Phthalates are industrial chemicals that are added to plastics to impart flexibility and resilience and are often referred to as plasticizers. Phthalates also are used as solubilizing or stabilizing agents in other applications. There are numerous products that may contain phthalates: adhesives; automotive plastics; detergents; lubricating oils; some medical devices and pharmaceuticals; plastic raincoats; solvents; vinyl tiles and flooring; and personal-care products, such as soap, shampoo, deodorants, lotions, fragrances, hair spray, and nail polish. Phthalates are often used in polyvinyl chloride type plastics, such as plastic packaging film and sheet, garden hoses, inflatable recreational toys, blood product storage bags, intravenous medical tubing, and toys (ATSDR, 2001, 2002). Because they are not chemically bound to the plastics to which they are added, phthalates can be released into the environment during use or disposal of the product. Various phthalate esters have been measured in specific foods, indoor and ambient air, indoor dust, water sources, and sediments (Clark et al., 2003).
People are exposed through ingestion, inhalation, and, to a lesser extent, dermal contact with products that contain phthalates. For the general population, dietary sources have been considered as the major exposure route, followed by inhaling indoor air. Infants may have relatively greater exposures from ingesting indoor dust containing some phthalates (Clark et al., 2003). Human milk can be a source of phthalate exposure for nursing infants (Calafat et al., 2004; Mortensen et al., 2005). The intravenous or parenteral exposure route can be important in patients undergoing medical procedures involving devices or materials containing phthalates. In settings where workers may be exposed to higher air phthalate concentrations than the general population, urinary metabolite and air phthalate concentrations are roughly correlated (Liss et al., 1985; Nielsen et al., 1985; Pan et al., 2006).
Phthalates are metabolized and excreted quickly and do not accumulate in the body (Anderson et al., 2001). Ingested phthalate diesters are initially hydrolyzed in the intestine to the corresponding monoesters, which are then absorbed (Albro et al., 1982; Albro and Lavenhar, 1989). Absorbed monoester metabolites are usually oxidized in the body and, in humans, excreted in urine largely as glucuronide conjugates (Albro et al., 1982; Dirven et al,. 1993). The table shows the phthalate diesters, corresponding monoester metabolites, and other oxidized metabolites included in the National Report on Human Exposure to Environmental Chemicals (CDC, 2013).
Human health effects from phthalates at low environmental doses or at biomonitored levels from low environmental exposures are unknown. Phthalates have low acute animal toxicity. In chronic rodent studies, several of the phthalates produced testicular injury, liver injury, liver cancer, and teratogenicity, but these effects either have not been demonstrated when tested in non-human primates or are yet to be studied. In vitro studies showed that certain phthalates can bind to estrogen receptors and may have weak estrogenic or anti-estrogenic activity (Coldham et al., 1997; Harris et al., 1997; Jobling et al., 1995), but in vivo studies did not support phthalates having estrogenic effects (Milligan et al., 1998; Okubo et al., 2003; Parks et al., 2000; Zacharewski et al., 1998); however, not all phthalates and metabolites have been tested. In animals, phthalates produced anti-androgenic effects by reducing testosterone production and, at very high levels, reducing estrogen production, effects that may be mediated by inhibiting testicular and ovarian steroidogenesis. High doses of di-2-ethylhexyl phthalate (DEHP), dibutyl phthalate (DBP), and benzylbutyl phthalate (BzBP) during the fetal period produced lowered testosterone levels, testicular atrophy, and Sertoli cell abnormalities in the male animals and, at higher doses, ovarian abnormalities in the female animals (Jarfelt et al., 2005; Lovekamp-Swan and Davis, 2003; McKee et al., 2004; NTP-CERHR, 2003a, 2003b, 2006). Phthalate urinary metabolite levels in men evaluated at an infertility clinic were associated with several measures of sperm function and morphology (Duty et al., 2004; Hauser et al., 2007), but similar findings were not present in young Swedish men with comparable or higher median levels of urinary metabolites (Jonsson et al., 2005).
The monoester metabolites are thought to mediate toxic effects for some of the phthalates, but there are known species-related differences in the hydrolysis of diester phthalates, efficiency of intestinal absorption, and extent of metabolite conjugation to glucuronide (Albro et al., 1982; Kessler et al., 2004; Rhodes et al., 1986). These differences may contribute to species-specific differences in toxicity (ATSDR, 2001, 2002). Also, phthalates have been shown to induce peroxisomal proliferation in rodents, which may be a pathway to the development of liver toxicity and cancers in these animals. However, peroxisomal proliferation may not be a relevant pathway in humans (Rusyn et al., 2006).
