by Kenneth J. Edwards, Jr., VP Alken-Murray Corp.
Phthalates, prominently used in many industries, including plastics, belong to a family of chemical compounds which are based on a benzene ring, to which is attached a pair of carbonyl groups in consecutive positions on the benzene ring.
Commentary & examples:
A benzene ring consists of a backbone of 6 carbons and 6 hydrogens. This structure can be represented as if it was alternating between 1,3,5-cyclohexatriene (a) and 2,4,6-cyclohexatriene (b), but the truth is that the extra electrons are not localized over the double bonds between the carbons. Single hydrogen atoms would also be associated with each carbon represented in the drawing. In reality, the electrons of the benzene ring are gathered into doughnut shaped clouds above and below the ring. Therefore, a simplified representation was adopted by chemists to represent the six carbons and their implicit six hydrogens, as a six sided figure (each point represents a carbon atom), and a circle in its center (the doughnut shaped electron cloud) (c)
Carbons which are double-bonded to an oxygen are called a "carbonyl" group. Thus, (d)has a carbonyl group which has a carbon on one side and an "R" group on the other side. "R" is used as a generic representation of an organic molecule or simply an atom, to enable chemists to make generalizations about families of chemicals which have certain parts of their chemical structure in common, except for the "R" group. (e) is the simplified version of (d).
All phthalates follow the pattern shown by (f), where R1 and R2 may represent various atoms or functional groups. The most common and simplest phthalate is phthalic acid(g). I find phthalic acid is of special interest. If one of the OH groups is neutralized with potassium hydroxide, to give an OK, instead of OH, the resulting compound is called potassium hydrogen phthalate. This compound is used as a pH 4.00 reference, to calibrate pH meter electrodes. It has exceptional buffering abilities and changes its pH very little with significant temperature variation.
(h) is still considered a member of the phthalate family because of the portion of the molecule shown in red.
Phthalic anhydride (i) and phthalimide (j) also belong to the phthalate family in spite of the fact that both carbonyl groups share an atom between them.
Those of our bacteria which can degrade phthalates should be able to degrade at least the phthalate portion of the molecules shown in g, h, i and j.
See the University of Minnesota Biocatalysis-Biodegradation pathway for Phthalates, to see the mechanism for biodegrading Phthalates.
Phthalates, Alkylphenols, Pesticides, Polybrominated Diphenyl Ethers,
and other Endocrine-Disrupting Compounds in
Indoor Air and Dust
Environmental Science & Technology (online) 13sep03
Ruthann A Rudel, * , † David E Camann, ‡ John D Spengler, § Leo R Korn , | and Julia G Brody †
† Silent Spring Institute, ‡ Southwest Research Institute, § Harvard University School of Public Health, | University of Medicine and Dentistry of New Jersey, * Corresponding author phone: (617)332-4288; fax: (617)332-4284; e-mail: email@example.com
Silent Spring Institute, 29 Crafts Street, Newton, Massachusetts 02458, Southwest Research Institute, 6220 Culebra Road, P.O. Box 28510, San Antonio, Texas 78228-0510, Environmental Science and Engineering Program, Harvard University School of Public Health, Landmark Center, 401 Park Drive, Boston, Massachusetts 02115, and Division of Biometrics, University of Medicine and Dentistry of New Jersey, School of Public Health, 335 George Street, Liberty Plaza, Suite 2200, New Brunswick, New Jersey 08903-2688
Chemicals identified as endocrine-disrupting compounds (EDCs) have widespread consumer uses, yet little is known about indoor exposure. We sampled indoor air and dust in 120 homes, analyzing for 89 organic chemicals identified as EDCs. Fifty-two compounds were detected in air and 66 were detected in dust. These are the first reported measures in residential environments for over 30 of the compounds, including several detected at the highest concentrations. The number of compounds detected per home ranged from 13 to 28 in air and from 6 to 42 in dust. The most abundant compounds in air included phthalates (plasticizers, emulsifiers), o-phenylphenol (disinfectant), 4-nonylphenol (detergent metabolite), and 4-tert-butylphenol (adhesive) with typical concentrations in the range of 50- 1500 ng/m3. The penta- and tetrabrominated diphenyl ethers (flame retardants) were frequently detected in dust, and 2,3-dibromo-1-propanol, the carcinogenic intermediate of a flame retardant banned in 1977, was detected in air and dust. Twenty-three pesticides were detected in air and 27 were detected in dust, the most abundant being permethrins and the synergist piperonyl butoxide. The banned pesticides heptachlor, chlordane, methoxychlor, and DDT were also frequently detected, suggesting limited indoor degradation. Detected concentrations exceeded government health-based guidelines for 15 compounds, but no guidelines are available for 28 compounds, and existing guidelines do not consider endocrine effects. This study provides a basis for prioritizing toxicology and exposure research for individual EDCs and mixtures and provides new tools for exposure assessment in health studies.
Current widespread interest in a range of health effects potentially associated with endocrine-disrupting compounds (EDCs) has made exposure assessment for these compounds a priority. Studies of potential health effects associated with EDCs have been hampered by lack of information about the major sources of exposure to EDCs. Furthermore, because many EDCs act additively through a common mechanism of action or have antagonistic or other interactive effects by operating at different points in cell signaling systems, consideration of exposure to mixtures is critical in studies of health effects (1-7). These questions are particularly important in relation to indoor environments, which have been identified as an important source of chemical exposures (8-11). People spend a large fraction of their time indoors, and indoor sources of chemicals, coupled with limited ventilation and slow chemical degradation processes, cause increased pollutant concentrations indoors. In fact, indoor air specifically has been described as “one of the most serious environmental risks to human health” (8).
Many high production volume chemicals—including some already identified as EDCs—have consumer uses (e.g., in plastics, detergents, and other household and consumer products) that make them potentially important indoor contaminants. While a number of comprehensive exposure studies have been conducted or are underway to characterize residential exposures to selected contaminants, particularly volatile organic compounds, pesticides, and polyaromatic hydrocarbons (PAHs), these studies have been limited to a small number of compounds and have focused on characterizing exposure pathways and sources (12-18). We were unable to locate exposure data for many of our compounds of interest, including alkylphenols, parabens, polybrominated diphenyl ethers (PBDEs), and many of the estrogenic phenolic compounds such as bisphenol A. We located only one (unpublished) study of substantial size that has characterized phthalate concentrations in indoor air (18).
The primary objective of this study is to provide an assessment of household exposure to a broad suite of organic chemicals that have been identified as EDCs. Indoor air and dust were selected for analysis because many EDCs are used in consumer products and building materials (6, 19), so these chemicals would be expected indoors. Indoor air has been identified as an important source of chemical exposure, while house dust has been demonstrated to be an important exposure pathway in young children (20). Dust also provides a record of chemicals that have been used in the home historically since degradation processes indoors are typically slow (21).
The chemicals targeted for analysis included phthalates, alkylphenols, pesticides, parabens, PBDEs, PAHs, polychlorinated biphenyls (PCBs), and other estrogenic phenols such as bisphenol A. These compounds were selected if there was evidence that they were EDCs, if they were reported to be present in commercial products or building materials, and/ or if they were compatible with one of two analytical methods being used for these samples. We previously reported on the selection of target compounds and methods for measuring them in air and dust (22).
This paper describes the analytical results for indoor air and house dust samples from 120 homes on Cape Cod, MA. Air and dust samples were analyzed for 89 target chemicals, many identified as EDCs. The large number of homes provides insight into population distributions of exposure to target compounds, and the large number of analytes provides insight into typical mixtures of EDCs to which people are exposed. Table 1 provides an overview of the study design. In addition to the air and dust samples, we collected a urine sample from a resident of the home and a detailed questionnaire about product use and home construction. We also used a geographic information system (GIS) to estimate the relative exposure at each home from historical wide-area pesticide use (23). Finally, air samples were extracted, and total estrogenic activity was determined using an MCF-7 cell proliferation assay (E-SCREEN) (24). Relationships across these measures will be reported separately. This household exposure study was conducted as part of a case-control epidemiologic study of breast cancer on Cape Cod, MA (25).
TABLE 1. Number of Analytes and Related Data Collection by Chemical Group for Samples Taken in 120 Homes on Cape Cod, MAa
no. of related analytes data collection chemical group dust air urineb interviewb GIS-basedb pesticides 38 39 13 + + alkylphenols 7 7 ~ phthalates 10 9 8 ~ PCBs, PAHs, PBDEs 10 10 ~ parabens 3 3 other estrogenic 18 20 phenols and misc. estrogenic activity + (E-SCREEN MCF-7 bioassay)b a+, data of this type were collected in this study. ~, limited questions related to sources of these compounds were included in the interview. b These data will be reported in subsequent papers.
