Estimating exposure to estrogens via drinking water requires information about the concentrations of estrogens in drinking water and the amount of water consumed by a typical person in the United States. The U.S. Environmental Protection Agency (EPA) recommended water ingestion rates of 0.87 L/day for children and 1.4 L/day for adults were used to estimate drinking water consumption (U.S. EPA 1997
). Predicted concentrations in drinking water were used instead of measured concentrations because few studies have measured estrogen concentrations in U.S. drinking water, and those that are available report primarily nondetected concentrations [see Supplemental Material, available online (doi:10.1289/ehp.0900654.S1 via http://dx.doi.org/
); see also Hannah et al. 2009
]. The predicted environmental concentrations (PECs) of synthetic estrogens and endogenous estrogens in drinking water resulting from human use and excretion were estimated using the Ph
ATE (Pharmaceutical Assessment and Transport Evaluation) model, version 2.1.1 (Anderson et al. 2004
ATE requires several compound-specific inputs, including the per capita use or excretion rate, metabolism, POTW removal rate, in-stream removal, and drinking water treatment removal. Removal rates were not available for all estrogens. When no removal information was available for a particular estrogen and a particular removal mechanism, we assumed that no removal occurred to assure that drinking water PECs were not underestimated. A summary of Ph
ATE inputs is provided in the Supplemental Material, Tables SM-4 through SM-9 (doi:10.1289/ehp.0900654.S1).
PhATE generates PECs for approximately 2,710 stream segments in 11 U.S. watersheds and for 184 drinking water treatment plants serving approximately 10,800,000 people located on the modeled stream segments. The PECs used in this analysis are those for each of the drinking water treatment systems included in PhATE. PhATE is able to generate PECs for both mean flow conditions and 7Q10 low-flow (i.e., the lowest consecutive 7-day low flow that occurs on average once every 10 years) conditions. To be conservative, in this analysis we used 7Q10 low-flow PECs because this is when estrogen concentrations in surface water used as source water for a drinking water treatment plant would be at their highest (as opposed to annual average flow when concentrations would be lower).
PECs generated by PhATE are grouped five ways to enable discrimination among different sources of estrogens in drinking water and among types of estrogens. Potential exposures are presented for prescribed synthetic estrogen alone (i.e., EE2), prescribed endogenous estrogens (i.e., E1, E2, and E3 prescribed for therapeutic use), total prescribed estrogens (i.e., E1, E2, E3, and EE2 prescribed for therapeutic use), naturally occurring endogenous estrogens (i.e., naturally occurring animal-derived E1, E2, and E3), and total estrogens (E1, E2, E3, and EE2 from all sources). For dietary and ADI comparisons that required combining exposure to individual estrogens, total estrogen exposure was expressed as estradiol equivalents (E2-eq), which were estimated based on alpha receptor binding for E1, E2, and E3 (assumed to have relative biological activity of 0.1, 1.0, and 0.038, respectively). To avoid potentially underestimating the estrogenic activity of synthetic estrogen, EE2 was assumed to have 10 times the activity of E2 (i.e., a relative potency of 10). More detailed discussion of the relative biological activity adjustments is presented in Supplemental Material (doi:10.1289/ehp.0900654.S1).
Estrogens have specific direct and indirect effects that may be functions of age as well as sex (Aksglaede et al. 2006
; Andersson and Skakkebaek 1999
; Dey et al. 2000
). Some potentially sensitive subpopulations include prepubescent males who have low natural serum estrogen levels and an expected low metabolic clearance rate (Andersson and Skakkebaek 1999
; Klein et al. 1996
), postmenopausal women with naturally low estrogen levels, and people with specific dietary preferences, such as vegetarians, that increase their exposure to phytoestrogens. We examined the differences in total intake of estrogens among several different age groups, between males and females, and among dietary preferences. Women taking prescribed estrogens for therapeutic use had the largest total daily exposure to estrogens, followed by infants fed soy-based formula, then by infants on breast milk or milk-based formula, and last by children and adults (data not shown).
Given the concern about exposure of prepubescent males (an assumed sensitive subpopulation) to estrogens in drinking water, this analysis used estrogen exposure from milk consumption of young children as one set of dietary benchmarks. The second set of dietary benchmarks is the daily exposure of an adult female eating an omnivorous diet. This is likely to be representative of the daily estrogen exposure for the majority of adults in the United States.