The National Toxicology Program's Office of Health Assessment and Translation, formerly Center for the Evaluation of Risks to Human Reproduction (NTP-CERHR) has reviewed the developmental and reproductive effects of specific phthalates (https://www.niehs.nih.gov/research/atniehs/dntp/ohat/index.cfm). Information about external exposure (i.e., environmental levels) and health effects is also available for some phthalates from ATSDR at https://www.atsdr.cdc.gov/toxprofiles/index.asp.
|Phthalates and Urinary Metabolites Measured in the National Biomonitoring Program|
|Phthalate name (CAS number)||Abbreviation||Urinary metabolite (CAS number)||Abbreviation|
|Benzylbutyl phthalate (85-68-7)||BzBP||
Mono-benzyl phthalate (2528-16-7)
(some mono-n-butyl phthalate)
|Dibutyl phthalates (84-74-2)||DBP||
Mono-n-butyl phthalate (131-70-4)
|Dicyclohexyl phthalate (84-61-7)||DCHP||Mono-cyclohexyl phthalate (7517-36-4)||MCHP|
|Diethyl phthalate (84-66-2)||DEP||Mono-ethyl phthalate (2306-33-4)||MEP|
|Di-2-ethylhexyl phthalate (117-81-7)||DEHP||
Mono-2-ethylhexyl phthalate (4376-20-9)
Mono-(2-ethyl-5-carboxypentyl) phthalate (40809-41-4)
|Di-isononyl phthalate (28553-12-0)||DiNP||Mono-isononyl phthalate||MiNP|
|Di-isodecyl phthalate||DiDP||Mono-(carboxynonyl) phthalate||MCNP|
|Dimethyl phthalate (131-11-3)||DMP||Mono-methyl phthalate (4376-18-5)||MMP|
|Di-n-octyl phthalate (117-84-0)||DOP||
Mono-n-octyl phthalate (5393-19-1)
Urinary levels of phthalate metabolites reflect recent exposure to the parent phthalate diester. The proportions of each metabolite for a given phthalate may vary by differing routes of exposure (Liss et al., 1985; Peck and Albro, 1982). Variation occurs from person to person in the proportions or amounts of a metabolite excreted after similar doses (Anderson et al., 2001); variation also occurs in the same person during repetitive monitoring (Fromme et al., 2007; Hauser et al., 2004; Hoppin et al., 2002). Population estimates of concentrations of specific phthalate metabolites may differ by age, gender, and race/ethnicity (Silva et al., 2004).
Finding a measurable amount of one or more phthalate metabolites in urine does not imply that the levels of the metabolites or the parent phthalate cause an adverse health effect. Biomonitoring studies on levels of phthalate metabolites provide physicians and public health officials with reference values so that they can determine whether people have been exposed to higher levels of phthalates than are found in the general population. Biomonitoring data can also help scientists plan and conduct research on exposure and health effects.
CAS No. 96-12-8
1,2-Dibromo-3-chloropropane (DBCP) is a liquid soil fumigant used until 1985 when the U.S. Environmental Protection Agency (EPA) banned applications (ATSDR, 1992). DBCP volatilizes from soil into the air after application. Recent surveys of U.S. public drinking water supplies have not detected DBCP (USGS, 2006).
Exposure to the general population is rare. In the past, inhalational and dermal exposure occurred primarily in formulators and applicators. DBCP can be absorbed by ingestion, inhalation, and dermal routes. After absorption, DBCP is shown in animal studies to distribute widely into most tissues. Metabolites are excreted in urine, feces, and, to a limited extent, exhaled air (ATSDR, 1992; MacFarland et al., 1984).
In animal studies, large acute doses of DBCP produce lethargy, ataxia, and convulsions. High chronic doses in laboratory animals demonstrate kidney toxicity, testicular injury and reduced sperm production, and altered estrus cycles and infertility (ATSDR, 1992; Lag et al., 1989; Rao et al., 1982). Male workers exposed during DBCP production have demonstrated oligospermia or azoospermia; sperm count recovery occurred generally with less than 3 years of workplace exposure (Potashnik, 1983; Potashnik and Yani-Inbar, 1987; Whorton et al., 1979; Lipschultz et al, 1980). In general populations, epidemiologic investigations found no association between exposure to previously contaminated drinking water and birth rates, birth outcomes, gastric cancer, or leukemia (Whorton et al., 1989; Wong et al., 1988).
An increased risk for certain cancers was found in several studies of workers exposed to DBCP (Olsen et al., 1995; Wesseling et al., 1996); however, these studies may have been confounded by other unmeasured exposures. Rodents that were administered DBCP developed tumors in the nasal cavity, lungs, and forestomach (NCI, 1978; NTP, 1982). The International Agency for Research on Cancer classified DBCP as a possible human carcinogen; the National Toxicology Program determined that DBCP was reasonably anticipated to be a human carcinogen. EPA established drinking water and other environmental standards and the Occupational Safety and Health Administration established workplace standards for DBCP. Information about external exposure (i.e., environmental levels) and health effects is available from ATSDR at https://www.atsdr.cdc.gov/toxprofiles/index.asp.
