. H., Qian, W., Putt, D. A., Hueni, S. E., Elfarra, A. A., Krause, R. J., and Parker, J. C. (2001). 

Renal and hepatic toxicity of trichloroethylene and its glutathione-derived metabolites in rats and 

mice: sex-, species-, and tissue-dependent differences. J Pharmacol Exp Ther 297, 155-64. 

Lee, K. M., Bruckner, J. V., Muralidhara, S., and Gallo, J. M. (1996). Characterization of presystemic 

elimination of trichloroethylene and its nonlinear kinetics in rats. Toxicol Appl Pharmacol 139, 262-

71. 

Lee, K. M., Muralidhara, S., Schnellmann, R. G., and Bruckner, J. V. (2000a). Contribution of direct 

solvent injury to the dose-dependent kinetics of trichloroethylene: portal vein administration to rats. 

Toxicol Appl Pharmacol 164, 46-54. 

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Lee, K. M., Muralidhara, S., White, C. A., and Bruckner, J. V. (2000b). Mechanisms of the dosedependent kinetics of trichloroethylene: oral bolus dosing of rats. Toxicol Appl Pharmacol 164, 55-

64. 

Lipscomb, J. C., Fisher, J. W., Confer, P. D., and Byczkowski, J. Z. (1998). In vitro to in vivo

extrapolation for trichloroethylene metabolism in humans. Toxicol Appl Pharmacol 152, 376-387. 

Lipscomb, J. C., and Garrett, C. M. (1998). Effect of organ procurement conditions on cytochrome P450 activity in rat liver microsomes. In Vitro Mol Toxicol 11, 265-270. 

Lipscomb, J. C., Garrett, C. M., and Snawder, J. E. (1997). Cytochrome P450-dependent metabolism of 

trichloroethylene: Interindividual differences in humans. Toxicol Appl Pharmacol 142, 311-318. 

Lipscomb, J. C., and Kedderis, G. L. (2002). Incorporating human interindividual biotransformation 

variance in health risk assessment. Sci Total Environ 288, 12-21. 

Lipscomb, J. C., Teuschler, L. K., Swartout, J. C., Striley, C. A. F., and Snawder, J. E. (2003a). 

Variance of microsomal protein and cytochrome P450 2E1 and 3A forms in adult human liver. 

Toxicol Mech Methods 13, 45-51. 

Lipscomb, J. C., Teuschler, L. K., Swartout, J., Popken, D., Cox, T., and Kedderis, G. L. (2003b). The 

impact of cytochrome P450 2E1-dependent metabolic variance on a risk-relevant pharmacokinetic 

outcome in humans. Risk Anal In press. 

Mensing, T., Welge, P., Voss, B., Fels, L. M., Fricke, H. H., Bruning, T., and Wilhelm, M. (2002). 

Renal toxicity after chronic inhalation exposure of rats to trichloroethylene. Toxicol Lett 128, 243-7. 

Moore, M. M., and Harrington-Brock, K. (2000). Mutagenicity of trichloroethylene and its metabolites: 

implications for the risk assessment of trichloroethylene. Environ Health Perspect 108 Suppl 2, 215-

23. 

Moser, V. C., MacPhail, R. C., and Gennings, C. (2003). Neurobehavioral evaluations of mixtures of 

trichloroethylene, heptachlor, and di(2-ethylhexyl)phthlate in a full-factorial design. Toxicology 188, 

125-37. 

Motohashi, N., Nagashima, H., and Molnar, J. (1999a). Trichloroethylene. I. Carcinogenicity of 

trichloroethylene. In Vivo 13, 211-4. 

Motohashi, N., Nagashima, H., and Molnar, J. (1999b). Trichloroethylene. II. Mechanism of 

carcinogenicity of trichloroethylene. In Vivo 13, 215-9. 

Nakashima, Y., and Ikegami, S. (2001). Guar gum reduces trichloroethylene accumulation in the body 

by reducing TCE absorption and fat tissue mass. J Ag Food Chem 49, 3499-505. 

Ohta, M., Saito, T., Saito, K., Kurasaki, M., and Hosokawa, T. (2001). Effect of trichloroethylene on 

spatiotemporal pattern of LTP in mouse hippocampal slices. Int J Neurosci 111, 257-71. 

Pastino, G. M., Yap, W. Y., and Carroquino, M. (2000). Human variability and susceptibility to 

trichloroethylene. Environ Health Perspect 108 Suppl 2, 201-14. 

Poet, T. S., Corley, R. A., Thrall, K. D., Edwards, J. A., Tanojo, H., Weitz, K. K., Hui, X., Maibach, H. 

I., and Wester, R. C. (2000). Assessment of the percutaneous absorption of trichloroethylene in rats 

and humans using MS/MS real-time breath analysis and physiologically based pharmacokinetic 

modeling. Toxicol Sci 56, 61-72. 

Rhomberg, L. R. (2000). Dose-response analyses of the carcinogenic effects of trichloroethylene in 

experimental animals. Environ Health Perspect 108 Suppl 2, 343-58. 

Rodenbeck, S. E., Sanderson, L. M., and Rene, A. (2000). Maternal exposure to trichloroethylene in 

drinking water and birth-weight outcomes. Arch Environ Health 55, 188-94. 