Participant Selection. Eligible women were either breast cancer cases or age-matched controls, were currently alive and residing on Cape Cod, and had lived in their home at least 10 yr at the time of the sampling. To enhance variability across subjects and improve the precision of estimates of upper and lower percentiles of exposure distributions for pesticides, we oversampled individuals with higher and lower potential for pesticide exposure based on self-reported pesticide use and a GIS-derived measure of historical wide area application of persistent pesticides. Sampling was conducted in two rounds of 60 homes per round, beginning in June 1999 and ending in September 2001. All sample collection and analyses were the same for both rounds, although minor changes were made to the target analyte list between rounds.
Air. The 24-h indoor air samples of particulate <5 µm and vapor phase materials were collected using a quiet indoor flow-controlled model SP-280 pump (Air Diagnostics and Engineering, Harrison, ME) modified to collect three parallel 160-mm URG personal pesticide sampling cartridges (University Research Glassware, Chapel Hill, NC). Each URG cartridge contained an impactor-equipped inlet (10 µm at 4 L/min) followed by a glass cartridge that was fitted with a 25-mm quartz fiber filter followed by a 3.0-g bed of XAD-2 resin sandwiched between two 113/16 in. diameter polyurethane foam plugs. Preparation of the URG cartridges is described in our earlier paper (22). Pumps were operated at a constant flow rate of 20-24 L/min. Flow control valves were used to control flow rates for the three parallel URG cartridges so that two samples were collected at flow rates of 8-9 L/min, and a third was collected at 4 L/min. Actual flow rates were determined at the beginning and end of the 24-h sample collection period using a high high-flow Gilian Gilibrator primary standard flow calibrator (Environmental Monitoring Supply). The two URGs collected at the higher flow rate were used for extraction and analysis by the two analytical methods, while the third URG was used to collect duplicate or other samples. The total volume of air sampled ranged from 10 to 14 m3 for the primary samples and from 4 to 6 m3 for the duplicate samples.
On day 1 of sample collection, the pump was placed in a frequently used room of the home, such as the living room or family room, and the URGs were suspended so that the intakes were directeddownward4 ft from the floor. The pump was then calibrated and turned on. On day 2, the URGs were disconnected, and the flow was checked. URGs were stored at -4 °C and then shipped on dry ice to Southwest Research Institute (SWRI) in San Antonio, TX, where they were extracted and analyzed.
Dust. Dust samples were collected using a Eureka Mighty-Mite vacuum cleaner, 9 amp, modified to collect dust into a 19 × 90 mm cellulose extraction thimble (Whatman Inc., Clifton, NJ). Because of the number of our target analytes associated with plastic materials, a custom crevice tool with a holder for the extraction thimble was constructed of PTFE Teflon so dust did not contact any plastic parts of the vacuum. Dust sample collection did not begin until the air sample collection was complete. Sample collection was accomplished by slowly and lightly drawing the crevice tool just above the surface of rugs, upholstery, wood floors, windowsills, ceiling fans, and furniture in each room. Sampling was conducted in the most frequently used rooms of the house, usually 4-5 rooms and including hallways. Unfinished/semifinished areas such as basements, attics, and garages were not sampled. Using this technique and collecting for 45-90 min, approximately4gof dust was collected per sample. Cellulose thimbles containing dust were removed and placed in precleaned, certified glass jars with Teflon-lined lids (Environmental Sampling Supply, Oakland, CA). Samples were stored at -4 °C until they were shipped overnight on dry ice to SWRI. Prior to extraction, dust was tapped out of the thimbles, weighed, and sieved to <150 µm. These samples were split into aliquots for extraction and analysis by each of the two methods. Fourteen samples were split into a larger number of aliquots, with and without spiking with target compounds to determine recovery efficiency. Final sample masses of aliquots used for extraction and analysis ranged from 0.047 to 1.6 g per method (median 0.385 g).
Chemical Analysis. Chemical analysis of air and dust samples was conducted at SWRI. Two GC/MS analytical methods were used to analyze a total of 88 target compounds in air and 86 compounds in dust samples (total of 89 different compounds). One method targets neutrally extracted pesticides, phthalates, PAHs, PBDEs, and PCBs. The second method, which requires derivitazation of the extract prior to analysis, targets alkylphenols—specifically 4-nonylphenol, 4-octylphenol, and their mono- and diethoxylates as well as parabens and other phenols and biphenyls identified as EDCs. The chlorpyrifos metabolite and degradation product 3,5,6-trichloropyridinol and the methoxychlor metabolite/ degradation product 2,2-bis(p-hydroxyphenyl)-1,1,1-trichloroethane (HPTE) and some chlorinated phenols were also included as target analytes of the phenols method. All target analytes are included in Supporting Information Tables S1 (air) and S2 (dust).
Neutrals/Phthalates Extraction and Analysis. Each sieved (<150 µm) dust sample was spiked with the required amount of surrogate solution, 40 ng/mL p-terphenyl-d14, and/or matrix spike solutions (in hexane) depending on the actual size of the dust sample. The spiked dust samples were equilibrated for 30 min at room temperature and then Soxhlet extracted using 6% diethyl ether in hexane for 16 h. The extracts were concentrated to 10 mL, and a 1-mL aliquot was cleaned by running through a florisil column (elution with 20 mL 10% acetone in hexane). When less than 2 g of sieved dust was available, proportionately smaller amounts of surrogates were spiked, and extracts were concentrated to proportionately smaller volumes. The florisil eluent was concentrated to a final volume of 2 mL with 10% ether in hexane for analysis by GC/MS.
The contents of each URG (XAD-2/PUF/filter) were Soxhlet extracted for 16 h in 150 mL of 6% ether in hexane solution with 100 mL of surrogate solution of p-terphenyld14 at 2.0 ng/mL. After being cooled, if water was visibly present in any of the extracts, the extract was passed through a glass drying tube containing sodium sulfate. The extracts were concentrated to 2mL and quantitatively transferred to a 3.7-mL vial, and the final volume was adjusted using 10% diethyl ether in hexane.
Analysis for the neutral target analytes was performed using an Agilent 6890/5973 (or a Thermoquest MD800) GC/ MS in selected ion monitoring (SIM) mode. A 60 m × 0.25 mm i.d. DB-5MS column was used as the GC analytical column. The GC/MS instrument was scanned to monitor two or four selected ions per analyte. The base peak ion (or the second most intense peak if there was interference with the base peak) was used as the quantification ion for each compound (22). Quantification was performed using labeled PAHs as internal standards (naphthalene-d8, acenaphthened10, phenanthrene-d10, chrysene –d12, perylene-d12). The percent relative standard deviation (% RSD) of each analyte was maintained within 30% during the initial five-point standard calibration. A continuing calibration standard was processed at the beginning and end of each sequence of 15 samples. The percent difference of each analyte in the midlevel standard was generally maintained within 40% of the initial calibration value during continuing calibrations.
Phenols Extraction and Analysis. Dust samples were extracted by acidifying with 1 mL of 1:1 sulfuric acid/water (after adding 2,4,6-tribromophenol as the surrogate standard and matrix spike solutions as required), equilibrating spiked samples for 30 min at room temperature, and extracting with three portions of 18 mL of optima-grade dichloromethane (DCM) (sonicated 10 min per extraction). The three extracts were combined and evaporated under nitrogen at less than 45 °C.
The contents of each URG (quartz filter/PUF/XAD-2) were extracted 3 times with 50 mL of optima-grade DCM, 10 min shaking per extraction (after adding 2,4,6-tribromophenol as the surrogate standard and matrix spike solutions as required). After each extraction, the DCM was decanted through a glass drying tube (1.5 in. diameter, 5 in. length, HGF Scientific, Inc., Stafford, TX) containing a glass wool plug. After the last extraction, the PUF was added to the drying tube to remove any residual DCM. The extracts were concentrated to 1.0 mL under nitrogen using a N-EVAP analytical evaporator at 35-40 °C. All glassware was washed with acidified DCM (3 mL of HCl/600 mL of DCM) prior to use.
Dust and air extracts were derivatized with N,O-bis-(trimethylsilyl) trifluoroacetamide (BSTFA) at 60 °C for 60 min. Analysis was performed using an Agilent 6890/5973 GC/MS system in SIM mode. A 30 m × 0.25 mm i.d. DB- 5.625 column was used as the GC analytical column. Quantification was performed using 3,4,5-trichlorophenol as the internal standard. A continuing calibration standard was processed at the beginning and end of each sequence of 15 samples. The percent difference of each analyte in the mid-level standard was maintained within 40% of the initial calibration value during continuing calibrations. QA/QC. Extensive QA/QC measures were conducted to ensure accuracy and reliability of measurements. Of particular concern was the possibility of field and laboratory contamination with ubiquitous target compounds in plastics and other common products, so a high proportion of blank samples was included in this study.
