Environ. Sci. Technol. 2006, 40, 2819-2825
Estrogenic Activity of Polychlorinated Biphenyls Present in Human Tissue and the Environment B . R E Y D E C A S T R O , * ,† S U S A N A . K O R R I C K , ‡,§ JOHN D. SPENGLER,‡ AND ANA M. SOTO⊥ Johns Hopkins Bloomberg School of Public Health, Department of Environmental Health Sciences, 615 North Wolfe Street, Room W6001B, Baltimore, Maryland 21205, Harvard School of Public Health, Department of Environmental Health, 677 Huntington Avenue, Boston, Massachusetts 02115, Channing Laboratory, Department of Medicine, Brigham and Women’s Hospital, Harvard Medical School, Boston, Massachusetts, and Tufts University School of Medicine, Department of Anatomy and Cellular Biology, 136 Harrison Avenue, Boston, Massachusetts 02111
This study evaluated the estrogenicity of polychlorinated biphenyls (PCBs) present in environmental media and human tissue and assessed exposure pathways for PCB-derived estrogenic potency in air, soil, and dust from New Bedford, MA, an area with a PCB-contaminated Superfund site. Thirtyfour PCB congeners were assayed for estrogenic potency using E-SCREEN, an assay based on the estrogen-dependent proliferation of MCF-7 cells in vitro. Childhood exposure to estradiol-equivalents via PCBs in environmental media was estimated by weighting previously reported New Bedford congener-specific concentrations by their relative estrogenic potency and published inhalation and soil ingestion rates. Thirteen congeners were weakly estrogenic in E-SCREEN: PCBs 17, 18, 30, 44, 49, 66, 74, 82, 99,103, 110, 128, and 179. These PCBs were typically 6 orders of magnitude less potent than 17β-estradiol, with proliferative potencies ranging from 0.0007% to 0.0040%. Of the environmental media assessed, air (inhalation) had the highest PCB-derived estradiol-equivalent exposure. PCB estrogenic potency information from this study provides an important resource both for preliminary estimation of routes of human exposure to xenoestrogens and for application to human health studies focused on estrogenresponsive health outcomes, such as reproductive development and related malignancies.
Introduction A wide variety of environmental pollutants are hormonally active and may exert adverse health effects by disrupting normal endocrine function (1). Among the health effects thought to be associated with hormonally active agents are impaired reproductive capability and behavior, altered fetal and child development, and the promotion of carcinogenesis. Many xenobiotic compounds have intrinsic hormonal activity * Corresponding author phone: (410)614-3893; fax: (410)955-9334; e-mail:
[email protected]. † Johns Hopkins Bloomberg School of Public Health. ‡ Harvard School of Public Health. § Harvard Medical School. ⊥ Tufts University School of Medicine. 10.1021/es051667u CCC: $33.50 Published on Web 03/14/2006
2006 American Chemical Society
(2), both agonist and antagonist (3, 4). Some environmental pollutants not only have affinity for hormone receptors but are also able to act via the receptor to induce effects in in vitro studies (5). Exposure to environmentally relevant lowdose xenoestrogens can affect the development of the genital tract (6-9) and mammary gland (10, 11), and these effects occur at doses several orders of magnitude lower than those required to evoke an effect in the uterotrophic rodent assay (12). Acute human (13) and wildlife (14) exposure to hormonally active agents can impair reproductive capability in the exposed and their offspring. Environmental studies also show that the potential releases are diverse and include wastewater (15), landfill leachate (16), and cattle feedlot runoff (17). Polychlorinated biphenyls (PCBs) are a class of chlorinated organic compounds that were widely used, particularly in the manufacturing of electrical equipment, because of their favorable dielectric properties and chemical stability. Poor disposal practices released PCBs into the environment where, because of their persistence and lipophilicity, they accumulated in biologic tissue and bioconcentrated at successively higher levels of the food chain. Despite longstanding prohibitions on their manufacture and use, PCBs are still ubiquitous in environmental media, wildlife, food, and humans. PCBs have been associated with reproductive abnormalities in animal models (18), and certain congeners have been shown to have estrogenic activity in experimental models (4, 19). However, the estrogenic activity of some PCB congeners commonly detected in environmental and biologic media has not been assessed. The E-SCREEN (20) assay quantifies estrogenic activity in terms of the proliferation induced in MCF-7-BOS cells, an established, estrogen-responsive human breast cancer cell line widely used to study estrogenic action. The goal of this study was to quantify the estrogenic potency of polychlorinated biphenyls commonly detected in human tissue and the environment using the E-SCREEN assay. Another aim was to estimate the contribution of different exposure pathways to PCB-derived estrogenic potency by combining results of this study with existing environmental monitoring data from New Bedford, MA, an area with a PCB-contaminated hazardous waste site.