Ruden, C. (2001a). Interpretations of primary carcinogenicity data in 29 trichloroethylene risk 

assessments. Toxicology 169, 209-25. 

Ruden, C. (2001b). The use and evaluation of primary data in 29 trichloroethylene carcinogen risk 

assessments. Reg Toxicol Pharmacol 34, 3-16. 

Ruden, C. (2002a). Scrutinizing three trichloroethylene carcinogenicity classifications in the European 

Union--implications for the risk assessment process. Int J Toxicol 21, 441-50. 

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Ruden, C. (2002b). The use of mechanistic data and the handling of scientific uncertainty in carcinogen 

risk assessments. The trichloroethylene example. Reg Toxicol Pharmacol 35, 80-94. 

Ruden, C. (2003). Science and transscience in carcinogen risk assessment--the European Union 

regulatory process for trichloroethylene. J Toxicol Environ Health B 6, 257-77. 

Schraml, P., Zhaou, M., Richter, J., Bruning, T., Pommer, M., Sauter, G., Mihatsch, M. J., and Moch, H. 

(1999). Untersuchung von Nierentumoren bei Trichlorathylen-exponierten Arbeitern mittels 

Comparative Genomic Hybridization und DNA Sequenzierung. Verh Dtsch Ges Pathol 83, 218-24. 

Shih, C. L., Chen, H. H., and Chiu, T. H. (2001). Acute exposure to trichloroethylene differentially 

alters the susceptibility to chemoconvulsants in mice. Toxicology 162, 35-42. 

Simmons, J. E., Boyes, W. K., Bushnell, P. J., Raymer, J. H., Limsakun, T., McDonald, A., Sey, Y. M., 

and Evans, M. V. (2002). A physiologically based pharmacokinetic model for trichloroethylene in 

the male long-evans rat. Toxicol Sci 69, 3-15. 

Simon, T. W. (1997). Combining physiologically based pharmacokinetic modeling with Monte Carlo 

simulation to derive an acute inhalation guidance value for trichloroethylene. Reg Toxicol 

Pharmacol 26, 257-70. 

Snawder, J. E., and Lipscomb, J. C. (2000). Interindividual variance of cytochrome P450 forms in 

human hepatic microsomes: correlation of individual forms with xenobiotic metabolism and 

implications in risk assessment. Reg Toxicol Pharmacol 32, 200-9. 

Stewart, B. W. (2001). Trichloroethylene and cancer: a carcinogen on trial. Med J Aust 174, 244-7. 

Tao, L., Yang, S., Xie, M., Kramer, P. M., and Pereira, M. A. (2000). Effect of trichloroethylene and its 

metabolites, dichloroacetic acid and trichloroacetic acid, on the methylation and expression of c-Jun 

and c-Myc protooncogenes in mouse liver: prevention by methionine. Toxicol Sci 54, 399-407. 

USEPA. (2001). Trichloroethylene Health Risk Assessment: Synthesis and Characterization. 

EPA/600/P-01/002A, August 2001, USEPA Office or Research and Development, Washington, DC 

20460. 

Walgren, J. E., Kurtz, D. T., and McMillan, J. M. (2000). The effect of the trichloroethylene metabolites 

trichloroacetate and dichloroacetate on peroxisome proliferation and DNA synthesis in cultured 

human hepatocytes. Cell Biol Toxicol 16, 257-73. 

Wartenberg, D., Reyner, D., and Scott, C. S. (2000). Trichloroethylene and cancer: epidemiologic 

evidence.[comment]. Environ Health Perspect 108 Suppl 2, 161-76. 

Waseem, M., Ali, M., Dogra, S., Dutta, K. K., and Kaw, J. L. (2001). Toxicity of trichloroethylene 

following inhalation and drinking contaminated water. J Appl Toxicol 21, 441-4. 

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Tetrachloroethylene

9.0 Introduction

Tetrachloroethylene is made by direct chlorination or oxychlorination of certain hydrocarbons. 

Tetrachloroethylene is used as a chemical intermediate, as solvent for metal cleaning and vapor 

degreasing, and for dry-cleaning and textile processing. (Aggazzotti et al. 1994) It is found in many 

household products, including paint removers, water repellents, silicone lubricants, spot removers, 

adhesives, and wood cleaners (ATSDR 1997). 

9.1 Toxic effects 

Liver, kidney, blood, and the central nervous system are the target organs for systemic effects (Calabrese 

1983; Chen et al. 2002; Echeverria et al. 1995; Ferroni et al. 1992; Umezu et al. 1997; Utzinger and 

Schlatter 1977; Zavon 1967). Exposure to high concentrations of tetrachloroethylene induces dizziness, 

headache, sleepiness, confusion, nausea, unconsciousness, and death. Irritation could occur when skin is 

exposed to tetrachloroethylene. Breathing the vapor may irritate the lungs, causing coughing and/or 

shortness of breath (Stewart et al. 1961). Animal studies showed that tetrachloroethylene can cause 

liver and kidney damage (Schimmelpfennig et al. 1987; Kylin et al. 1963; Kylin et al. 1965; Lash et al.

2002). The developing fetus and children may be particularly susceptible to the toxic effects of 

tetrachloroethylene (Ahlborg 1990; Fredriksson et al. 1993; Motohashi et al. 1993; Spector et al. 1999). 