Air. Potential sample contamination by target compounds was evaluated using both laboratory solvent and matrix (URG contents including quartz filter/PUF/XAD-2) blanks as well as field matrix blanks shipped to the laboratory with samples. Analysts were blinded to the identity of field blanks. A total of 36 neutrals and 35 phenols blank samples were analyzed along with the 120 field samples reported here. These included field blanks (n = 7), matrix blanks (n = 23), and solvent blanks (6 neutrals, 5 phenols). The nominal analyte reporting limit in this study was the analyte level in the lowest standard of the initial five-point calibration curve. When an interfering compound was present so that the presence of a target analyte at the detection limit was obscured, the reporting limit of the analyte was raised to the size of the false interfering peak. Method Reporting Limits (MRLs) are listed in Table 2 (detected analytes) and in Table S1 in Supporting Information (all analytes). Phthalates, alklyphenols, and bisphenol A were the only compounds detected in any blanks. Target analytes were reported as not detected in samples if they were present at less than the mean + 3 SD of the amount in blank samples.
Method performance was evaluated using matrix spike samples. Over the course of the sample collection, 16 phenols or 17 neutrals PUF/XAD-2 preparations were spiked with target compounds. Average recoveries ranged from 40% to 220%; data in tables and figures are qualified for any compounds with average recoveries less than 60% or greater than 150% or for compounds with highly variable recoveries (>50% of spikes outside the 60-150% recovery range). Full-scan confirmational analyses were performed on two air sample extracts to verify large quantifications of o-phenyl phenol, propoxur, and phthalates. In addition, the two air samples with highest concentrations of 2,3-dibromo-1-propanol were confirmed by full scan.
Duplicate air samples (field duplicates; n = 10) were also analyzed by both neutrals and phenols methods to characterize reproducibility. Percent differences for field duplicate samples were typically between 15 and 25%. For a few compounds, average percent differences between field duplicates were higher than 30% (carbaryl, 33%; piperonyl butoxide, 39%; pentachlorophenol, 42%; 2,3-dibromo-1- propanol, 41%). The analyte o-phenyl phenol was included as a target analyte in both analytical methods for air and for dust samples as another check of the reliability of these methods. Percent differences between measurements by the two methods averaged 31%, and the two measures were well correlated (Pearson correlation coefficient, 0.87), although the phenols method tended to report slightly lower values than the neutrals method for this compound.
Breakthrough was not specifically evaluated, however “sandwich” combinations of XAD-2 between two layers of PUF have been shown to efficiently trap semivolatile organic chemicals with vapor pressures up to 10-3 kPa (26), so we expect these target compounds to be efficiently trapped with this preparation.
Dust. Potential sample contamination by target compounds was evaluated for dust samples by running 27 neutrals and 22 phenols solvent blanks. Matrix or field blanks are not readily available for house dust samples. Certain phthalates, nonyl- and octylphenol diethoxylate, and 2-sec-butylphenol were the only target compounds detected in solvent blanks. These target analytes were reported as not detected in samples if they were present at less than the mean + 3 SD of the blank samples.
Method performance (percent recoveries) was evaluated using matrix spiked (n = 14) samples. Average recoveries ranged from 40% to 220%.
Full-scan confirmational analyses were performed on nine dust sample extracts to verify large quantifications of bendiocarb, carbaryl, chlordane, chlorpyrifos, cypermethrin, DDT, methoxychlor, permethrin, piperonyl butoxide (PBO), propoxur, phthalates, PCB congeners, andPBDE99. The 2,3- dibromo-1-propanol detects were also confirmed by full-scan GC/MS of three dust samples.
Duplicate dust samples (laboratory splits; n = 4) were also analyzed to characterize reproducibility. Average percent differences between duplicates were less than 20% with the exception of carbaryl (59%), bis(2-ethylhexyl) adipate (30%), benz[a]anthracene (39%), benz[a]pyrene (40%), and piperonyl butoxide (22%).
Data Analysis. The unadjusted descriptive statistics were calculated using the standard formulas for simple random samples. Data below the limit of detection were set equal to zero, which will cause the sample mean to be biased low.
Adjusted geometric mean concentrations and confidence intervals were calculated for target compounds after adjusting for stratified sampling. To achieve this, data and detection limits were log transformed. If there were no data below the limit of detection in a stratum, the usual within stratum arithmetic mean and standard deviation were calculated. When there were data below the limit of detection in a stratum, the normal distribution maximum likelihood estimates for the mean and standard deviation, assuming left censoring at the log detection limit were calculated. If there were no values above the detection limit within a stratum, the previous estimate does not exist.
After the within strata estimates were obtained, the adjusted means and their standard errors were calculated using the standard formulas for stratified samples (27). Since the data were sampled separately from cases and controls and participants in the first round were limited to women over 65 yr, the sample is more complex than a stratified random sample from the nine exposure cells. However, for the purposes of summarizing the data, they were assumed to have the simple stratified structure.
The 95% confidence intervals for the adjusted means were calculated using a t-distribution, with the Satterthwaite approximation to the degrees of freedom. These confidence intervals assume a normal distribution within the population. Since this assumption is probably not true for this population, the confidence intervals should be regarded as only approximate. The mean, standard error, and confidence intervals were exponentiated back to the original scale of the concentration data. It is important to realize that the estimate of the geometric mean in the original scale is consistent for the median of a log-normal distribution rather than the mean. The confidence interval in the original scale should be interpreted as a confidence interval for the median of the concentration values.
Results and Discussion
Summary Statistics. Summary data for all detected compounds are shown in Tables 2 (air) and 3 (dust), and Tables S1 and S2 in Supporting Information provide more detailed statistics and include target compounds that were not detected. Chemicals are divided into the following groups: (1) alkylphenols; (2) phthalates; (3) parabens; (4) PAHs, PCBs, and PBDEs; (5) pesticides; and (6) phenols and miscellaneous. The summary tables (Tables 2 and 3) show the number of samples tested for each detected analyte, the percent of samples with detectable levels, the method reporting limit, and the median and range for the raw data. Tables S1 and S2 in Supporting Information include additional descriptive statistics for the raw data (arithmetic mean, range of detects, and the median, 75th, and 90th percentile concentrations detected). In addition, Tables S1 and S2 (Supporting Information) present geometric means and confidence intervals for the data after (i) adjusting for stratification in the participant selection process based on self-reported and GIS-based opportunities for pesticide exposure and (ii) using maximum likelihood estimates with left censoring for nondetects. Comparison of the adjusted geometric means with the medians of the raw data show few differences, suggesting that the adjustments and parametric assumptions are in agreement with the raw results.
In all, 52 of 88 target compounds were detected in indoor air and 66 of 86 compounds were detected in house dust. The most frequently detected compounds were phthalates, which are ubiquitous in plastics, building materials, food packaging, and personal care products, and alkylphenols, which are impurities or degradation products of the alkylphenol polyethoxylates that are used in detergents and personal care products and as inert ingredients in pesticide formulations. Three phthalates were detected in air in 100% of homes, and three different phthalates were detected in dust in 100% of homes. Nonylphenol was also detected in air in 100% of homes. Other frequently detected chemicals in air and dust samples include methyl paraben, which is used in personal care products; PBDEs, which are flame retardants with properties similar to PCBs; and bisphenol A, which is a constituent of polycarbonate plastics. Pesticides detected in at least half the homes included DDT, methoxychlor, pentachlorophenol, permethrin, and the synergist piperonyl butoxide (PBO) (dust) and chlordane and pentachlorophenol (air). The disinfectant o-phenyl phenol was detected in air in 100% of homes and was detected in a majority of dust samples. The number of target chemicals detected per sample ranged from 13 to 28 for air samples (mean 19) and from 6 to 42 for dust samples (mean 26). Figures 1 and 2 show concentration distributions for the most commonly detected compounds in air and dust, grouped by chemical class; and chemicals and pesticides detected at highest concentrations are summarized in Table 4.
Most Abundant Compounds.
Phthalates. Phthalates, many of which have been characterized as EDCs due to their ability to interfere with androgen action (28, 29), were detected at the highest concentrations in both air and dust, although different phthalates dominated the two media. In indoor air, diethyl phthalate (DEP) and di-n-butyl phthalate (DBP) were present at the highest concentrations. The 90th percentile concentrations in indoor air were 1560 and 426 ng/m3 for DEP and DBP, respectively. These are the same phthalates observed to be most abundant in human urine samples reported by theCDCfor a cross-section of U.S. adults (30). In dust, diethyl hexyl phthalate (DEHP) and butyl benzyl phthalate (BBP) were the chemicals detected at the highest concentrations. The 90th percentile concentrations for these phthalates in dust were 854 and 277 µg/g dust, respectively. In addition, high concentrations of an unidentified phthalate with >7 carbon chain were detected (approximate concentration range 4-800 µg/g), and this compound interfered with detection of diisononyl phthalate.