Materials and Methods PCB Screening Panel. Thirty-four PCB congeners were selected for screening based on three criteria. Twenty-three were selected because of their presence in human tissue (21, 22) and their likely affinity for the estrogen receptor: PCBs 8, 28, 44, 49, 52, 66, 70, 74, 82, 84, 87, 99, 101, 105, 118, 128, 138, 153, 170, 179, 180, 183, and 187. These congeners have chlorination patterns associated with estrogen receptor affinity (23, 24): ortho chlorination, which imparts rotational rigidity to the biphenyl structure, akin to the planar rigidity of the endogenous hormone, 17β-estradiol (E2); ortho-para chlorination, where the para chlorine(s) resembles 17βestradiol’s hydroxyl groups and the ortho chlorine(s) impart rotational rigidity; and nonchlorinated vicinal para-meta positions, which are susceptible to metabolic hydroxylation with consequent estrogen receptor affinity. Six PCB congenerss15, 17, 18, 31, 77, and 110swere selected because they were prevalent in air, soil, or dust samples from greater New Bedford, MA, a region well-studied because of PCB contamination in New Bedford Harbor (2529). Five congeners were selected because they are used in analytical chemistry procedures as internal standards for VOL. 40, NO. 8, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2819
FIGURE 1. Cell proliferation vs dose for 17β-estradiol. E-SCREEN data from 30 17β-estradiol control plates run concurrently with the PCB screening runs discussed in this paper. Each dose level comprises 180 observations; error bars represent one standard error. quantitation and monitoring recovery efficiency: 11, 30, 103, 112, 166. These were screened to identify internal PCB standards that would not interfere with estrogenicity screening. These congeners have not been reported in human tissue or environmental samples, nor have they been found in commercial PCB preparations (30). E-SCREEN Assay. Estrogenic potency was quantified using the E-SCREEN cell culture bioassay, which quantifies estrogenicity by measuring the proliferation of MCF-7-BOS cells induced by the test material. The MCF-7 human breast cancer cells used in E-SCREEN were obtained from the Michigan Cancer Foundation and cloned by the Sonnenschein and Soto laboratories at Tufts University School of Medicine in 1983; they are now called MCF-7-BOS cells (31). The cells were maintained in Dulbeccos’s modification of Eagle’s medium (DMEM; ICN Biomedicals, Costa Mesa, CA) with phenol red and supplemented with 5% fetal bovine serum (5% FBS; HyClone Laboratories, Logan, UT). The cells were kept in a controlled atmosphere of 6% CO2/94% air and saturating humidity at 37 °C. To assay a test substance, MCF7-BOS cells were trypsinized and seeded into 12-well plates (Falcon, Lincoln Park, NJ) at concentrations of 30 000-40 000 cells per well. The cells were incubated in medium (5% FBS DMEM with phenol red) for 24 h to permit attachment to the bottom of wells. The seeding medium was then washed once with phenol red-free DMEM then replaced with experimental medium composed of 5% charcoal-dextran-stripped fetal bovine serum in phenol red-free DMEM with penicillin (5% CD FBS). Charcoal-dextran stripping eliminates over 99% of sex steroids from bovine serum (32, 33). Neat preparations of all screened PCB congeners were obtained from EM Science (Gibbstown, NJ) and were manufacturer-certified >99.9% pure suitable for gas chromatography/mass spectroscopy. The neat congeners were dissolved in ethanol to a concentration of 10 mM. Just before use, the stock PCB solutions were diluted serially in DMEM without phenol red and added to the experimental medium, with two replicate wells at each dose level. Positive and null controls were included with every E-SCREEN run. The concurrent positive control was one 12well cell culture plate prepared as above with 17β-estradiol in the dose range of 1 × 10-13 to 1 × 10-9 M at one-log dose intervals. The estradiol control provided baseline data for calculating the relative proliferative potency and efficacy of the test compound and confirmed the sensitivity of the cell culture conditions to estradiol (Figure 1). In addition, every plate had two hormone-free controls wells containing only 5% CD FBS experimental medium. 2820
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 40, NO. 8, 2006
Each PCB congener was incubated in MCF-7-BOS cell culture for 5 days (corresponding to the exponential phase of proliferation for this cell line) in a controlled atmosphere of 6% CO2/94% air and saturating humidity at 37 °C. After incubation, experimental medium was aspirated from the wells and the cells were lysed in situ using a solution of 10% ethylhexadecyl-dimethylammonium bromide (Eastman Kodak Company, Rochester, NY) in 0.5% Triton X-100, 2 mM MgCl2, 15 mM NaCl, 5 mM phosphate buffer, pH 7.4. The cell nuclei were counted in a Coulter Counter apparatus, model ZM (Coulter Electronics, Hialeah, FL). Three cell counts were taken for each well with two replicate wells per dose; the cell count coefficient of variation at each dose was typically below 10%. Each estrogenic PCB had six confirmatory E-SCREEN runs, except that PCB 30 had five and PCB 49 had seven confirmatory runs. All confirmatory E-SCREEN runs used the same dose levels. Nonestrogenic PCBs had a minimum of three confirmatory runs. The potency and efficacy of each PCB were calculated relative to 17β-estradiol using data obtained from the concurrent estradiol plate. The relative proliferative potency (RPP; 17β-estradiol RPP ≡ 100%) was calculated from assay data as 100 times the ratio of the lowest doses of PCB and estradiol that induce half-maximal cell proliferation (i.e., the M50 doses). The M50 for each PCB was estimated from data for all assay runs with a logit-log regression of the linear portion of the dose-response curve, while an M50 of 3 × 10-11 M was used for estradiol based on previous baseline experiments. However, for PCBs 17, 18, and 30 the maximum proliferative effect occurred at their lowest experimental dose level, so estimation of the M50 by logit-log regression was not possible. Instead, the RPPs for these PCBs were computed as 100 times the ratio of the minimal dose of the test material and estradiol needed to attain maximal cell yield (i.e., the M100 doses); the M100 for estradiol was set at 1 × 10-10 M (32). Also calculated were the proliferative effect (PE) and the relative proliferative effect (RPE; 17β-estradiol RPE ≡ 100%). The PE is the ratio of the maximum cell yield of each PCB to that of the hormone-free control (5% CD FBS), and the RPE is 100 times the ratio of the PEs for each PCB and estradiol. The PE at 1 × 10-10 M 17β-estradiol was used for all RPE calculations since this is the lowest 17β-estradiol dose at which maximal MCF-7-BOS cell proliferation is induced (32) (Figure 1). The RPE characterizes the agonistic effect of each PCB relative to 17β-estradiol, where an RPE of 100% indicates a full agonist and lower values indicate a partial agonist. Statistical Analysis. Test material evaluated with the E-SCREEN assay is considered estrogenic if it induces a statistically significant increase in cell yield over the hormonefree control. Statistical significance was assessed at the 5% level. Differences in cell yield were evaluated using analysis of variance with Dunnett’s adjustment for multiple comparisons between multiple treatments and a single control. The one-way form of Dunnett’s method was used to test whether there was a statistically significant elevation in mean cell number at each dose level compared to the mean of the control wells. In addition, because the cell count data were nested within each plate, each plate’s data can be expected to be correlated, particularly with respect to each plate’s initial seeding density. This nesting within plates was accounted for by including an indicator variable for plate as a random effect, dose as a fixed effect, and cell number as the dependent variable. The PROC MIXED routine of the Statistical Analysis System (34) implemented the mixed model analyses with Dunnett’s test. Because RPE values less than 20% may represent random variation of the assay rather than estrogenic activity, confirmatory analyses were done with generalized linear mixed modeling (GLMM) techniques (35).