Exposure to pregnant rodents induces behavioral deficits in pups (Mattsson et al. 1998; Seeber 1989). 

The neurotoxicities of tetrachloroethylene may result from the alterations of fatty acid patterns in the 

brain (ATSDR 1997; Burger et al. 1991). In contrast to the nervous system, the effects on the liver 

including cancer are thought to be a result of the metabolite, trichloroacetic acid (ATSDR 1997). It is 

believed that trichloroacetic acid may play a role in inducing hepatocellular peroxisomes, resulting in 

the production of hydrogen peroxide as a by-product (Bentley et al. 1993). The increased hydrogen 

peroxide may increase DNA damage. Kidney cancer may in part be a result of the formation of the 

genotoxic metabolites from S-(1,2,2-trichlorovinyl) glutathione by β-lyase (Birner et al. 1997; Cooper et 

al. 2002; Green et al. 1990). Tetrachloroethylene is classified as a group 2A carcinogen (probably 

carcinogenic to human) (Aschengrau et al. 1993; Aschengrau et al. 1998; Aschengrau et al. 2003; 

Wartenberg et al. 2000). 

9.2 Pharmacokinetics

Tetrachloroethylene is readily absorbed through oral, skin, and inhalation exposure (Ward et al. 1988). 

Once it is absorbed, tetrachloroethylene is distributed to fatty tissues because of high lipophilicity 

(fat/blood partition coefficient is about 140) (Dallas et al. 1994c). The half-life of tetrachloroethylene in 

fat tissues is 55 hours (ATSDR 1997). One to three percent of absorbed tetrachloroethylene is 

metabolized to trichloroacetic acid in the liver (ACGIH 1991). Unmetabolized tetrachloroethylene is 

exhaled (ATSDR 1997). This is the primary route of excretion. Trichloroacetic acid is excreted in the 

urine (ATSDR 1997). 

9.2.1 Absorption

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Tetrachloroethylene is readily absorbed in the G.I. tract and lungs. Pulmonary uptake is proportional to 

ventilation rate, duration of exposure, and the concentration in the inspired air (ATSDR 1997). In rats, 

the proportion absorbed was approximately 55-70% after 1 minute, gradually declining to 40-50% after 

2 hours (Dallas et al. 1994b). Dermal absorption has been studied in guinea pigs (Bogen et al. 1992). 

9.2.2 Distribution

Tetrachloroethylene is preferentially stored in fat tissues. In rats, distribution to brain, liver, and kidneys 

has also been demonstrated (Frantz and Watanabe 1983; Dallas et al. 1994a; Dallas et al. 1994b). In 

animal studies, transplacental and lactational transport of unchanged tetrachloroethylene has been 

reported (Byczkowski et al. 1994; Hamada and Tanaka 1995). 

9.2.3 Metabolism

The metabolic pathways of tetrachloroethylene are summarized in Figure 9.1. The overlapping 

pathways with the other three volatile organics in Mixture 2 can be seen in Figure 8.1 under 

trichloroethylene. 

Fig 9.1 Metabolic pathways 

of tetrachloroethylene 

(ATSDR, 1997) 

Human pharmacokinetic studies have been performed in volunteers and workers. The pharmacokinetics 

of tetrachloroethylene by inhalation exposure has been described (Ikeda 1977; Monster et al. 1979; 
USEPA Contract No. 3C-R102-NTEX Principal Investigator/Program Director (Last, first, middle): Yang, Raymond S. H. 

Ohtsuki et al. 1983; Imbriani et al. 1988). One study described the pharmacokinetics of 

tetrachloroethylene in a boy who ingested the chemical (Koppel et al. 1985). 

The pharmacokinetics of tetrachloroethylene following inhalation exposure have been described for 

rodents (Pegg et al. 1978; Schumann and Watanabe 1979). The dermal pharmacokinetics of 

tetrachloroethylene in hairless guinea pigs was also studied (Bogen et al. 1992). Pharmacokinetics of 

tetrachloroethylene following oral exposure were reported in several studies including rats, mice, and 

dogs (Frantz and Watanabe 1983; Dallas et al. 1994c). 

In vitro metabolic studies of tetrachloroethylene have been conducted using rat hepatic microsome and 

other subcellular systems (Huang et al. 2001; Costa and Ivanetich 1980; Reitz et al. 1996; Dekant et al.

1998). Some studies focused on the interaction of tetrachloroethylene with rat hepatic microsomal P450 

enzymes (Hanioka et al. 1995a; Hanioka et al. 1995b; Hanioka et al. 1997). 

9.2.4 Excretion

In humans and animals, the major part of the absorbed amount is exhaled unchanged. In humans, 80-

100% of the amount was exhaled as parent compound. In rats, about 70% was exhaled in same 

conditions (ATSDR 1997). Excretion of metabolites in urine is 2% of exposed dose with a half-life of 

75-80 hours (Ikeda et al. 1972; Imbriani et al. 1988; ATSDR 1997). In rats, elimination via maternal 

milk was high (Byczkowski et al. 1994; Byczkowski and Fisher 1995). 