In the absence of data, most estimates of exposure to phthalates have concluded that inhalation is not an important route of exposure (29). However, the high indoor air concentrations detected here and the correspondence between phthalates abundant in air and urine suggest that inhalation exposures may be important. While exposure estimates based on ambient air concentrations may appear to be an insignificant portion of total exposure, actual exposure by inhalation is likely to be higher than would be estimated on the basis of ambient indoor air concentrations because phthalate-containing product use may result in personal air concentrations that are much higher than ambient concentrations.
Alkylphenols. Alkylphenols, particularly 4-nonylphenol (4- NP) and its mono- and diethoxylates, were also among the most abundant compounds detected (4-NP 90th percentile in air, 230 ng/m3; NP2EO in dust, 18.9 µg/g) (see Tables 2 and 3 and Tables S1 and S2 in Supporting Information). In addition to being present at high concentrations relative to other compounds detected, 4-NP was detected in 100% of indoor air samples. These data provide the first evidence that 4-NP is an important contaminant of indoor air, although lower concentrations have been reported in outdoor air (31). This result contrasts with conclusions by others that 4-NP is not volatile and would be unlikely to be a significant air contaminant (32, 33). Nonylphenol, octylphenol, and their small ethoxylates have been identified as EDCs because of their ability to mimic estrogen action (24).
Parabens and Phenols. Several other estrogenic compounds, presumably originating from consumer products, were commonly detected in air. These include the disinfectant o-phenyl phenol (90th percentile, 440 ng/m3), 4-tert-butyl phenol (90th percentile, 43ng/m3), and methyl paraben (90th percentile, 11 ng/m3).
Pesticides. Pesticides detected at the highest concentrations include the currently used pesticide permethrin and the synergist piperonyl butoxide (PBO) in dust (Table 4). Other pesticides detected at relatively high concentrations include heptachlor, propoxur, chlordane, chlorpyrifos, and pentachlorophenol in air and methoxychlor, DDT, pentachlorophenol, chlorpyrifos, carbaryl, and propoxur in dust (Table 4, Figures 1 and 2). The 90th percentile concentrations for these pesticides ranged from 10 to 19 ng/m3 in air and from 1.7 to 17 µg/g in dust. The prevalence indoors of pesticides that have been banned or restricted for many years, such as DDT, chlordane, heptachlor, methoxychlor, dieldrin, and pentachlorophenol, suggests that degradation indoors is negligible. This observation is further supported by the abundance of DDT in dust relative to its degradation product DDE (Figure 2).
Brominated Flame Retardants. PBDEs, which are flame retardants widely used in foams and other plastics, were detected in dust samples with a concentration distribution similar to the carcinogenic PAHs, benzo[a]pyrene, and benz- [a]anthracene (Figure 2), with 90th percentile concentrations ranging from 0.7 to 4.1 µg/g dust. We targeted tetra- and pentabrominated BDEs, which originate from polyurethane foams. PCBs, which have a similar mechanism of endocrine toxicity to PBDEs, were also detected in air and dust samples but at somewhat lower concentrations (Figure 2).
Another notable finding in this study was detects of the mutagen and carcinogen 2,3-dibromo-1-propanol (34) in both dust and air samples. This chemical is described as an intermediate in the production of the flame retardant TRIS (tris(2,3-dibromo-1-propyl)phosphate), which was banned in 1977, and also as a urinary metabolite of TRIS (34). We detected it in both indoor air (9% of 85 homes with detects and a wide range of concentrations with maximum of 200 ng/m3) and house dust (6% of 88 homes with maximum of 42.8 µg/g dust).
Toxicity Data and Implications. For over 30 EDCs that we detected in indoor air and dust, including alkylphenols, PBDEs, 2,3-dibromo-1-propanol, parabens, and some phenols (e.g., bisphenol A, 4-tert-butyl phenol), our measurements are the first that we know of in these media. In some cases, these are the first we are aware of in any media. The exposure data reported here provide a basis for prioritizing EDCs for more comprehensive toxicity testing and for assessing potential risks once toxicity testing is complete. The compounds listed in Table 4, for example, provide a starting point for prioritization based on chemical concentrations, and consideration of preliminary toxicity data would suggest prioritization of additional compounds, such as the brominated flame retardants.
Comparison with Available Government Risk Evaluations. We sought to compare our detected concentrations with risk-based media concentrations that have been developed for air, and we compared our dust concentrations with residential soil risk-based concentrations, which are designed to protect a small child from toxicant exposure via soil ingestion. Of the measurements that we were able to compare with EPA risk-based concentrations (35, 36), measurements in our study exceeded risk-based concentrations in at least one home for DEHP, PCBs, DDT, chlordane, dieldrin, heptachlor, and lindane (dust and air) and for benzo[a]pyrene, benz- [a]anthracene, chlorpyrifos, dicofol, and pentachlorophenol (dust only). However, because these EPA guidelines do not consider endocrine effects, these comparisons are of limited usefulness. In addition, we were unable to locate any risk-based media concentrations for 28 of the chemicals that we detected in homes in this study, including alkylphenols, parabens, some phthalates and pesticides, and most of the phenolic compounds, so we cannot evaluate the potential health risks associated with the detected concentrations using these types of data. Given the evidence of exposure reported here for EDCs, it is important to note the limitations in available toxicity data so that further work in this area can be prioritized. Furthermore, given that we detected so many EDCs and others report that mixtures at sub-threshold concentrations act additively (4, 7), our results provide additional evidence that consideration of mixtures is important in assessing EDC exposure.
Indoor Sources. For virtually all the target compounds where comparison data are available, levels detected in indoor air are higher than those reported by others for outdoor air (9, 12, 14, 22, 37, 38), confirming that most of these chemicals originate in household products and materials. For example, one study of outdoor air in urban New York/New Jersey reported that average levels of 11 nonylphenol isomers combined were in the range of 10 ng/m3 (31), while in our study the average concentration of 4-nonylphenol was 130 ng/m3. Median outdoor concentrations ofDBPwere reported to be 18 ng/m3 in a suburban California location (18) as compared with a median indoors in our study of 210 ng/m3. While environmental regulatory programs have traditionally focused on outdoor ambient air, surface water, drinking water, and hazardous industrial processes, little attention has been paid to the home environment.
Regional Variation. Comparison of these data with other studies can provide insights about regional, demographic, and temporal patterns in exposure to these compounds. Where comparison data were available (primarily for pesticides, PCBs, PAHs, and some phthalates), levels detected in our study (on Cape Cod, MA) are similar to levels reported elsewhere—especially for air concentrations (9, 12-14, 18, 39-42). Some regional differences observed for dust levels were reported in ref 40. Briefly, dust concentrations of PAHs on Cape Cod appear lower than on Long Island, NY, but higher than in many other regions of the United States (Iowa; Seattle, WA; Los Angeles, CA); levels of PCBs in dust appear higher on Cape Cod than in Iowa and Los Angeles, CA, but similar to or lower than Seattle, WA; Detroit, MI; and Long Island NY; levels of pesticides in Cape Cod house dust appear higher than other regions for DDT, carbaryl, chlordane, methoxychlor, pentachlorophenol, and propoxur; and levels appear lower than other regions for diazinon and permethrin. For chlorpyrifos and o-phenyl phenol in dust, Cape Cod levels are higher than some regions and lower than others (40). Compared with PBDE levels in indoor dust reported from Germany (43) and the United Kingdom (44), PBDE levels reported here were 5-10 times higher. These comparisons must be interpreted with caution considering differences between studies in methods of sample collection and demographics of study populations.
Individuals with Highest Measurements. As is typical for environmental measurement data, the exposure distributions for most analytes are highly skewed. Thus, the maximum concentration detected is often much higher than even the 90th or 95th percentiles. This finding suggests that (for each analyte) a small proportion of the population (e.g., 1%) receives substantially higher exposures than the majority. Since most health-based standards are derived to protect the 90th or 95th percentile-exposed individual in a population, these standards may not be adequately protective of the highest exposed1%of the population who have exposures that are substantially higher, sometimes by orders of magnitude. For example, the maximum air concentrations for DDT and diazinon were 58 and 61 times higher than the 90th percentile concentrations, respectively. In dust samples, maximum concentrations for diazinon, chlorpyrifos, and PCB 153 were 228, 122, and 89 times higher than 90th percentile concentrations, respectively. The flame retardant 2,3-dibromo-1-propanol, while it was detected in fewer than 10% of the homes, was detected over a very large concentration ranges the maximum detected concentrations in both air and dust were at least 200 times higher than the MRL.