FIGURE 2. PCBs 17, 18, 30, 44, 49, 66, 74, 82, 99, 103, 110, 128, and 179. Proliferative effect (PE) vs dose for 13 estrogenic PCBs. Error bars represent one standard deviation, and each curve represents at least six replications. Note the larger PE scale for PCB30. PEs for volatile PCBs 17, 18, 30, 44, 49, and 103 are based on concurrent estradiol controls. CNTL: 5% CD FBS control. By explicitly accounting for contributions of experimental design to variation in the data and more fully specifying the variance structure, GLMM maximizes the precision of parameter estimates thereby improving statistical power for detecting subtle dose-response effects.
Results Estrogenic PCBs. Thirteen of the 34 congeners studied were estrogenic: PCBs 17, 18, 44, 49, 66, 74, 82, 99, 110, 128, and 179 and PCBs 30 and 103, which are used as internal standards
in analytical chemistry procedures (Figure 2). The 21 nonestrogenic congeners were PCBs 8, 15, 28, 31, 52, 70, 77, 84, 87, 101, 105, 118, 138, 153, 170, 180, 183, 187, as well as three internal standardssPCBs 11, 112, and 166. Table 1 lists the screened PCBs and their estrogenicity parameters, and Figure 2 displays the PCB dose dependence of PE for the estrogenic congeners. For PCBs 17, 18, 66, 74, 103, and 110, cytotoxicity at higher doses exceeded xenoestrogenic induction of MCF7-BOS cell proliferation, causing a net decrease in cell number. VOL. 40, NO. 8, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2821
TABLE 1. E-SCREEN Results for PCB Congeners estrogenicity parameters [%] PCBa
chlorination
8: 11:d 15: 17: 18: 28: 30:d 31: 44: 49: 52: 66: 70: 74: 77: 82: 84: 87: 99: 101: 103:d 105: 110: 112:d 118: 128: 138: 153: 166:d 170: 179: 180: 183: 187:
2,4′ 3,3′ 4,4′ 2,2′,4 2,2′,5 2,4,4′ 2,4,6 2,4′,5 2,2′,3,5′ 2,2′,4,5′ 2,2′,5,5′ 2,3′,4,4′ 2,3′,4′,5 2,4,4′,5 3,3′,4,4′ 2,2′,3,3′,4 2,2′,3,3′,6 2,2′,3,4,5′ 2,2′,4,4′,5 2,2′,4,5,5′ 2,2′,4,5′,6 2,3,3′,4,4′ 2,3,3′,4′,6′ 2,3,3′,5,6 2,3′,4,4′,5 2,2′,3,3′,4,4′ 2,2′,3,4,4′,5′ 2,2′,4,4′,5,5′ 2,3,4,4′,5,6 2,2′,3,3′,4,4′,5 2,2′,3,3′,5,6,6′ 2,2′,3,4,4′,5,5′ 2,2′,3,4,4′,5′,6 2,2′,3,4′,5,5′,6
RPEb
RPPc
11.4 27.6
0.0040 0.0040
52.8
0.0040
25.6 13.1
0.0011 0.0014
9.3
0.0008
16.8
0.0007
46.9
0.0014
14.9
0.0008
38.0
0.0015
17.8
0.0007
14.0
0.0007
29.9
0.0014
a IUPAC congener designations. b RPE ) relative proliferative effect. RPP ) relative proliferative potency. d Internal standards for analytic chemistry procedures. c
For six of the estrogenic PCBss17, 18, 30, 44, 49, and 103smean cell counts in their respective 5% CD FBS control wells exceeded the concurrent estradiol control wells at a statistically significant level, which we determined was due to PCBs volatilizing during incubation. The in vitro volatilization of each of these PCB congeners was confirmed in separate experiments (35). The PEs for the volatilized PCBs were calculated using cell count data from concurrent estradiol 5% CD FBS control wells. Consistent with the weak potency observed for other estrogenic pollutants, the estrogenic PCBs were typically 6 orders of magnitude less potent than the endogenous hormone 17β-estradiol, with RPPs in the range of 0.00070.0040%. At their maximum potency the estrogenic environmental congeners induced MCF-7-BOS cell proliferation from 9.3% of 17β-estradiol for PCB66 to as high as 46.9% for PCB82. All the estrogenic PCBs identified in this study have three to seven total chlorines and are ortho chlorinated. PCBs 17, 30, 49, 66, 74, 82, 99, 103, 110, and 128 have the orthopara chlorination pattern, while PCBs 17, 18, 30, 44, 49, 82, 103, 110, and 179 have at least one pair of nonchlorinated vicinal para-meta positions (Figure 3). Exposure Pathways in New Bedford, MA. We combined PCB E-SCREEN results with previously collected environmental exposure data from the Greater New Bedford area to assess environmental pathways for PCB-derived estrogenic potency. In the prior New Bedford exposure assessment (2528), house dust, yard soil, and indoor and outdoor air were collected from a sample of 34 homes both downwind and upwind from the contaminated New Bedford Harbor. Homes 2822
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 40, NO. 8, 2006
were sampled in 1994-1995 during remediation (dredging) of contaminated Harbor sediments. The estrogenic PCBs 82 and 179 were not quantified in these studies, and two pairs of congeners coeluted in the GC/MS analysis: nonestrogenic PCB77 with estrogenic PCB110 and nonestrogenic PCB15 with estrogenic PCB17. Table 2 displays the concentrations of estrogenic PCBs detected in the New Bedford environmental survey and our estimates of daily PCB-derived estrogenic exposure (estradiol-equivalents, E2EQ). A PCB congener’s percent-molar concentration is proportional to the total PCB moles for all PCB congeners detected in the sample. Estradiol-equivalents were computed by multiplying RPPs from E-SCREEN (Table 1) by congener-specific PCB daily exposures for each medium, based on median molar concentrations and reported inhalation and ingestion rates for children, a population presumed to be particularly susceptible. Maximal inhalation exposures were estimated for a 1-2-year-old child with 6.8 m3/day inhalation rate