9.3 Interactions with other chemicals

The hepatic monooxygenase system is mainly responsible for oxidation of tetrachloroethylene. Thus, 

chemicals that affect the monooxygenase system could affect the metabolism and toxicity of 

tetrachloroethylene. Two papers were published dealing with pharmacokinetic interactions between 

tetrachloroethylene and other chlorinated contaminants (Dobrev et al. 2001, 2002). Toxicological 

interactions between tetrachloroethylene and ethanol or other chemicals were also reported (Koizumi et 

al. 1982; Dobrov and Poluekto 1971; Kobayashi et al. 1982; Seiji et al. 1989; Giovannini et al. 1992). 

9.4 PBPK models

Several PBPK models for the disposition of tetrachloroethylene were presented in animals and humans 

(Gelman et al. 1996; Haddad et al. 2000; Ward et al. 1988; Koizumi 1989; Bois et al. 1990; Gearhart et 

al. 1993; Dallas et al. 1994b; Dallas et al. 1994c; Wilson and Knaak 1994; Dallas et al. 1995; Reitz et 

al. 1996; Poet et al. 2000; Loizou 2001; Poet et al. 2002). The majority of the available PBPK models 

are concerned with the carcinogenesis of tetrachloroethylene. One model has been developed to predict 

brain concentrations following exposure to tetrachloroethylene during showering (Rao and Brown 

1993). PBPK models for the lactational transfer of tetrachloroethylene through breast milk were 

developed to estimate the risk of tetrachloroethylene exposure to infants (Byczkowski et al. 1994; 

Byczkowski and Fisher 1995). 

9.5 Literature Cited

ACGIH (1991). Documentation of the threshold limit values and biological exposure indices. American 

Conference of Governmental Industrial Hygienists, Cincinnati. 

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Aggazzotti, G., Fantuzzi, G., Righi, E., Predieri, G., Gobba, F. M., Paltrinieri, M., and Cavalleri, A. 

(1994). Occupational and environmental exposure to perchloroethylene (Pce) in dry cleaners and 

their family members. Arch Environ Health 49, 487-493. 

Ahlborg, G. (1990). Pregnancy outcomes among women working in laundries and dry- cleaning shops 

using tetrachloroethylene. Am J Ind Med 17, 567-575. 

Aschengrau, A., Ozonoff, D., Paulu, C., Coogan, P., Vezina, R., Heeren, T., and Zhang, Y. Q. (1993). 

Cancer risk and tetrachloroethylene-contaminated drinking-water in Massachusetts. Arch Environ 

Health 48, 284-292. 

Aschengrau, A., Paulu, C., and Ozonoff, D. (1998). Tetrachloroethylene-contaminated drinking water 

and the risk of breast cancer. Environ Health Perspect 106, 947-953. 

Aschengrau, A., Rogers, S., and Ozonoff, D. (2003). Perchloroethylene-contaminated drinking water 

and the risk of breast cancer: Additional results from Cape Cod, Massachusetts, USA. Environ 

Health Perspect 111, 167-173. 

ATSDR (1997). Toxicological profile for tetrachloroethylene. US Department of Health and Human 

Services. 

Bentley P, Calder I, and C, E. (1993). Hepatic peroxisome proliferation in rodents and its significance 

for humans. Food Chem Toxicol 31, 857-907. 

Birner, G., Bernauer, U., Werner, M., and Dekant, W. (1997). Biotransformation, excretion and 

nephrotoxicity of haloalkene- derived cysteine S-conjugates. Arch Toxicol 72, 1-8. 

Bogen, K. T., Colston, B. W., and Machicao, L. K. (1992). Dermal absorption of dilute aqueous 

chloroform, trichloroethylene, and tetrachloroethylene in hairless guinea- pigs. Fundam Appl Toxicol

18, 30-39. 

Bois, F. Y., Zeise, L., and Tozer, T. N. (1990). Precision and sensitivity of pharmacokinetic models for 

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300-315. 

Burger, A., Gehrisch, S., Jaross, W., Dietz, E., Sucker, M. L., and Gutewort, T. (1991). Investigations on 

the effect of tetrachloroethylene on the lipoprotein metabolism of man by taking into account the 

alcohol uptake. Zeitschrift Fur Klinische Medizin-Zkm 46, 671-674. 

Byczkowski, J. Z., and Fisher, J. W. (1995). A computer-program linking physiologically-based 

pharmacokinetic model with cancer risk assessment for breast- fed infants. Comput Meth Programs 

Biomed 46, 155-163. 

Byczkowski, J. Z., Kinkead, E. R., Leahy, H. F., Randall, G. M., and Fisher, J. W. (1994). Computersimulation of the lactational transfer of tetrachloroethylene in rats using a physiologically-based 

model. Toxicol Appl Pharmacol 125, 228-236. 

Calabrese, E. J. (1983). Tetrachloroethylene in drinking water - a toxicologists perspective - discussion. 

J Am Water Work Assoc 75, 190-190. 

Chen, H. H., Chan, M. H., and Fu, S. H. (2002). Behavioural effects of tetrachloroethylene exposure in 

rats: acute and subchronic studies. Toxicology 170, 201-209. 

Cooper, A. J. L., Bruschi, S. A., and Anders, M. W. (2002). Toxic, halogenated cysteine S-conjugates 

and targeting of mitochondrial enzymes of energy metabolism. Biochem Pharmacol 64, 553-564. 

Costa, A. K., and Ivanetich, K. M. (1980). Tetrachloroethylene metabolism by the hepatic-microsomal 

cytochrome-P-450 system. Biochem Pharmacol 29, 2863-2869. 