Levels of DBCP in blood reflect recent exposure. DBCP was not detected in the NHANES 2003-2006 subsamples, similar to other studies (Ashley et al. 1994; Churchill et al. 2001). Finding a measurable amount of DBCP in the blood does not imply that the level of DBCP causes an adverse health effect. Biomonitoring studies of DBCP in the blood can provide physicians and public health officials with reference values so that they can determine whether people have been exposed to higher levels of DBCP than are found in the general population. Biomonitoring data can also help scientists plan and conduct research on exposure and health effects.
Agency for Toxic Substances and Disease Registry (ATSDR). Toxicological profile for 1,2-dibromo-3-chloropropane. 1992 [online]. Available at URL: https://www.atsdr.cdc.gov/toxprofiles/tp.asp?id=852&tid=166 8/3/12
Ashley DL, Bonin MA, Cardinali FL, McCraw JM, Wooten JV. Blood concentrations of volatile organic compounds in a nonoccupationally exposed US population and in groups with suspected exposure. Clin Chem 1994;40(7 Pt 2):1401-1404.
Churchill JE, Ashley DL, Kaye WE. Recent chemical exposures and blood volatile organic compound levels in a large population-based sample. Arch Environ Health 2001;56(2):157-166.
Lag M, Omichinski JG, Soderlund EJ, Brunborg G, Holme JA, Dahl JE, et al. Role of P-450 activity and glutathione levels in 1,2-dibromo-3-chloropropane tissue distribution, renal necrosis and in vivo DNA damage. Toxicology 1989;56:273-288.
Lipshultz LI, Ross CE, Ehorton D, Milby T, Samith R, Joyner RE. Dibromochloropropane and its effect on testicular function in man. J Urol 1980;124:464–468.
MacFarland RT, Gandolfi AT, Sipes IC. Extra-hepatic GSH-dependent metabolism of 1,2-dibromoethane (DBE) and 1,2-dibromo-3-chloropropane (DBCP) in the rat and mouse. Drug Chem Toxicol 1984;7:213-227.
National Cancer Institute (NCI). Bioassay of dibromochloropropane for possible carcinogenicity (CAS no. 96-12-8). Technical Report Series No. 28. 1978 [online]. Available at URL: https://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr028.pdf. 8/3/12
National Toxicology Program (NTP). Carcinogenesis bioassay of 1,2-dibromo-3-chloro-propane (CAS no. 96-12-8) inF344/N rats and B6C31F mice (inhalation studies). Technical Report Series No. 206. 1982 [online]. Available at URL: https://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr206.pdf. 8/3/12
Olsen GW, Bodner KM, Stafford BA, Cartmill JB, Gondek MR. Update of the mortality experience of employees with occupational exposure to 1,2-dibromo-3-chloropropane (DBCP). Am J Ind Med 1995;28(3):399-410.
Potashnik G. A four-year reassessment of workers with dibromochloropropane-induced testicular dysfunction. Andrologia 1983;15(2):164-170.
Potashnik G, Yanai-Inbar I. Dibromochloropropane (DBCP): an 8-year reevaluation of testicular function and reproductive performance. Fertil Steril 1987;47(2):317-323.
Rao, K.S., J.D. Burek, F. Murray, John JA, Schwetz BA, Beyer JE, Parker CM. Toxicologic and reproductive effects of inhaled 1,2-dibromo-3-chloropropane in male rabbits. Fund Appl Toxicol 1982;2(5): 241-251.
United States Geological Survey (USGS). Volatile Organic Compounds in the Nation's Ground Water and Drinking-Water Supply Wells. 2006 [online]. Available at URL: https://pubs.usgs.gov/circ/circ1292/. 8/3/12
Wesseling C, Ahlbom A, Antich D, Rodriguez AC, Castro R. Cancer in banana plantation workers in Costa Rica. Int J Epidemiol 1996;25(6):1125-1131.
Whorton D, Milby TH, Krauss RM, Stubbs HA. Testicular function in DBCP exposed pesticide workers. J Occup Med 1979;21(3):161-166.
Whorton DM, Wong O, Morgan RW, Gordon N. An epidemiologic investigation of birth outcomes in relation to dibromochloropropane contamination in drinking water in Fresno County, California, USA. Int Arch Occup Environ Health 1989;61:403-407.
Wong O, Whorton MD, Gordon N, Morgan RW. An epidemiologic investigation of the relationship between DBCP contamination in drinking water and birth rates in Fresno County, California. Am J Public Health 1988;78:43-46.