and 10 kg body mass, and solid-phase exposures were for a 1-6year-old child with 200 mg/day ingestion rate and 10 kg mass (36). For complete representation of estrogenic PCBs quantified in the New Bedford samples, Table 2 includes PCB136, which was previously found to be estrogenic in E-SCREEN (37). Our estimates of PCB-derived estrogenic potency indicate that both solid- and vapor-phase environmental media are potential exposure sources. However, the estimated PCBderived estradiol-equivalent exposure from air was at least 1 order of magnitude higher than estimated exposures from dust or soil (Table 2). Estimated total vapor-phase exposure to estradiol-equivalents was 1.86 × 10-6 (indoor air) and 6.30 × 10-6 (outdoor air) nmol E2EQ/day-kg, while exposure from yard soil was 1.38 × 10-7 nmol E2EQ/day-kg, and that from house dust was 6.02 × 10-8 nmol E2EQ/day-kg. PCBs 17 (coeluting with the nonestrogenic PCB15) and 18 contributed over 80% of estradiol-equivalent exposure in air samples. Over 60% of the estradiol-equivalents in yard soil was attributable to PCBs 18, 99, and PCB110 (where congener 110 accounted for most of the coeluting measurement with nonestrogenic PCB77). PCBs 17 (coeluting with the nonestrogenic PCB15), 18, 44, 49, and 110 (coeluting with nonestrogenic PCB77) together contributed more than 75% of total estradiol-equivalents in dust samples.
Discussion We identified potential estrogen agonist activity for a number of PCB congeners commonly detected in environmental samples and human tissues but for which there are limited prior measures of estrogenicity. These findings should contribute to efforts to improve the sensitivity of health effect studies by grouping toxicologically similar PCBs. Furthermore, of the environmental media assessed, air (inhalation pathway) was by far the highest PCB-derived estradiolequivalent exposure source among the New Bedford samples (Table 2), thereby underscoring the potential toxicologic importance of inhalational PCB exposure. E-SCREEN is a relatively rapid and low-cost assay and is therefore an efficient tool for identifying potential estrogenic activity in a range of contaminants and matrixes. However, unlike animal models, MCF-7 cells have only limited metabolic activity. By using parent PCB compound, we did not assess the potential role of metabolic transformation on estrogen activity. For example, PCB52 has nonchlorinated vicinal para-meta positions susceptible to metabolic hydroxylation that would, in turn, confer potential estrogen activity. Although PCBs 52 and 153 were not estrogenic in our analyses, they were estrogenic in models measuring rat uterine weight (4, 38, 39). The potency of PCBs that are hydrophobic and have relatively high vapor pressure is difficult to assess in multiwell cell culture plates, such as
FIGURE 3. Chemical structures of the estrogenic PCBss17, 18, 30, 44, 49, 66, 74, 82, 99, 103, 110, 128, 179, and 17β-estradiol, as well as a diagram of the IUPAC PCB naming scheme.
TABLE 2. PCB Concentrations from 34 Homes in the Greater New Bedford, MA Area (Adapted from Refs 26-28) and Estimated Estradiol-Equivalents Exposurea Potential PCB-Derived Estradiol-Equivalent Vapor-Phase Exposures for 1-2-Year-Old Child, 6.8 m3/day Inhalation, 10 kg Body Mass (36) indoor air
outdoor air
PCB
percentmolarb
concn [nmol/m3]
PCB exposure [nmol/day-kg]
E2EQ exposure [nmol/day-kg]c
PCB
percentmolarb
concn [nmol/m3]
PCB exposure [nmol/day-kg]
E2EQ exposure [nmol/day-kg]c
18: 15/17:d 44: 77/110:d 49: 66: 74: 99: 136: 128:
8.78 7.25 2.88 2.87 2.48 1.27 1.16 1.15 0.21 0.13
3.31 × 10-2 2.46 × 10-2 1.06 × 10-2 9.62 × 10-3 9.56 × 10-3 5.17 × 10-3 3.66 × 10-3 3.83 × 10-3 1.14 × 10-3 1.66 × 10-4
2.25 × 10-2 1.67 × 10-2 7.21 × 10-3 6.54 × 10-3 6.50 × 10-3 3.52 × 10-3 2.49 × 10-3 2.60 × 10-3 7.75 × 10-4 1.13 × 10-4
9.00 × 10-7 6.69 × 10-7 7.87 × 10-8 4.87 × 10-8 9.40 × 10-8 2.95 × 10-8 1.65 × 10-8 1.98 × 10-8 7.75 × 10-10 7.76 × 10-10
18: 15/17: 49: 44: 77/110: 66: 99: 74: 136: 128:
9.42 7.69 3.59 2.93 2.04 1.24 1.08 0.99 0.16 0.04
1.07 × 10-1 8.07 × 10-2 4.20 × 10-2 4.17 × 10-2 3.82 × 10-2 1.76 × 10-2 2.03 × 10-2 1.50 × 10-2 2.91 × 10-3 8.31 × 10-4
7.28 × 10-2 5.49 × 10-2 2.86 × 10-2 2.84 × 10-2 2.60 × 10-2 1.20 × 10-2 1.38 × 10-2 1.02 × 10-2 1.98 × 10-3 5.65 × 10-4
2.91 × 10-6 2.20 × 10-6 4.13 × 10-7 3.10 × 10-7 1.93 × 10-7 1.00 × 10-7 1.05 × 10-7 6.76 × 10-8 1.98 × 10-9 3.89 × 10-9
total:
1.86 × 10-6
total:
6.30 × 10-6
Potential PCB-Derived Estradiol-Equivalent Solid-Phase Exposures for 1-6-Year-Old Child, 200 mg/day Soil Ingestion, 10 kg Body Mass (36) house dust PCB
percentmolarb
concn [nmol/g]
77/110:d 66: 99: 44: 49: 74: 18: 128: 15/17:d 136:
8.04 2.30 1.93 1.84 1.63 1.37 1.17 1.01 0.92 0.25
6.67 × 10-2 3.03 × 10-2 2.28 × 10-2 3.07 × 10-2 2.62 × 10-2 2.07 × 10-2 1.58 × 10-2 1.18 × 10-2 1.29 × 10-2 2.85 × 10-3
yard soil
PCB exposure [nmol/day-kg]
E2EQ exposure [nmol/day-kg]c
1.33 × 10-3 6.06 × 10-4 4.56 × 10-4 6.14 × 10-4 5.24 × 10-4 4.14 × 10-4 3.16 × 10-4 2.36 × 10-4 2.58 × 10-4 5.70 × 10-5
9.93 × 10-9 5.08 × 10-9 3.47 × 10-9 6.70 × 10-9 7.58 × 10-9 2.74 × 10-9 1.26 × 10-8 1.62 × 10-9 1.03 × 10-8 5.70 × 10-11
total:
6.02 × 10-8
PCB 77/110: 99: 128: 49: 66: 44: 74: 15/17: 18: 136:
percentmolarb
concn [nmol/g]
9.68 2.68 1.75 1.42 1.42 1.36 0.64 0.27 0.40
3.73 × 10-1 1.10 × 10-1 7.00 × 10-2 3.89 × 10-2 5.02 × 10-2 4.61 × 10-2 2.38 × 10-2 1.06 × 10-2 1.79 × 10-2 ND
PCB exposure [nmol/day-kg]
E2EQ exposure [nmol/day-kg]c
7.46 × 10-3 2.20 × 10-3 1.40 × 10-3 7.78 × 10-4 1.00 × 10-3 9.22 × 10-4 4.76 × 10-4 2.12 × 10-4 3.58 × 10-4 NA
5.55 × 10-8 1.67 × 10-8 9.63 × 10-9 1.13 × 10-8 8.42 × 10-9 1.01 × 10-8 3.16 × 10-9 8.48 × 10-9 1.43 × 10-8 NA
total:
1.38 × 10-7
NB: estrogenic PCBs 82 and 179 not quantified in refs 25-28. Percent-molar is the proportion of total PCB moles, based on median concentrations. E2EQ: estradiol-equivalents, based on relative estrogenic potency (RPP) from E-SCREEN assay. d Total moles for coeluting congeners (PCBs 15 with 17 and 77 with 110) were calculated using the molecular weight of the heaviest congener. a
b
c
those used in the E-SCREEN assay. This is due to volatilization that may induce false positive responses in nearby control wells (40). However, we were able to avoid this problem by using uncontaminated controls from concurrent estradiol control plates when potential control contamination on the PCB plates were identified. These limitations notwithstanding, where prior estrogenicity assessments have been done, our findings are generally in agreement. PCBs 8, 11, and 15 were not estrogenic by our E-SCREEN assay, consistent with the absence of a conclusive estrogenic effect in a study of uterotrophic activity in Wistar
rats (39). PCB30 was estrogenic in this study, which is consistent with results in a 14-day MCF-7 foci assay, and the RPEs of both studies were comparable: RPE ) 52.8% in E-SCREEN and estimated RPE ) 48% in Gierthy et al. (19). PCB105 was not estrogenic in this study, but was previously reported to have antiestrogenic activity based on decreased 17β-estradiol-induced procathepsin-D production in MCF-7 cells (41). PCB77 did not induce any detectable cell proliferation in the E-SCREEN assay consistent with its antiestrogenic activity in rodent models (3, 4). Elsewhere, however, this congener VOL. 40, NO. 8, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2823
has been estrogenic in a variety of assays, including an MCF-7 cell proliferation assay (42, 43). Nesaretnam et al. (42) comprehensively evaluated the estrogenicity of PCB77, including assaying competitive binding to cell-free human estrogen receptor, proliferation and luciferase induction in estrogen-responsive cell lines (MCF-7 and ZR-75-1), and uterine weight gain in immature female CD1 mice, and the results from these experiments indicated that PCB77 has estrogenic potency. In a later study by Ramamoorthy et al. (3), PCB77 was found to be antiestrogenic, with the discrepancy of these findings with Nesaretnam et al. (42) attributed to possible impurities in the PCB77 standard used by Nesaretnam et al. The Ramamoorthy team synthesized their own PCB77, while Nesaretnam et al. obtained their PCB77 from Promochem (St. Albans, Herts, England), which they reported to be certified sufficiently pure for GC/MS. In the present E-SCREEN study, PCB77 was obtained as a neat preparation from EM Science and manufacturer-certified to be >99.9% pure and suitable for GC/MS. Because of differences in body size, diet, metabolism, and exposure risk, children generally experience higher relative exposures to environmental contaminants than adults. For example, using our E-SCREEN results and the New Bedford environmental monitoring data (Table 2), a 1-2-year-old child’s total vapor-phase daily exposure would be approximately 4.08 × 10-3 pmol E2EQ/day-kg (assuming time spent equally indoors and outdoors), whereas a 70-kg adult male’s exposure (assuming 15.2 m3/day inhalation rate (44)), would be approximately 1.30 × 10-3 pmol E2EQ/day-kg, about 30% that of a young child’s. Even in a group at high risk for exposure (1-5-year-old children), compared to the serum estradiol reference range at this age (11-37 pmol E2/L for boys and 18-37 pmol E2/L for girls (45)), daily PCBderived estrogen-equivalent exposure from New Bedford air or soil/dust does not appear to be near physiologic levels, assuming complete PCB absorption and simple additive estrogenic potencies. There are other important sources of PCB exposure, however, including consumption of contaminated fish, other foods, or breast milk that were not assessed in this study. Furthermore, over 97% of serum estradiol is bound to proteins, and it is unbound estradiol that is believed to be the physiologically active hormone (46). Most xenoestrogens such as PCBs do not have high binding affinity for sex hormone binding globulin (SHBG), the main estrogen binder in human serum, so xenoestrogens may be more readily available to the estrogen-target cells than natural estrogens (47, 48). Thus, although estradiol-equivalents deriving from PCBs in New Bedford air, soil, and dust samples were estimated to be at very low picomole levels, they may nonetheless contribute to undesirable hormonal effects through enhanced bioavailability, bioaccumulation, and combination with other xenoestrogen exposures. Furthermore, there may be critical periods in early