Dallas, C. E., Chen, X. M., Muralidhara, S., Varkonyi, P., Tackett, R. L., and Bruckner, J. V. (1994a). 

Use of tissue disposition data from rats and dogs to determine species-differences in input 

parameters for a physiological model for perchloroethylene. Environ Res 67, 54-67. 

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Dobrev, I. D., Andersen, M. E., and Yang, R. S. H. (2002). In silico toxicology: simulating interaction 

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Hanioka, N., Omae, E., Yoda, R., Jinno, H., Nishimura, T., and Ando, M. (1997). Effect of 

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tetrachloroethylene and chloroform - single exposure. Act Pharmacol Toxicol 20, 16-&. 

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tetrachloroethylene using a set of time series analyses. Neurotoxicol Teratol 15, 3-10. 

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metabolize tetrachloroethylene. Int Arch Occup Environ Health 51, 381-390. 

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H. I., and Wester, R. C. (2000). Assessment of the percutaneous absorption of trichloroethylene in 

rats and humans using MS/MS real-time breath analysis and physiologically based pharmacokinetic 

modeling. Toxicol Sci 56, 61-72. 

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Poet, T. S., Weitz, K. K., Gies, R. A., Edwards, J. A., Thrall, K. D., Corley, R. A., Tanojo, H., Hui, X. 

Y., Maibach, H. I., and Wester, R. C. (2002). PBPK modeling of the percutaneous absorption of 

perchloroethylene from a soil matrix in rats and humans. Toxicol Sci 67, 17-31. 

Rao H, and Brown DR (1993). A physiologically based pharmacokinetic assessment of 

tetrachloroethylene in ground water for a bathing and showering determination. Risk Anal 13, 37-49. 

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of perchloroethylene metabolism for physiologically based pharmacokinetic modeling in rats, mice, 

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hepatotoxicity by tetrachloroethylene. Zeitschrift Fur Klinische Medizin-Zkm 42, 2241-2243. 

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metabolism and hepatic macromolecular binding of tetrachloroethylene. Toxicol Appl Pharmacol 48, 

A89-A89. 

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Teratol 11, 579-583. 

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and Ikeda, M. (1989). Dose excretion relationship in tetrachloroethylene-exposed workers and the 

effect of tetrachloroethylene co-exposure on trichloroethylene metabolism. Am J Ind Med 16, 675-

684. 

Spector, J., Lewandowski, A. G., Mott, J. A., and Schreiber, J. S. (1999). Neuropsychological and 

behavioral functioning in tetrachloroethylene-exposed pre-school children and controls. Arch Clin 

Neuropsychol 14, 661-662. 

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tetrachloroethylene vapor - relationship of expired air and blood-concentrations to exposure and 

toxicity. Arch Environ Health 2, 516-&. 

Umezu, T., Yonemoto, J., Soma, Y., and Miura, T. (1997). Behavioral effects of trichloroethylene and 

tetrachloroethylene in mice. Pharmacol Biochem Behav 58, 665-671. 

Utzinger, R., and Schlatter, C. (1977). Review on toxicity of trace amounts of tetrachloroethylene in 

water. Chemosphere 6, 517-524. 

Ward, R. C., Travis, C. C., Hetrick, D. M., Andersen, M. E., and Gargas, M. L. (1988). 

Pharmacokinetics of tetrachloroethylene. Toxicol Appl Pharmacol 93, 108-117. 

Wartenberg, D., Reyner, D., and Scott, C. S. (2000). Trichloroethylene and cancer: Epidemiologic 

evidence. Environ Health Perspect 108, 161-176. 

Wilson, J. D., and Knaak, J. (1994). Health risk assessment of tetrachloroethylene using a 

physiologically-based pharmacokinetic model. Abstr Pap Am Chem Soc 207, 62-ENVR. 

Zavon, M. R. (1967). Liver disease from tetrachloroethylene. JAMA 199, 135-&. 

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1, 1, 1-Trichloroethane 

10.0 Introduction

1, 1, 1-Trichloroethane (TCA; Figure 10.1) is a common organic solvent, often used commercially for 

industrial degreasing as well as dry-cleaning. The 2001 Comprehensive Environmental Response, 

Compensation, and Liability Act (CERCLA) Priority List of Hazardous Substances includes TCA in the 

top 100 hazardous substances based upon its environmental distribution, especially at hazardous waste 

sites (ATSDR, 2001a). Furthermore, TCA ranks 13th in the CERCLA Completed Exposure Pathway; 

therefore, humans are frequently exposed to TCA (ATSDR, 2001b). Because of its ability to induce 

central nervous system depression, TCA has been abused, and thus purposeful human exposure also 

occurs. TCA is considered to be a group III carcinogen due to lack of adequate evidence of 

carcinogenicity in rodents and humans (IARC, 1999). 

Cl H 

C C 

Cl H 

Cl H 

Figure 10.1. Structure of 1, 1, 1-trichloroethane. 