Tools for Health Studies and Source Identification. There is great interest in conducting epidemiologic studies to evaluate effects of exposures to EDCs, but limitations in exposure assessment tools have impeded progress. Our study was designed in part to develop improved exposure tools for EDCs and to address some key data gaps—for example, these data provide a basis for prioritizing the development of exposure biomarkers. Data on key sources of these compounds and factors that affect exposure levels allow for further development of exposure assessment and source reduction tools and provide insight into exposure characterizations in health studies that have already been completed.
FIGURE 1. Cumulative frequency distributions of measured levels of the most frequently detected chemicals in indoor air samples from 120 homes. Distributions are truncated at the reporting level, and concentrations are shown on a log scale on the x-axis. Footnotes for specific chemicals refer to notes in Table 2. Chemicals are grouped into classes: (a) PAHs, PCBs, and misc.; (b) pesticides; (c) alkylphenols; and (d) phthalates.
Concentration In Indoor Air (ng/m3)
Concentration In Indoor Air (ng/m3)
Concentration In Indoor Air (ng/m3)
Concentration In Indoor Air (ng/m3)
FIGURE 2. Cumulative frequency distributions of measured levels of frequently detected chemicals in indoor dust samples from 120 homes. Distributions are truncated at the reporting level, and concentrations are shown on a log scale on the x-axis. Footnotes for specific chemicals refer to notes in Table 3. Chemicals are grouped into classes: (a) PAHs, PCBs, PBDEs, and misc.; (b) pesticides; (c) alkylphenols; and (d) phthalates.
Concentration (micrograms/g dust)
Concentration (micrograms/g dust)
Concentration (micrograms/g dust)
Concentration (micrograms/g dust)
TABLE 2. Summary Data for Detected Chemicals in Indoor Air (ng/m3)a
no. of homes % chemical sampled MRLb >RL min median max ___________________________________________________________________________________ Alkylphenols and Alkylphenol Ethoxylates 4-nonylphenol 120 3 100 21 110 420 nonylphenol monoethoxylate 120 6 95 <RL 17 73 nonylphenol diethoxylate 120 4 33 <RL <RL 26 nonylphenol ethoxycarboxylate 30 18 7 <RL <RL 18 octylphenol monoethoxylate 120 10 93 <RL 8.6 50 octylphenol diethoxylate 120 8 5 <RL <RL 120 Phthalates diethyl phthalatec 120 75 100 130 590 4300 di-n-butyl phthalated 120 21 100 52 220 1100 benzyl butyl phthalate 120 31 44 <RL <RL 480 bis(2-ethylhexyl) phthalate 102 59 68 <RL 77 1000 dicyclohexyl phthalate 102 2 21 <RL <RL 280 bis(2-ethylhexyl) adipate 120 3 99 <RL 9.0 66 di-n-propyl phthalate 120 3 15 <RL <RL 27 diisobutyl phthalate 120 2 100 11 61 990 Parabens butyl paraben 120 4 8 <RL <RL 3.2 ethyl paraben 120 1 3 <RL <RL 4.0 methyl paraben 120 1 67 <RL 2.9 21 Polycyclic Aromatic Hydrocarbons (PAHs) anthracene 90 1 1 <RL <RL 3.7 pyrene 90 1 27 <RL <RL 3.4 Polychlorinated Biphenyls (PCBs) and Polychlorinated Diphenyl Ethers (PBDEs) PCB 52 120 1 31 <RL <RL 25 PCB 105 116 1 3 <RL <RL 3.6 PCB 153 119 1 6 <RL <RL 6.7 Pesticides 4,4'-DDD 90 1 3 <RL <RL 3.5 4,4'-DDE 90 1 2 <RL <RL 5.1 4,4'-DDT 90 1 10 <RL <RL 30 bendiocarb 90 6 4 <RL <RL 120 carbaryl 120 2 11 <RL <RL 22 α-chlordane 120 1 51 <RL 0.10 61 g-chlordane 120 1 53 <RL 0.22 83 chlorothalonil 90 1 17 <RL <RL 36 chlorpyrifos 120 1 38 <RL <RL 92 3,5,6-trichloro-2-pyridinold 120 1 13 <RL <RL 7.3 diazinon 120 1 40 <RL <RL 550 dieldrin 90 2 4 <RL <RL 3.0 heptachlor 120 1 44 <RL <RL 71 lindane 90 2 1 <RL <RL 110 methyl parathiond 90 2 6 <RL <RL 92 pentachlorophenold 120 1 58 <RL 1.6 34 cis-permethrin 120 1 3 <RL <RL 3.7 trans-permethrin 120 2 3 <RL <RL 5.4 o-phenylphenol 120 1 100 12 71 970 (neutrals method) o-phenylphenol 120 1 100 9.8 70 590 (phenols method) piperonyl butoxide 90 1 6 <RL <RL 110 prometon 90 2 1 <RL <RL 4.3 propoxure 120 4 47 <RL <RL 110 trifluralind,f 90 1 10 <RL <RL 23 Phenols and Miscellaneous 2,3-dibromo-1-propanol 85 1 9 <RL <RL 200 2,4-dihydroxybenzophenoned 85 1 1 <RL <RL 1.2 4,4'-methylenediphenold 120 1 3 <RL <RL 4.9 4-tert-butylphenol 120 1 100 3.4 16 290 p-phenylphenol 120 1 1 <RL <RL 1.5 2,4-dichlorophenol 120 1 28 <RL <RL 6.0 4-nitrophenol 120 1 17 <RL <RL 7.0 a Additional summary statistics in Table S1 in Supporting Information. b MRL is the typical method reporting limit (RL) reported as median reporting limit for nondetect samples. Some samples had higher or lower RLs due to smaller or larger sample sizes, respectively, or due to interferences. For chemicals with detects in all samples, MRL is derived from matrix blank samples and assumes typical sample size (11.6 m3). For chemicals detected in blanks, MRL is the mean + 3 SD of the levels in matrix blanks and assumes typical sample size. c Average of matrix spike recoveries was high (150-220%). d Matrix spike recoveries were variable (>50% of spikes outside the range of 60-150%). e Interference from XAD-2 breakdown affects propoxur identification and quantification. f Average of matrix spike recoveries was low (40-60%).
