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 (http://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 http://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. 117-81-7
Di-2-ethylhexyl phthalate (DEHP) is primarily used to produce flexibility in plastics, mainly polyvinyl chloride, which is used for many consumer products, toys, packaging film, and blood product storage and intravenous delivery systems. Concentrations in plastic materials may reach 40% by weight. DEHP has been removed from or replaced in most toys and food packaging in the United States.
Following ingestion, DEHP is metabolized to more than 30 metabolites which are rapidly eliminated in urine, and in humans, as glucuronide conjugates (Albro et al., 1982; Albro and Lavenhar, 1989; ATSDR, 2002; Peck and Albro, 1982). Four metabolites have been measured in the National Report on Human Exposure to Environmental Chemicals: mono-(2-ethyl-5-hexyl) phthalate (MEHP), mono-(2-ethyl-5-oxohexyl) phthalate (MEOHP), mono-(2-ethyl-5-hydroxyhexyl) phthalate (MEHHP) and mono-(2-ethyl-5-carboxypentyl) phthalate (MECPP).
MEHP is primarily formed by the hydrolysis of DEHP in the gastrointestinal tract and then absorbed. DEHP present in medical devices and parenteral delivery systems results in the diester rather than the monoester form being directly introduced into the blood. After parenteral administration, hydrolysis of DEHP most likely also occurs in the blood, and subsequent metabolism is similar to that following ingestion (Koch et al., 2005a, 2005b, 2005c). MEOHP, MEHHP, and MECPP are produced by the oxidative metabolism of MEHP and are present at roughly three- to five-fold higher concentrations than MEHP in urine (Barr et al., 2003; Fromme et al., 2007; Koch et al., 2003).
MEHP is the putative toxic metabolite of DEHP. Liver toxicity, decreased testicular weight, and testicular atrophy have been observed in rodents fed high doses over a short term or with chronic dosing (McKee et al., 2004; NTP-CERHR, 2006). In contrast, marmoset monkeys fed high dose DEHP for longer than a year did not demonstrate testicular or liver toxicity (NTP-CERHR, 2006). Very high doses of DEHP have suppressed estradiol production in female rats (Lovecamp-Swan and Davis, 2003). The Food and Drug Administration determined that in adults, the amounts of DEHP or MEHP received from intravenous delivery systems or blood transfusions (DEHP is hydrolyzed to MEHP in stored blood) would result in short-term elevations similar to background levels (FDA, 2001). However, critically ill neonates and infants receiving selected or multiple intensive procedures, such as exchange transfusions, extracorporeal membrane oxygenation, and parenteral nutrition, could receive higher exposures than the general population (Calafat et al., 2004; FDA, 2001; Loff et al., 2000; Weuve et al., 2006).
OSHA has established a workplace air standard for external exposure to DEHP; NIOSH and ACGIH have established guidelines for workplace air exposure to DEHP. IARC considers DEHP to be unclassifiable with respect to human carcinogenicity. NTP determined that DEHP is reasonably anticipated to be a human carcinogen.
The levels of MEHP reported in NHANES 1999-2000, 2001-2002, and 2003-2004 appear roughly comparable to those reported previously in several small U.S. studies involving adults (Blount et al., 2000), pregnant women in New York City (Adibi et al., 2003), and low income African-American women in Washington, DC (Hoppin et al., 2002). In contrast, a sample of South Korean women had higher urine MEHP levels: the geometric mean was about ten times higher than for females in each of the NHANES survey periods (CDC, 2012; Koo and Lee, 2005). Median urine MEHP levels in a small group of Japanese adults, in a group of Swedish male military recruits, and in samples of men attending an infertility clinic were similar to median values for adults and males, respectively, in NHANES 1999-2000 and 2001-2002 subsamples (Duty et al., 2004, 2005; Itoh et al., 2005).
In another sample of men attending an infertility clinic, the median and 95th percentile values of urinary MEHP were similar, but MEHHP and MEOHP were about three to five times higher than comparable values found in males in two NHANES survey periods (1999-2000, 2001-2002) (CDC, 2012; Hauser et al., 2007). Compared with the U.S. population in the National Report on Human Exposure to Environmental Chemicals (the Report), urinary MEHP, MEOHP, and MEHHP levels were similar or up to twofold higher in a sample of German residents (Koch et al., 2003; Preuss et al., 2005) and German children (Becker et al., 2004; Koch et al., 2004). During 2001-2003, median levels of urinary MEOHP and MEHHP appeared to be similar in samples of German university students and the adults in the Report (CDC, 2013; Wittasek et al., 2007).
In separate analyses of NHANES 1999-2000 and NHANES 2001-2002, the adjusted geometric mean levels of urinary MEHP were significantly higher in children compared with adolescents and adults, and in females compared with males (Silva et al., 2004). South Korean children had geometric mean urine MEHP levels that were about three times higher than the U.S. children in the Report (CDC, 2013; Koo and Lee, 2005). Younger children eliminate higher proportions of urinary MEHHP and MEOHP relative to MEHP, with the difference increasing as age decreases; this may be the result of differences in metabolism and/or excretion (NTP-CERHR, 2006). Studies of hospitalized neonates have reported urinary geometric mean levels of MEHP, MEOHP, and MEHHP that were two to five times higher, or more (depending on the intensity of DEHP-product exposure), than the geometric means of children in the NHANES subsamples for all three survey periods (Calafat et al., 2004; Weuve et al., 2006). Small studies of plasma and platelet donors have reported very high levels of MEHP, MEOHP, MEHHP and MECPP in urine collected shortly after these procedures (Koch et al., 2005b, 2005c).
Finding a measurable amount of one or more DEHP metabolites in urine does not imply that the levels of the metabolites or the parent compound cause an adverse health effect. Biomonitoring studies on levels of urinary DEHP metabolites provide physicians and public health officials with reference values so that they can determine whether people have been exposed to higher levels of DEHP than are found in the general population. Biomonitoring data can also help scientists plan and conduct research on exposure and health effects.
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