10.1 Toxic effects 

TCA has various systemic effects, most notably central nervous system (CNS) depression, 

hepatotoxicity, and cardiovascular complications. Central nervous system depression is the principal 

CNS effect observed in individuals and animals following exposure to TCA (Hall and Hine, 1966; Stahl

et al., 1969; Jones and Winter, 1983; Bowen and Balster, 1998; Bowen et al., 1998; Bruckner et al., 

2001). Descriptions of plausible mode(s) of action for CNS depression are given in numerous reports 

(Rosengren et al., 1985; Nilsson, 1986b; Nilsson, 1986a; Nilsson, 1987; Fernicola et al., 1991; 

Beckstead et al., 2000; Warren et al., 2000; You and Dallas, 2000; Beckstead et al., 2001; Okuda et al., 

2001; Beckstead et al., 2002; Wiley et al., 2002; Lopreato et al., 2003). 

Various reports cite changes in serum enzyme chemistry, which serve as indicators of hepatotoxicity for 

both humans and animals (Halevy et al., 1980; Hodgson et al., 1989). Another marker for hepatotoxicity 

observed following exposure to TCA is accumulation of fat in the liver (Hall and Hine, 1966; Caplan et 

al., 1976; Hodgson et al., 1989). However, many contradictory studies on both humans and animals 

report failure of serum enzymes to change or extremely mild changes, indicating no apparent 

hepatotoxic effects (Domette and Jones, 1960; Carlson, 1973; Kramer et al., 1978; Kelafant et al., 1994; 

Wang et al., 1996). Although observed hepatic alterations are reversible, they tend to indicate mild 

hepatotoxicity induced by TCA and/or a metabolite (Halevy et al., 1980; Bruckner et al., 2001). Cardiac 

sensitization to epinephrine, resulting in arrhythmia, has been linked with exposure to TCA in both 

humans and animals (Clark and Tinston, 1973; Guberan et al., 1976; Macdougall et al., 1987). 

Additionally, cardiac depression, resulting in decreased blood pressure, is caused by exposure to TCA. 

Toraason and coworkers demonstrated decreases in contractility of cultured cardiac cells occurred in a 

dose-dependent manner following treatment with TCA (Toraason et al., 1990). Some reproductive 

effects have also been reported, ranging from increased mammary adenocarcinomas to decreases in 

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sperm motility, however effects were usually slight and in some cases confounded by exposure to 

chemical mixtures (Rudolph and Swan, 1986; Swan et al., 1989; Yang, 1993; Coleman et al., 1999; 

Lemasters et al., 1999; NTP, 2000; Wang et al., 2002). Other studies found no association between TCA 

exposure and reproductive effects (George et al., 1989; Wrensch et al., 1990a; Wrensch et al., 1990b) 

10.2 Pharmacokinetics

10.2.1 Absorption 

Exposure to TCA primarily occurs through inhalation, and has been described in both humans and 

animals (Morgan et al., 1972a; Morgan et al., 1972b; Monster et al., 1979; Hobara et al., 1982; 

Jakobson et al., 1982; Hobara et al., 1983; Koizumi et al., 1983; Nolan et al., 1984; Dallas et al., 1989; 

Boman et al., 1995). Dermal and gastrointestinal exposures are plausible as TCA is a groundwater 

contaminant, though due to the volatility of TCA, the most common exposure route is inhalation 

(ATSDR, 1995). Alternative routes of TCA exposure have been explored by many researchers, as TCA 

is efficiently and rapidly absorbed via the lung, skin, and gastrointestinal tract of humans and animals 

(Stewart and Dodd, 1964; Riihimaki and Pfaffli, 1978; Mitoma et al., 1985; RTI, 1987; Reitz et al., 

1988; Morgan et al., 1991; Yoshida et al., 1998; Giardino et al., 1999; Kezic et al., 2000; Poet et al., 

2000; Kezic et al., 2001). Steady-state blood levels in rats exposed to 50 or 500 ppm TCA were 

approached at 2 hours following initiation of continuous exposure (Dallas et al., 1989). Reitz and 

colleagues noted achievement of maximal blood levels of TCA at 10-15 minutes following 

administration of a 14.2 mg/kg dose of TCA in water via gavage (Reitz et al., 1988). Following the 

initial phases, absorption rates plateau as steady-state levels are approached in blood and tissues; 

generally, blood levels approach steady-state within a few hours following onset of exposure (Monster

et al., 1979; Nolan et al., 1984). 

10.2.2 Distribution

TCA is widely distributed, with preferential distribution to fatty tissues due to its lipophilic nature, 

regardless of exposure scenario (Takahara, 1986; RTI, 1987; Shimada, 1988; Katagiri et al., 1997; You 

and Dallas, 1998). Detectable levels of TCA are found in the fat, liver, kidney, spleen, blood, lung, 

heart, brain, placenta, and fetus following inhalation exposure (Danielsson et al., 1986; Takahara, 1986; 

Shimada, 1988). In mice exposed for 1 hour to 1,000 ppm TCA, tissue concentrations of TCA 

immediately following exposure resulted in preferential accumulation of TCA (in descending order) in 

the fat, liver, kidney, spleen and blood, followed by lung, heart and brain (Takahara, 1986). Schumann 

and coworkers supported these findings, as they reported significantly higher TCA concentrations in 

fatty tissues than in the liver and kidneys following exposure of mice and rats to either 150 or 1,500 ppm 

TCA for 6 hours (Schumann et al., 1982b). Distribution of TCA is regulated by various factors, 

including tissue blood flow rate, tissue volume and tissue:blood partition coefficients, the latter likely 

being most influential (ATSDR, 1995). 