TABLE 3. Summary Statistics for Household Dust Samples (µg/g)a
no. of homes % chemical sampled MRLb >RL min median max ___________________________________________________________________________________ Alkylphenols and Alkylphenol Ethoxylates 4-nonylphenol 118 1 80 <RL 2.58 8.68 nonylphenol monoethoxylate 118 2 86 <RL 3.36 15.6 nonylphenol diethoxylate 118 2 86 <RL 5.33 49.3 nonylphenol ethoxycarboxylate 30 3 93 <RL 2.12 9.45 4-octylphenol 118 0.2 2 <RL <RL 0.090 octylphenol monoethoxylate 118 0.5 50 <RL 0.13 1.99 octylphenol diethoxylate 118 0.2 69 <RL 0.306 2.12 Phthalates diethyl phthalate 119 4 89 <RL 4.98 111 di-n-butyl phthalate 119 24 98 <RL 20.1 352 benzyl butyl phthalatee 119 3 100 3.87 45.4 1310 bis(2-ethylhexyl) phthalatee 101 8 100 16.7 340 7700 dicyclohexyl phthalate 101 0.8 77 <RL 1.88 62.7 bis(2-ethylhexyl) adipatec,d 119 0.4 100 0.935 5.97 391 di-n-hexyl phthalate 119 0.1 76 <RL 1.1 30.6 diisobutyl phthalate 119 1 95 <RL 1.91 39.1 Parabens butyl paraben 118 0.2 22 <RL <RL 3.92 methyl paraben 118 0.3 90 <RL 0.978 8.24 ethyl paraben 118 0.2 9 <RL <RL 2.18 Polycyclic Aromatic Hydrocarbons (PAHs) anthracene 89 0.2 13 <RL <RL 3.05 benz[a]anthracene 119 0.3 76 <RL 0.499 10.0 pyrene 89 1.2 96 <RL 1.33 39.8 benzo[a]pyrene 119 0.4 85 <RL 0.712 18.1 Polychlorinated Biphenyls (PCBs) and Polybrominated Diphenyl Ethers (PBDEs) PCB 52 119 0.2 8 <RL <RL 15.7 PCB 105 119 0.2 9 <RL <RL 16.5 PCB 153 119 0.2 16 <RL <RL 35.3 PBDE 47 89 0.4 45 <RL <RL 9.86 PBDE 99 89 0.4 55 <RL 0.304 22.5 PBDE 100 89 0.3 20 <RL <RL 3.40 Pesticides 4,4'-DDD 119 0.2 9 <RL <RL 0.718 4,4'-DDE 119 0.2 13 <RL <RL 0.738 4,4'-DDT 119 0.3 65 <RL 0.279 9.61 alachlor 119 0.3 1 <RL <RL 0.221 bendiocarbc,d 114 0.2 12 <RL <RL 40.7 carbarylc,d 119 0.4 43 <RL <RL 34.4 α-chlordane 119 0.3 39 <RL <RL 9.97 y-chlordane 119 0.3 41 <RL <RL 10.6 chlorothalonil 119 0.2 19 <RL <RL 3.20 chlorpyrifos 119 0.2 18 <RL <RL 228 3,5,6-trichloro-2-pyridinol 118 0.2 31 <RL <RL 44.7 cypermethrinc 119 1 5 <RL <RL 172 diazinon 119 0.2 14 <RL <RL 51.0 dicofol (ketone form) 119 0.4 6 <RL <RL 3.54 dieldrin 119 0.4 12 <RL <RL 4.89 lindane 119 0.4 2 <RL <RL 1.04 heptachlor 119 0.2 3 <RL <RL 0.549 malathion 119 0.2 3 <RL <RL 1.48 methoxychlor 119 0.5 54 <RL 0.240 12.9 pentachlorophenol 118 0.3 86 <RL 0.793 7.96 methyl parathion 119 0.3 3 <RL <RL 0.992 cis-permethrin 119 0.3 45 <RL <RL 61.9 trans-permethrin 119 0.4 53 <RL 0.387 98.0 o-phenylphenol (neutrals method)119 0.4 67 <RL 0.283 1.67 o-phenylphenol (phenols method) 118 0.3 73 <RL 0.303 2.40 piperonyl butoxided 119 0.2 66 <RL 0.426 624 prometon 119 0.3 1 <RL <RL 0.095 propoxurc 119 0.2 42 <RL <RL 12.6 Phenols and Miscellaneous 2,3-dibromo-1-propanol 88 0.2 6 <RL <RL 42.8 2,4-dihydroxybenzophenone 88 0.7 63 <RL 0.515 9.36 3-biphenylol 118 0.2 2 <RL <RL 0.170 4,4'-biphenyldiold 118 0.3 6 <RL <RL 3.89 4,4'-methylenediphenol 118 0.2 7 <RL <RL 0.934 4-cumylphenol 118 0.2 3 <RL <RL 0.542 4-tert-butylphenol 118 0.2 5 <RL <RL 1.12 bisphenol Ad 118 0.2 86 <RL 0.821 17.6 p-phenylphenol 118 0.2 5 <RL <RL 2.40 2,4-dichlorophenol 118 0.2 5 <RL <RL 0.227 4-nitrophenolc 118 0.4 42 <RL <RL 4.25 a Additional summary statistics in Table S2 in Supporting Information. b MRL is the typical method reporting limit (RL) reported as median reporting limit for nondetect samples. Some samples had higher or lower RLs due to smaller or larger sample sizes, respectively, or due to interferences. For chemicals with detects in all samples, MRL is derived from solvent blank samples and assumes typical sample size (0.38 g). For chemicals with detects in solvent blanks, MRL is the mean + 3 SD of the levels in blanks and assumes typical sample size. c Average of matrix spike recoveries was high (150-220%). d Matrix spike recoveries were variable (>50% of spikes outside the range of 60-150%). e Spike recovery not determined.
TABLE 4. Most Abundant Chemicals
Ten Chemicals with Highest 90th Percentile Concentrations air (ng/m3)a dust (µg/g)a diethyl phthalate (1,600) 100 bis(2-ethylhexyl)phthalate (854) 100 o-phenylphenol (440) 100 benzyl butyl phthalate (277) 100 di-n-butyl phthalate (430) 100 di-n-butyl phthalate (43.9) 98 4-nonylphenol (230) 100 nonylphenol diethoxylate (18.9) 86 bis(2-ethylhexyl) phthalate (210) 68 bis(2-ethylhexyl) adipate (16.6) 100 diisobutyl phthalate (150) 100 trans-permethrin (16.5) 53 benzyl butyl phthalate (68) 44 piperonyl butoxide (15.1) 66 4-tert-butylphenol (43) 100 diethyl phthalate (10.8) 89 nonylphenol monoethoxylate (41) 95 nonylphenol monoethoxylate (8.55) 86 bis(2-ethylhexyl) adipate (22) 99 cis-permethrin (7.04) 45 10 Pesticides with Highest 90th Percentile Concentrations air (ng/m3)a dust (µg/g)a o-phenylphenol (440) 100 trans-permethrin (16.5) 53 heptachlorb (19) 44 piperonyl butoxide (15.1) 66 propoxur (16) 49 cis-permethrin (7.04) 45 y-chlordaneb (12) 53 methoxychlorb (3.38) 54 chlorpyrifos (12) 38 4,4'-DDTb(3.19) 65 pentachlorophenolb (10) 58 pentachlorophenolb (2.42) 86 diazinon (9.0) 40 chlorpyrifosb (1.87) 18 α-chlordaneb (8.8) 51 carbaryl (1.72) 43 chlorothalonil (3.4) 17 propoxur (1.70) 42 3,5,6-trichloro-2-pyridinol (1.1) 13 bendiocarb (1.11) 12 a Percent detection in italics. b Indicates banned or restricted-use pesticide (at time of sample collection).
The authors thank the following individuals for their substantial contributions to this effort: Nancy Ho, Jennifer Roberts Kachajian, Patricia Pajaron, and Christopher Swartz for coordinating and implementing the sample and field data collection; Wen Ye for data management and consulting on statistical issues; Jose Vallarino for customizing our sampling equipment; Alice Yau and Michelle Zuniga for chemical analytical support; and Karen Reece and Caitlin Willoughby for literature review and technical assistance in preparing the manuscript. We especially acknowledge the immeasurable contributions of Cheryl Osimo, Cape Cod outreach coordinator, and the women of Cape Cod who contributed their time and energy to participate in the sampling program. This research was funded by an appropriation of the Massachusetts Legislature administered by the Massachusetts Department of Public Health. Manuscript preparation was supported by the Boston Affiliate of the Susan Komen Breast Cancer Foundation, and the Susan S. Bailis Breast Cancer Research Fund.
Supporting Information Available
More detailed summary statistics for these data and a list of chemicals that were not detected in this study (Tables S1 and S2). This material is available free of charge via the Internet at http://pubs.acs.org.
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Received for review December 20, 2002. Revised manuscript received May 16, 2003. Accepted June 18, 2003.
The Estrogenic Activity of Phthalate Esters In Vitro
Environmental Health Perspectives v.105, n.8, Aug97
Catherine A. Harris, 1 Pirkko Henttu, 2 Malcolm G. Parker, 2 and John P. Sumpter 1
1 Department of Biology and Biochemistry, Brunel University, Uxbridge, Middlesex, United Kingdom
2 Molecular Endocrinology Laboratory, Imperial Cancer Research Fund, London, United Kingdom
A large number of phthalate esters were screened for estrogenic activity using a recombinant yeast screen. A selection of these was also tested for mitogenic effect on estrogen-responsive human breast cancer cells. A small number of the commercially available phthalates tested showed extremely weak estrogenic activity. The relative potencies of these descended in the order butyl benzyl phthalate (BBP)>dibutyl phthalate (DBP)>diisobutyl phthalate (DIBP)>diethyl phthalate (DEP)>diisononyl phthalate (DINP). Potencies ranged from approximately 1 10 6 to 5 10 7 times less than 17ß-estradiol. The phthalates that were estrogenic in the yeast screen were also mitogenic on the human breast cancer cells. Di(2-ethylhexyl) phthalate (DEHP) showed no estrogenic activity in these in vitro assays. A number of metabolites were tested, including mono-butyl phthalate, mono-benzyl phthalate, mono-ethylhexyl phthalate, mono- n -octyl phthalate; all were found to be inactive. One of the phthalates, ditridecyl phthalate (DTDP), produced inconsistent results; one sample was weakly estrogenic, whereas another, obtained from a different source, was inactive. Analysis by gel chromatography-mass spectrometry showed that the preparation exhibiting estrogenic activity contained 0.5% of the ortho -isomer of bisphenol A. It is likely that the presence of this antioxidant in the phthalate standard was responsible for the generation of a dose-response curve–which was not observed with an alternative sample that had not been supplemented with o , p ´-bisphenol A–in the yeast screen; hence, DTDP is probably not weakly estrogenic. The activities of simple mixtures of BBP, DBP, and 17ß-estradiol were assessed in the yeast screen. No synergism was observed, although the activities of the mixtures were approximately additive. In summary, a small number of phthalates are weakly estrogenic in vitro . No data has yet been published on whether these are also estrogenic in vivo ; this will require tests using different classes of vertebrates and different routes of exposure. Key words : contaminated standards, estrogenicity, MCF-7, metabolites, phthalates, recombinant yeast screen, ZR-75.Environ Health Perspect 105:802-811 (1997).