development that are particularly sensitive to low-level exposures. The likelihood of potential effects at low doses was suggested by publications showing that developmental effects in animals occur after in utero exposure to low doses of estrogenic chemicals (6, 8). For example, a recent study (11) showed that bisphenol-A administered perinatally altered mouse mammary gland development at a dose 4 000 000-fold lower than the effective dose in the uterotrophic assay (12). But, findings among low-dose exposure studies have been inconsistent with some showing no effect (49, 50) and others showing enhanced estrogenic effects with low-dose exposure mixtures (51, 52). This highlights the challenges of inferring the net biologic activity of complex mixtures of agonists and antagonists that characterize most environmental and biological specimens. As a consequence, the human health risk posed by pollutant-derived estrogenic potency remains an 2824
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 40, NO. 8, 2006
area of active inquiry. PCB estrogenic potency information available from this study provides an important resource both for preliminary estimation of routes of human exposure to xenoestrogens and for application to human health studies focused on estrogen-responsive health outcomes such as reproductive development and related malignancies.
Acknowledgments B.R.d.C. received support from the Harvard School of Public Health and the Tufts University School of Medicine. This work was supported in part by NIEHS Grant Nos. ES000002, ES05947, and ES08314. The content of this report is solely the responsibility of the authors and does not necessarily represent the official views of the NIEHS or NIH. We thank Carlos Sonnenschein, Larisa Altshul, Maria Luizzi, Cheryl Michaelson, and Nancy Prechtl for their advice.
Literature Cited (1) ) NRC. Hormonally Active Agents in the Environment; National Research Council: Washington, DC, 1999. (2) Soto, A. M.; Chung, K. L.; Sonnenschein, C. The pesticides endosulfan, toxaphene, and dieldrin have estrogenic effects on human estrogen-sensitive cells. Environ. Health Perspect. 1994, 102 (4), 380-383. (3) Ramamoorthy, K.; Gupta, M. S.; Sun, G.; McDougal, A.; Safe, S. H. 3,3′4,4′-Tetrachlorobiphenyl exhibits antiestrogenic and antitumorigenic activity in the rodent uterus and mammary cells and in human breast cancer cells. Carcinogenesis 1999, 20 (1), 115-123. (4) Jansen, H. T.; Cooke, P. S.; Porcelli, J.; Liu, T. C.; Hansen, L. G. Estrogenic and antiestrogenic actions of PCBs in the female rat: in vitro and in vivo studies. Reprod. Toxicol. 1993, 7 (3), 237248. (5) Krishnan, V.; Porter, W.; Santostefano, M.; Wang, X.; Safe, S. Molecular mechanism of inhibition of estrogen-induced cathepsin D gene expression by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in MCF-7 cells. Mol. Cell. Biol. 1995, 15 (12), 67106719. (6) vom Saal, F. S.; Timms, B. G.; Montano, M. M.; Palanza, P.; Thayer, K. A.; Nagel, S. C.; Dhar, M. D.; Ganjam, V. K.; Parmigiani, S.; Welshons, W. V. Prostate enlargement in mice due to fetal exposure to low doses of estradiol or diethylstilbestrol and opposite effects at high doses. Proc. Natl. Acad. Sci. U.S.A. 1997, 94 (5), 2056-2061. (7) Ulrich, E. M.; Caperell-Grant, A.; Jung, S. H.; Hites, R. A.; Bigsby, R. M. Environmentally relevant xenoestrogen tissue concentrations correlated to biological responses in mice. Environ. Health Perspect. 2000, 108 (10), 973-97. (8) Ramos, J. G.; Varayoud, J.; Sonnenschein, C.; Soto, A. M.; Mun ˜ oz de Toro, M.; Luque, E. H. Prenatal exposure to low doses of bisphenol A alters the periductal stroma and glandular cell function in the rat ventral prostate. Biol. Reprod. 2001, 65, 12711277. (9) Markey, C. M.; Wadia, P. R.; Rubin, B. S.; Sonnenschein, C.; Soto, A. M. Long-term effects of fetal exposure to low doses of the xenoestrogen bisphenol A in the female mouse genital tract. Biol. Reprod. 2005, 72 (6), 1344-1351. (10) Markey, C. M.; Luque, E. H.; Munoz de Toro, M.; Sonnenschein, C.; Soto, A. M. In utero exposure to bisphenol A alters the development and tissue organization of the mouse mammary gland. Biol. Reprod. 2001, 65, 1215-1223. (11) Munoz-de-Toro, M.; Markey, C. M.; Wadia, P. R.; Luque, E. H.; Rubin, B. S.; Sonnenschein, C.; Soto, A. M. Perinatal exposure to bisphenol A alters peripubertal mammary gland development in mice. Endocrinology 2005, 146 (9), 4138-4147. (12) Markey, C. M.; Michaelson, C. L.; Veson, E. C.; Sonnenschein, C.; Soto, A. M. The mouse uterotrophic assay: a reevaluation of its validity in assessing the estrogenicity of bisphenol A. Environ. Health Perspect. 2001, 109 (1), 55-60. (13) Giusti, R. M.; Iwamoto, K.; Hatch, E. E. Diethylstilbestrol revisited: a review of the long-term health effects. Ann. Intern. Med. 1995, 122 (10), 778-788. (14) Facemire, C. F.; Gross, T. S.; Guillette, L. J., Jr. Reproductive impairment in the Florida panther: nature or nurture? Environ. Health Perspect. 1995, 103 (Suppl 4), 79-86. (15) Schiliro, T.; Pignata, C.; Fea, E.; Gilli, G. Toxicity and estrogenic activity of a wastewater treatment plant in Northern Italy. Arch. Environ. Contam. Toxicol. 2004, 47 (4), 456-62.