10.2.3 Metabolism

Metabolism of TCA has been studied extensively (Carlson, 1973; Ivanetich and Van den Honert, 1981; 

Casciola and Ivanetich, 1984; Takano et al., 1985; Takano et al., 1988; Kawai et al., 1991; Durk et al., 

1992; Baker and Ronnenberg, 1993). Regardless of exposure route, TCA is metabolized at low rates 

(<10%), mainly to four metabolites: trichloroethanol, trichloroethanol glucuronide, trichloroacetic acid, 

and carbon dioxide (Monster, 1979; Schumann et al., 1982a; Nolan et al., 1984; Mitoma et al., 1985; 

Reitz et al., 1988; Dallas et al., 1989; Kawai et al., 1991); Figure 10.2). Oxidative metabolism of TCA 

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by the cytochrome P-450 mixed-function oxidase system combined with other metabolic dehydrogenase 

enzymes forms trichloroethanol and trichloroacetic acid; trichloroethanol may be further metabolized 

via conjugation to form a glucuronide derivative. Cytochrome P450 2E1, specifically, is believed to play 

a role in TCA metabolism (Nakajima and Sato, 1979; Guengerich et al., 1991; Kaneko et al., 1994). 

Monster and coworkers found in humans exposed to 70 or 145 ppm of TCA for 4 hours, trichloroethanol 

and trichloroacetic acid excreted in the urine only accounted for 2 and 0.5%, respectively, of absorbed 

TCA (Monster et al., 1979). Another byproduct, acetylene, may also be formed from TCA in mammals 

via reductive dechlorination, though only under hypoxic conditions (Durk et al., 1992). The metabolic 

pathway of TCA is shown in Figure 10.2, as well as in Figure 8.1 where interactive reaction network 

with the other three volatile organics is evident. 

Cl

Cl

Cl

H

H

H

Cl

Cl

Cl

OH

H

H

Cl

Cl

Cl

OH

O

1,1,1 trichloroethane 1,1,1 trichloroethanol 1,1,1 trichloroacetic acid

glucuronide conjugates

CO2

Figure 10.2. Metabolism of 1, 1, 1-Trichloroethane, reproduced from Agency for Toxic Substances 

and Disease Registry’s toxicological profile (ATSDR, 1995). 

10.2.4 Elimination

The primary route of TCA elimination is exhalation of the parent compound, which occurs fairly rapidly 

following exposure due to TCA’s highly volatile nature (Monster, 1979; Monster et al., 1979; Hobara et 

al., 1982; Schumann et al., 1982c; Schumann et al., 1982a; Schumann et al., 1982b; Nolan et al., 1984). 

In humans exposed to 35 or 350 ppm for 6 hours, more than 91% of TCA absorbed was eliminated, 

unchanged, in exhaled air (Nolan et al., 1984). Similarly, in animals given 20 daily doses of TCA by 

gavage in vegetable oil followed by a single 14C-labeled bolus, 85.1 and 92.3% of TCA was excreted as 

parent compound via exhalation, from rats and mice, respectively, (Mitoma et al., 1985). Both 

acetylene and carbon dioxide are excreted in expired air (Durk et al., 1992; ATSDR, 1995). The other 

major metabolites, trichloroethanol, trichloroethanol glucuronide, and trichloroacetic acid are mainly 

eliminated via urinary excretion, though fecal excretion has also been observed (Caperos et al., 1982; 

Mitoma et al., 1985; Ghittori et al., 1987; Imbriani et al., 1988; Kawai et al., 1991). 

10.2.5 Species variations

Because physiologically based pharmacokinetic (PBPK) modeling attempts to extrapolate between 

various species for risk assessment purposes, variations among species can affect model precision and 

accuracy. Most aspects of TCA pharmacokinetics are similar among species, including absorption and 

elimination route. However, quantitative differences in blood:air partition coefficients as well as 

metabolism rates have been noted (Schumann et al., 1982b). Specifically, mice tend to have higher rates 

of TCA metabolism compared to rats and humans. Furthermore, blood:air partition coefficients, which 

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dramatically effect inhalation absorption differ: 2.53, 5.76, and 10.8 for humans, rats, and mice 

respectively (Reitz et al., 1988). 

10.3 PBPK modeling

Attempts to construct PBPK models appropriate for TCA’s disposition have commonly used approaches 

similar to those developed in 1984 by Ramsey and Andersen (Ramsey and Andersen, 1984; Reitz et al., 

1988). Based upon the Ramsey and Andersen model (RAM), a modified model was used to estimate 

metabolic kinetic constants using a closed, recirculated atmosphere representative of those used for gas 

uptake studies (Gargas et al., 1986). This study found that to adequately describe TCA disposition, its 

metabolism required only a first-order pathway, which was abolished when oxidative microsomal 

metabolism is inhibited. Reitz and colleagues utilized a model similar to the RAM to simulate exposure 

to TCA via inhalation, intravenous administration, bolus gavage, and in drinking water, and 

demonstrated the plausibility of using PBPK models in TCA risk assessment, based upon successful 

interspecies extrapolation (Reitz et al., 1988). 