Address correspondence to C.A. Harris, Department of Biology and Biochemistry, Brunel University, Uxbridge, Middlesex, UB8 3PH, U.K.
We are very grateful to the Natural Environment Research Council for funding this work (project no. RO5761). We thank BP Chemical Ltd., EXXON Chemical Ltd., and Monsanto Europe S.A., who supplied us with commercial preparations of phthalate esters, Rob Bos and Monsanto Europe S.A. for the provision of various metabolites, Dow Europe S.A. for the synthesis and donation of o , p ´-bisphenol A, and the Ministry of Agriculture, Fisheries and Food (Fisheries Laboratory) for the analysis of chemical standards. We also thank David Cadogan of ECPI for arranging for the chemical analysis of DTDP.
Received 18 February 1997; accepted 1 May 1997.
In recent years there have been a plethora of publications discussing man-made estrogen-mimicking chemicals, the so-called xenoestrogens. Reports of declining semen quality ( 1 ) have been followed by hypotheses that this phenomenon may be linked to an increase in the exposure of humans to xenoestrogens, specifically in utero ( 2 ). Suspect chemicals originate from a variety of backgrounds, many being anthropogenic in origin, such as pesticides, detergents, and plasticizers. One of the earliest endocrine disruptors to be identified was the pesticide DDT, the effects of which are discussed by Fry and Toone ( 3 ). Other man-made chemicals have since been recognized as possessing estrogenic properties. For example, 4-nonylphenol is the degradation product of one group of nonionic surfactants, the nonylphenol polyethoxylates, and exposure to it has been demonstrated to induce estrogenic effects both in vitro ( 4 – 6 ) and in vivo ( 7 ). However, naturally occurring xenoestrogens–including phytoestrogens, such as coumestrol and genistein, and mycoestrogens, such as zearalenone–also exist; these may be found in plant food stuffs, to which humans have always been exposed ( 8 ).
Phthalates are just one of the many classes of chemicals that have been implicated as having estrogenic properties. Evidence of the estrogenic behavior of certain phthalates in vitro has previously been reported ( 9 – 11 ). Furthermore, an in vivo study has shown the adverse effects of butyl benzyl phthalate (BBP) on rat testes size and sperm production ( 12 ). A report concerning anin vivo multigenerational study investigating the reproductive toxicity of dibutyl phthalate (DBP) in Sprague-Dawley rats has recently been published. In this study, Wine et al. ( 13 ) found that a number of reproductive parameters were adversely affected by exposure to DBP administered via feed and that, critically, the second generation appeared more adversely affected than the first generation in that most of the F 1 males were infertile. The mechanisms underpinning these adverse reproductive effects are unclear presently, but one possibility is that some phthalates are estrogenic in vivo and hence disrupt normal male development.
Phthalates are essentially used as plasticizers in the production of polymeric materials such as polyvinyl chloride (PVC), imparting flexibility and workability, both during the manufacturing process and to the end product. When used in this way, they are not chemically bound to the product ( 14) and may therefore leach into the surrounding medium ( 15 ).
Table 1 – All the parent phthalate esters tested using recombinant yeast screen.
Phthalates are produced in extremely large volumes [the most widely used being di(2-ethylhexyl) phthalate (DEHP), at 400-500 thousand tons per annum in Europe alone; see Table 1] and have been identified in all environmental compartments. For example, they have been reported in water, sediment, air and biota sampled from the Gulf of Mexico ( 16 ), and river water and sewage effluent samples from the Greater Manchester area, United Kingdom ( 17 ). Food samples contaminated with phthalates have also been reported ( 18 – 21 ). The lipophilic nature of these chemicals indicates that fatty foods such as cream, cheese, and butter are most likely to be subject to contamination. Sharman et al. ( 21 ) discovered levels of up to 114 mg/kg total phthalate in cheese samples; however, the majority of samples contained 0.6-3.0 mg/kg DEHP and 4-20 mg/kg total phthalate. The authors suggested that these high levels might have arisen from environmental sources (for example, from the wrappers surrounding the cheese) rather than as a result of the diluted presence of the contaminant in the raw commodity, followed by its distillation in the fatty phase ( 21 ). Although these chemicals are no longer used in most direct contact food plastics and use in such materials has been regulated for many years based on toxicological data available and the fat content of the food concerned ( 22 ), it is possible that other sources of contamination during the manufacturing process, and from certain printing inks and adhesives used in packaging, may contribute to levels of phthalates found in more recently sampled foods ( 19 ).
The possibility of such extensively used chemicals as the phthalates having a detrimental influence on reproductive systems, of either humans or wildlife, clearly causes public concern, as is evident from the considerable media coverage of this issue. However, when phthalates are discussed, they are often mistakenly referred to as a single group of chemicals, with the assumption that they all have similar properties, for example estrogenic activity. In this paper we investigate the ability of individual phthalate esters to produce an estrogenic response in vitro and attempt to relate this factor to their occurrence as environmental contaminants, as a partial contribution to an assessment of their risk as endocrine disruptors.
Chemicals tested. 17ß-estradiol was purchased from Sigma, Poole, United Kingdom.
Thirty-five phthalates, encompassing a variety of structural and behavioral differences and including the major phthalates of commercial importance, were purchased from Greyhound Chemservice, Merseyside, United Kingdom (Table 1). These were of 97-99% purity.
For comparison, a number of commercial preparations were also obtained as gifts from companies as follows: dibutyl phthalate (DBP, 99.7% pure), diisobutyl phthalate (DIBP, 99.6% pure), diethyl phthalate (DEP, >99.7% pure), and dioctyl phthalate (DOP, 99.9% pure), from BP Chemical Ltd., Hull, United Kingdom; diisodecyl phthalate (DIDP, 99.9%) and diisononyl phthalate (DINP, 99.9%) from EXXON Chemical Ltd., Fareham, United Kingdom; ditridecyl phthalate (DTDP) from EXXON Chemical Ltd., Courbevoie, France; and butyl benzyl phthalate (BBP, >98.5%) from Monsanto Europe S.A., Louvain-la-Neuve, Belgium. Purity of these preparations is given as provided by the company.
Various phthalate metabolites were donated by R. Bos of the Department of Toxicology, University of Nijmegen, The Netherlands. These were mono-hexyl phthalate (MHP), mono-ethylhexyl phthalate (MEHP), mono-pentyl phthalate (MPP), mono- n -octyl phthalate (MnOP) and metabolites V, VI, and IX of DEHP ( 23 ). Also donated (by Monsanto Europe S.A.) were the primary metabolites of BBP, mono-butyl phthalate and mono-benzyl phthalate.
4-Nonylphenol, supplied by Schenectady International Inc. (Schenectady, NY), bisphenol A (Aldrich, Poole, U.K.), o , p ´-DDT (Greyhound Chemservice, Merseyside, U.K.), and genistein (Sigma, Poole, U.K.) were tested in the recombinant yeast screen only, in order to demonstrate the activity and potency of some known xenoestrogens.
o , p ´-Bisphenol A was supplied by Dow Europe S.A., Horgen, Switzerland, and was tested in the recombinant yeast screen to assess the possible significance of its presence as a contaminant in one of the phthalate samples and its effect on the apparent estrogenicity of that sample.
The recombinant yeast screen. All chemicals were assessed for estrogenic activity using a recombinant yeast screen. This is a cost-effective, sensitive, and specific process for detecting estrogenic activity. The screen has been described and extensively validated elsewhere [see Routledge and Sumpter ( 24 ) for full details]. Essentially, a gene for the human estrogen receptor has been integrated into the main yeast genome and is expressed in a form capable of binding to estrogen response elements and controlling the expression of the reporter gene lac-Z . Thus, on activation of the receptor, the lac-Z gene is expressed, producing the enzyme ß-galactosidase, which is secreted into the medium where it causes a color change of the chromogenic substance chlorophenol red-ß-d-galactopyranoside (CPRG) from yellow to red. The intensity of the red color can be measured by absorbance.
The screen is highly specific for estrogens; androgens, progesterones and corticosteroids are either completely inactive in the screen or very weakly active at very high concentrations ( 24 ).