(16) Behnisch, P. A.; Fujii, K.; Shiozaki, K.; Kawakami, I.; Sakai, S. Estrogenic and dioxin-like potency in each step of a controlled landfill leachate treatment plant in Japan. Chemosphere 2001, 43 (4-7), 977-84. (17) Soto, A. M.; Calabro, J. M.; Prechtl, N. V.; Yau, A. Y.; Orlando, E. F.; Daxenberger, A.; Kolok, A. S.; Guillette, L. J., Jr.; le Bizec, B.; Lange, I. G.; Sonnenschein, C. Androgenic and estrogenic activity in water bodies receiving cattle feedlot effluent in Eastern Nebraska, U.S.A. Environ. Health Perspect. 2004, 112 (3), 34652. (18) Lind, P. M.; Eriksen, E. F.; Sahlin, L.; Edlund, M.; Orberg, J. Effects of the antiestrogenic environmental pollutant 3,3′,4,4′,5pentachlorobiphenyl (PCB #126) in rat bone and uterus: diverging effects in ovariectomized and intact animals. Toxicol. Appl. Pharmacol. 1999, 154 (3), 236-44. (19) Gierthy, J. F.; Arcaro, K. F.; Floyd, M. Assessment of PCB estrogenicity in a human breast cancer cell line. Chemosphere 1997, 34 (5-7), 1495-1505. (20) Soto, A. M.; Lin, T.-M.; Justicia, M.; Silvia, R. M.; Sonnenschein, C. An “In Culture” Bioassay to Assess the Estrogenicity of Xenobiotics (E-SCREEN). In Chemically-Induced Alterations in Sexual and Functional Development: The Wildlife/Human Connection; Colburn, T., Clement, C., Eds.; Princeton Scientific Publishing Co., Inc.: Princeton, NJ, 1992; pp 295-309. (21) Gladen, B. C.; Doucet, J.; Hansen, L. G. Assessing human polychlorinated biphenyl contamination for epidemiologic studies: lessons from patterns of congener concentrations in Canadians in 1992. Environ. Health Perspect. 2003, 111 (4), 43743. (22) Orloff, K. G.; Dearwent, S.; Metcalf, S.; Kathman, S.; Turner, W. Human exposure to polychlorinated biphenyls in a residential community. Arch. Environ. Contam. Toxicol. 2003, 44 (1), 12531. (23) Korach, K. S.; Sarver, P.; Chae, K.; McLachlan, J. A.; McKinney, J. D. Estrogen receptor-binding activity of polychlorinated hydroxybiphenyls: conformationally restricted structural probes. Mol. Pharmacol. 1988, 33 (1), 120-126. (24) Katzenellenbogen, B. S.; Bhardwaj, B.; Fang, H.; Ince, B. A.; Pakdel, F.; Reese, J. C.; Schodin, D.; Wrenn, C. K. Hormone binding and transcription activation by estrogen receptors: analyses using mammalian and yeast systems. J. Steroid Biochem. Mol. Biol. 1993, 47 (1-6), 39-48. (25) Cullen, A. C.; Vorhees, D. J.; Altshul, L. M. Influence of harbor contamination on the level and composition of polychlorinated biphenyls in produce in Greater New Bedford, Massachusetts. Environ. Sci. Technol. 1996, 30, 1581-1588. (26) Vorhees, D. J. Multi-Media Human Exposure to Polychlorinated Biphenyls. Ph.D. Thesis, Harvard School of Public Health, Boston, MA, 1996. (27) Vorhees, D. J.; Cullen, A. C.; Altshul, L. M. Exposure to polychlorinated biphenyls in residential indoor air and outdoor air near a Superfund site. Environ. Sci. Technol. 1997, 31 (12), 3612-3618. (28) Vorhees, D. J.; Cullen, A. C.; Altshul, L. M. Polychlorinated biphenyls in house dust and yard soil near a Superfund site. Environ. Sci. Technol. 1999, 33 (13), 2151-2156. (29) Choi, A. L.; Levy, J. I.; Dockery, D. W.; Ryan, L. M.; Tolbert, P. E.; Altshul, L. M.; Korrick, S. A. Does living near a Superfund site contribute to higher polychlorinated biphenyl (PCB) exposure? Environ Health Perspect., in press. (30) Jones, K. C. Determination of polychlorinated biphenyls in human foodstuffs and tissues: suggestions for a selective congener analytical approach. Sci. Total Environ. 1988, 68, 141159. (31) Villalobos, M.; Olea, N.; Brotons, J. A.; Olea-Serrano, M. F.; Ruiz de Almodovar, J. M.; Pedraza, V. The E-SCREEN assay: a comparison of different MCF7 cell stocks. Environ. Health Perspect. 1995, 103 (9), 844-850. (32) Soto, A. M.; Sonnenschein, C. The role of estrogens on the proliferation of human breast tumor cells (MCF-7). J. Steroid Biochem. 1985, 23 (1), 87-94. (33) Heringa, M. B.; van der Burg, B.; van Eijkeren, J. C.; Hermens, J. L. Xenoestrogenicity in in vitro assays is not caused by