We estimated dietary exposure to endogenous estrogens by combining the reported concentrations of estrogens in foodstuffs with the average consumption rate of the foodstuff [see Supplemental Material, Tables SM-1 through SM-3 (doi:10.1289/ehp.0900654.S1)]. Endogenous estrogen intake is likely biased low because concentration data for one or more estrogens in a particular food may be absent and data for other foodstuffs expected to contain estrogens have not been published. Additionally, this analysis does not include the contribution of phytoestrogens to the background estrogenic activity present in a typical U.S. diet. Given that adult premenopausal women are reported to have a total isoflavone intake of 1.78 mg/day (Horn-Ross et al. 2002
) and 2.17 mg/day (Huang et al. 2000
) and a total lignan intake of 0.108 mg/day (Horn-Ross et al. 2002
), 0.525 mg/day (McCann et al. 2003
), and 0.645 mg/day (deKleijn et al. 2002
), omission of phytoestrogens understates background dietary estrogen exposure. Although biologically active (Masutomi et al. 2004
; Montani et al. 2008
; Rimoldi et al. 2007
), phytoestrogens were excluded because data on their concentrations in many foods are not available and their estrogenic potency relative to the animal-derived endogenous estrogens remains difficult to quantify.
In addition to comparing estrogen exposure via drinking water with background dietary exposure, this analysis compares exposure to estrogens in drinking water with four independently derived ADIs (or sets of ADIs) available in the literature.
The WHO (2000)
derived an ADI of 0.05 μg E2/kg body weight (BW)/day based on a no observed effect level (NOEL) of 0.3 mg E2/person/day associated with changes in several hormone-dependent parameters. The WHO divided the NOEL by an uncertainty factor (UF) of 10 to account for “normal variation” among individuals and a second UF of 10 to account for “sensitive populations.” In the United States, the assumed adult body weight is 60 kg, so the adjusted WHO whole-body ADI is 3 μg E2/person/day.
The TTC is an acceptable daily exposure presented by Kroes et al. (2004)
based on a review of toxicity data from a variety of chemicals. In the present study, we conservatively assumed that estrogens are structurally active compounds and evaluated them using the TTC of 0.15 μg/person/day derived by Kroes et al. (2004)
for compounds with a structural alert. Such compounds were assigned the lowest TTC. Had a higher (i.e., less conservative) TTC been used, the MOS estimated in this analysis would have been higher.
Because EE2 is produced in a manufacturing environment, its makers developed occupational exposure limits for the protection of workers using all available toxicologic and pharmacologic data to protect against potential hazards, primarily from dust inhalation. The myriad biologic responses to estrogen exposure argue for use of such an integrated assessment of effect in risk assessment. The occupational exposure limit is established to protect humans against all biologically significant effects and is based on in vivo studies and human experience. The EE2 occupational exposure limit of 0.01 μg/m3 used in our analysis is the most recent and lowest of five occupational exposure limits developed by different manufacturers (Johnson & Johnson, unpublished data). It was converted to an allowable dose of 0.07 μg EE2/person/day by adjusting from an allowable air concentration to an ADI by multiplying by an assumed inhalation rate of 10 m3/person/day and multiplying by 5/7 to account for the difference in number of days a worker is exposed per week versus a member of the general public. An additional 10-fold reduction to account for sensitive populations, in this case potential effects on the developing infant, results in an ADI of 0.007 μg EE2/person/day. Similarly, ADIs of 0.07, 0.02, and 0.07 μg/person/day for E1, E2, and E3, respectively, were derived from their respective occupational exposure limits [0.1 μg/m3, 0.029 μg/m3, and 0.1 μg/m3 (Caldwell DJ, personal communication; Johnson & Johnson, unpublished data)] using the same approach as presented for EE2.
Australia developed water reuse guidelines for estrogens (EPHC et al. 2008
). Australia used the WHO ADI to develop the E2 guideline, however; the ADIs for E1, E3, and EE2 were derived by applying a 10,000-fold safety factor to the lowest therapeutic dose, including a safety factor of 10 to account for sensitive populations. The resulting ADIs for E1, E2, E3, and EE2 are 0.052, 3, 0.084, and 0.0026 μg/person/day, respectively.