Bogen and Hall used a derivation of the RAM with an additional compartment for skin to assess risk 

associated with TCA in drinking water, and found PBPK modeling predicted nontoxic TCA 

concentrations lower than the existing NOAELs (Bogen and Hall, 1989). Attempts to determine 

metabolic constants via PBPK modeling concluded that, due to low metabolism of TCA, gas uptake 

study techniques were too insensitive to sufficiently form a TCA PBPK model (Gargas and Andersen, 

1989). Absorption and elimination of TCA across time following an inhalation exposure was measured, 

and a PBPK model was built to predict TCA levels in blood and expired air (Dallas et al., 1989). As 

TCA contaminates both water and soil, percutaneous absorption has been modeled in rats and humans, 

including simulations specific for exposure to children (Poet et al., 2000). Notably, combination of 

quantitative structure-property relationships with traditional PBPK modeling has successfully predicted 

inhalation pharmacokinetics for TCA, as well as other volatile organic chemicals (Beliveau et al., 2003). 

Within the context of utilizing biological monitoring to assess exposure, especially in a work 

environment, PBPK models of TCA have been applied. Droz and coworkers first developed a 

population physiological model to investigate variability in biological monitoring, then applied the 

model to assess how alterations in components such as workload, organ function, and body build 

affected the model’s ability to accurately determine TCA exposure (Droz et al., 1989a,b). Comparison 

of various exposure scenarios on alterations in biological monitoring using PBPK modeling has been 

used to determine which biological indices, i.e., parent compound versus metabolite in various 

biological media, are best suited to assess exposure to TCA (Lapare et al., 1995). A linear fourcompartment mass-balance model was used to not only assess uptake and elimination of TCA in human 

subjects at environmentally feasible levels, but also predict exhaled TCA concentrations in another 

human study (Wallace et al., 1997). Analysis of various PBPK models for a series of chemicals, 

including TCA, has allowed analysis of pharmacokinetic model output sensitivity to variability in both 

biochemical and metabolic input parameters (Hetrick et al., 1991). 

Because exposure to a single chemical compound in industrial or environmental exposure settings is 

unlikely, examination of chemical in mixtures is necessary. Koizumi and coworkers performed 

inhalation studies to investigate co-exposure of TCA with perchloroethylene, and found significant 

decreases in formation of TCA metabolites due to co-exposure with perchloroethylene (Koizumi et al., 

1983). Further, Tardif and Charest-Tardif noted decreases in excretion of TCA metabolites following 

co-exposure with m-xylene (Tardif and Charest-Tardif, 1999). Dobrev and associates successfully 

modeled competitive inhibition of trichloroethylene by TCA and tetrachloroethylene, likely due to a 

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shared metabolic pathway with limited enzymatic capacity, specifically the cytochrome P450s (Dobrev

et al., 2001). Further work by Dobrev and associates also used a combination of tetrachloroethylene, 

perchloroethylene, and TCA to assess possible interactions which might change observed toxicity 

(Dobrev et al., 2002). The findings indicated that co-exposure to the three chlorinated hydrocarbons lead 

to a nonlinear increase in toxic conjugative metabolites of tetrachloroethylene (which are associated 

with renal toxicity and/or carcinogenicity), possibly indicating a greater than additive risk associated 

with exposure to the chemical mixture. Although metabolism of TCA is relatively low (<10%), its 

ability to interact with essential metabolic enzymes may confer TCA the ability to inhibit or decrease 

metabolism/detoxification of other chemicals, especially other organic solvents. Alternatively, because 

of comparatively low affinity for the cytochrome P450s, formation of TCA metabolites (especially 

trichloroethanol and trichloroacetic acid) may be reduced due to enzymatic inhibition caused by coexposure with other chemicals. 

10.4 Literature Cited

ATSDR (1995). Toxicological profile for 1,1,1-trichloroethane. Agency for Toxic Substances and 

Disease Registry. U.S. Department of Health and Human Services, Public Health Service, 

Atlanta, GA. 

ATSDR (2001a). 2001 CERCLA priority list of hazardous substances. Agency for Toxic Substances and 

Disease Registry. U.S. Department of Health and Human Services, Public Health Service, 

Atlanta, GA. 

ATSDR (2001b). 2001 Substances most frequently found in completed exposure pathways (CEPs) at 

hazardous waste sites. Agency for Toxic Substances and Disease Registry. U.S. Department of 

Health and Human Services, Public Health Service, Atlanta, GA. 

Baker, M. T., and Ronnenberg, W. C., Jr. (1993). Contrasting effects of 1,1,1-trichloroethane on 

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Acknowledgments 

This work was funded by the USEPA under Contract No. 3C-R102-NTEX, “Method of Analysis to 

Perform a Tissue-Based Cumulative Risk Assessment for Mixtures of Chemicals.” Numerous 

individuals at Colorado State University, Center for Toxicology and Technology participated in the 

production of the work. Jim Dennison, Manupat Lohitnavy, Ornrat Lohitnavy, Yasong Lu, and Damon 

Perez, assisted with the review of organophosphorus pesticides. Amanda Ashley, Sun Ku Lee, and Ken 

Liao reviewed the literature for chlorinated hydrocarbons. Jim Dennison served as Co-Principal 

Investigator and worked throughout the project on every phase. Linda Monum provided administrative 

support and Christina Barinque provided valuable support with the bibliography.