(34) (35) (36)
(37)
(38)
(39) (40)
(41)
(42) (43) (44) (45) (46)
(47)
(48)
(49) (50) (51)
(52)
displacement of endogenous estradiol from serum proteins. Toxicol. Sci. 2004, 82 (1), 154-63. SAS. SAS/STAT Software; SAS Institute: Cary, NC, 1996. deCastro, B. R. Environmental Estrogens: Biologic Activity and Exposure Assessment. Ph.D. Thesis, Harvard School of Public Health, Boston, MA, 2000. US EPA. Child-Specific Exposure Factors Handbook (Interim Report); EPA-600-P-00-002B; National Center for Environmental Assessment, Office of Research and Development: Washington, DC, 2002. Soto, A. M.; Sonnenschein, C.; Chung, K. L.; Fernandez, M. F.; Olea, N.; Serrano, F. O. The E-SCREEN assay as a tool to identify estrogens: an update on estrogenic environmental pollutants. Environ. Health Perspect. 1995, 103 (Suppl 7), 113-122. Desaulniers, D.; Leingartner, K.; Wade, M.; Fintelman, E.; Yagminas, A.; Foster, W. G. Effects of acute exposure to PCBs 126 and 153 on anterior pituitary and thyroid hormones and FSH isoforms in adult Sprague Dawley male rats. Toxicol. Sci. 1999, 47 (2), 158-169. Ecobichon, D. J.; MacKenzie, D. O. The uterotropic activity of commercial and isomerically pure chlorobiphenyls in the rat. Res. Commun. Chem. Pathol. Pharmacol. 1974, 9 (1), 85-95. Ciapetti, G.; Granchi, D.; Verri, E.; Savarino, L.; Stea, S.; Savioli, F.; Gori, A.; Pizzoferrato, A. False positive results in cytotoxicity testing due to unexpectedly volatile compounds. J. Biomed. Mater. Res. 1998, 39 (2), 286-291. Krishnan, V.; Safe, S. Polychlorinated biphenyls (PCBs), dibenzop-dioxins (PCDDs), and dibenzofurans (PCDFs) as antiestrogens in MCF-7 human breast cancer cells: quantitative structureactivity relationships. Toxicol. Appl. Pharmacol. 1993, 120 (1), 55-61. Nesaretnam, K.; Corcoran, D.; Dils, R. R.; Darbre, P. 3,4,3′,4′Tetrachlorobiphenyl acts as an estrogen in vitro and in vivo. Mol. Endocrinol. 1996, 10 (8), 923-936. Nesaretnam, K.; Darbre, P. 3,5,3′,5′-Tetrachlorobiphenyl is a weak oestrogen agonist in vitro and in vivo. J. Steroid Biochem. Mol. Biol. 1997, 62 (5-6), 409-418. US EPA. Exposure Factors Handbook; EPA/600/P-95/002Fa,b,c; National Center for Environmental Assessment, Office of Research and Development: Washington, DC, 1997. Williams, R. H., Larsen, P. R., Kronenberg, H. M., Melmed, S., Polonsky, K. S., Eds. Williams Textbook of Endocrinology; W. B. Saunders: Philadelphia, PA, 2003; p 1820. Hammond, G. L.; Nisker, J. A.; Jones, L. A.; Siiteri, P. K. Estimation of the percentage of free steroid in undiluted serum by centrifugal ultrafiltration-dialysis. J. Biol. Chem. 1980, 255 (11), 5023-5026. Hodgert Jury, H.; Zacharewski, T. R.; Hammond, G. L. Interactions between human plasma sex hormone-binding globulin and xenobiotic ligands. J. Steroid Biochem. Mol. Biol. 2000, 75 (2-3), 167-176. Nagel, S. C.; vom Saal, F. S.; Welshons, W. V. Developmental effects of estrogenic chemicals are predicted by an in vitro assay incorporating modification of cell uptake by serum. J. Steroid Biochem. Mol. Biol. 1999, 69 (1-6), 343-57. Singleton, D. W.; Khan, S. A. Xenoestrogen exposure and mechanisms of endocrine disruption. Front. Biosci. 2003, 8, s110-8. Witorsch, R. J. Low-dose in utero effects of xenoestrogens in mice and their relevance to humans: an analytical review of the literature. Food Chem. Toxicol. 2002, 40 (7), 905-12. Silva, E.; Rajapakse, N.; Kortenkamp, A. Something from “nothing”seight weak estrogenic chemicals combined at concentrations below NOECs produce significant mixture effects. Environ. Sci. Technol. 2002, 36 (8), 1751-176. Rajapakse, N.; Silva, E.; Kortenkamp, A. Combining xenoestrogens at levels below individual no-observed-effect concentrations dramatically enhances steroid hormone action. Environ. Health Perspect. 2002, 110 (9), 917-21.
Received for review August 22, 2005. Revised manuscript received February 3, 2006. Accepted February 6, 2006. ES051667U
VOL. 40, NO. 8, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2825