Details of the preparation of medium components and yeast stocks have been published previously (24 ).
Growth medium was prepared by adding 5 ml 20% w/v glucose solution, 1.25 ml 4 mg/ml l-aspartic acid solution, 0.5 ml vitamin solution, 0.4 ml 24 mg/ml l-threonine solution, and 0.125 ml 20 mM copper (II) sulfate solution to 45 ml single strength minimal medium. The yeast culture was then prepared by seeding 50 ml growth medium with 125 µl yeast stock and incubating this overnight at 28°C on an orbital shaker. Assay medium contained 0.5 ml 10 mg/ml chlorophenol red-ß-d-galactopyranoside added to 50 ml growth medium seeded with 1 ml of the above yeast culture.
All glassware was thoroughly washed with solvent. Test chemicals were made up in ethanol to 2 10 -2 M (phthalates), 2 10 -4 M (4-nonylphenol, bisphenol A, genistein, o , p ´-DDT), or 2 10 -7M (17ß-estradiol) stock solutions and stored at 4°C.
Stock solutions were serially diluted in ethanol, and 10 µl of each dilution was transferred to a 96-well microtiter plate (Linbro/Titertek, ICN FLOW, Bucks, U.K.). This gave a final concentration of 10-3 M to 4.8 10 -7 M for the phthalates and their metabolites, 10 -5 M to 4.8 10 -9 M for other xenoestrogens, or 10 -8 M to 4.8 10 -12 M for 17ß-estradiol. Solvent controls were set up on each plate using 10-µl aliquots of ethanol. The ethanol was allowed to evaporate and 200-µl aliquots of assay medium (containing the yeast) was then added to each well. The plates were then sealed with autoclave tape, shaken for 2 min on a titer-plate shaker, and incubated at 32°C for 4-6 days in a naturally ventilated oven (WTB binder, BD-series; Jencons Scentific Ltd., Bedfordshire, U.K.). Plates were shaken on day 1 of incubation and again approximately 1 hr before taking absorbance readings (540 nm for color and 620 nm for turbidity), using a Titertek Multiskan MCC/340 plate reader (Life Sciences Int., Basingstoke, U.K.).
Mammalian cells. For comparison, the proliferative effects of all commercially available phthalates showing estrogenic activity in the recombinant yeast screen, as well as those that were negative but of major volume use, were tested using two estrogen-responsive human breast cancer cell lines, MCF-7 and ZR-75. As these cell lines are of human origin, they may be of particular relevance when considering the wide exposure of humans to the phthalates, which are ubiquitous in the environment ( 25 ) and can be found in such domestic products as vinyl flooring, children’s toys, printing inks, and cosmetics ( 26 ).
The phthalate samples used in these assays were the analytical standards as supplied by Greyhound Chemservice. Cells were cultured in phenol red-free medium containing 5% v/v charcoal dextran stripped serum (DCC). They were then plated in 6-well microtiter plates (Falcon, Becton Dickinson, Lincoln Park, NJ) into the aforementioned medium 3-4 days prior to commencing the experiment. For the MCF-7 cells, medium was replaced with treated medium containing either 0.1% vehicle solvent (ethanol) as a negative control, 10 -8 M 17ß-estradiol as a positive control, or 10 -5 M of each respective phthalate. Cells were trypsinized and counted using a Coulter Counter (Coulter Electronics, Harpenden, Herts, U.K.) on days 0, 2, 5, 8, and 12. Treatments were duplicated and the experiment was repeated twice. For the ZR75 cells, the treatments (control, 10 -8 M, 10 -10 M, and 10 -12 M 17ß-estradiol and 10 -5 M, 10 -6 M, and 10 -7 M of individual phthalates) were done in triplicate. Cells were counted at a single endpoint on day 11.
Table 1 lists the phthalate esters tested, together with their consumption figures in Europe, to give an idea of their importance relative to one another as industrial chemicals. Some phthalates generated a dose-dependent increase in ß-galactosidase production in the yeast screen, indicating weak estrogenic activity.
Figure 1 . Estrogenic activity of some known environmental estrogens in the recombinant yeast screen. 17ß-estradiol serially diluted from 10 -8 M and ethanol were used as positive and negative controls, respectively. 4-Nonylphenol, o´p´-DDT, bisphenol A, and genistein are shown as standard curves serially diluted from 10 -5 M.
In order to relate the significance of the activity of the estrogenic phthalates to that of other environmental estrogens, we assessed the response of the yeast screen to a range of environmental estrogens. The chemicals tested were bisphenol A (an antioxidant), genistein (a phytoestrogen), 4-nonylphenol (the degradation product of a surfactant), and o , p ´-DDT (a pesticide); the results are shown in Figure 1. These chemicals were tested over a concentration range of 10 -5 M to 5 10 -9 M, and were found to have potencies varying from approximately 10 4-10 5 times less than that of the main natural estrogen, 17ß-estradiol.
Figure 2 . The estrogenic activity in the yeast screen of phthalate esters at concentrations ranging from 10 -3 M to 5 10 -7 M, compared to 17ß-estradiol (serially diluted from 10 -8 M). A) Illustrates the estrogenic activity of phthalates consumed in major volumes in Europe. B) Illustrates the estrogenic activity of DEP and DTDP, which are used commercially in Europe, and DPhP, BCHP, and IHBP, which are of negligible commercial usage. C) Portrays the lack of estrogenic activity observed in the yeast screen when the cells were incubated with certain phthalates. Abbreviations: BBP, butyl benzyl phthalate; DBP, dibutyl phthalate; DIBP, diisobutyl phthalate; DEHP, bis(2-ethylhexyl) phthalate; DIDP, diisodecyl phthalate; DINP, diisononyl phthalate; DEP, diethyl phthalate; DTDP, ditridecyl phthalate; DPhP, diphenyl phthalate; BCHP, butyl cyclohexyl phthalate; IHBP, isohexylbenzyl phthalate; DHP, dihexyl phthalate; DIHP, diisohexyl phthalate; DMP, dimethyl phthalate; DUP, diundecyl phthalate.
The estrogenic activities of the major volume usage phthalates (those exceeding 20,000 ton/annum in Europe) in the yeast screen are shown in Figure 2A. Of these six major volume use phthalates, three possessed estrogenic activity (BBP, DBP, and DIBP), two did not (DEHP and DIDP), and one (DINP) behaved unreproducibly in the screen. The former three phthalates were the most active of all those tested, and the latter three are the most extensively used in industry. Two dose-response curves were produced for DINP due to the slightly unreproducible behavior of this chemical in the yeast screen. DINP ii (Fig. 2A) shows the mean response of two standard curves in which a detectable increase in ß-galactosidase production was observed. This pattern was reproduced in three separate assays, but differed in a further three in which DINP appeared completely inactive (DINP i).
The phthalates of relatively low or negligible use in Europe (29 different ones) were assessed for estrogenic activity using the yeast screen only. Relatively few of these (five in total) possessed any estrogenic activity; all others were inactive, even at the highest concentration tested (10 -3 M) (Fig. 2B, 2C). The results obtained from the five phthalates that showed estrogenic activity are illustrated in Figure 2B. Of these, only two (DEP and DTDP) are used commercially in Europe.
Table 2 – Phthalates found to give an estrogenic response in yeast screen.
All of the phthalates that showed activity were very weak estrogens. The most potent, BBP, was approximately 1 million-fold less potent than estradiol (Table 2), making it considerably less potent than other environmental estrogens such as bisphenol A, nonylphenol, and o , p ´-DDT. When chemicals are so weakly estrogenic, it is entirely feasible that it is not the chemical (in this case the phthalate) itself which is intrinsically estrogenic, but rather that an impurity in the chemical is estrogenic. Thus, before labeling a chemical as a weak estrogen, it is necessary to exclude the possibility that the chemical is contaminated with an estrogenic impurity. One way to address this issue is to test a number of samples, of different origin, of each phthalate possessing estrogenic activity. If all samples of a phthalate possess the same degree of estrogenic activity, it is likely that that particular phthalate is intrinsically active, whereas if the different samples of a phthalate possess considerably different potencies, it is then likely that the phthalate itself is not estrogenic, but that some samples contain varying proportions of one or more contaminants that are estrogenic.
To assess this possibility–that estrogenic contaminants might be present in some phthalates–commercial preparations of all the major volume usage phthalates, including DTDP and DEP, were assessed for estrogenic activity and their potencies compared to that of their respective analytical standards (data not shown). With one exception, no differences were observed; the estrogenic activities of the commercial preparations were equivalent to those of their respective analytical standards. However, contrary to the analytical standard, the commercial preparation of DTDP failed to produce a response, even when present at 10 -3 M. Both samples of DTDP were subsequently analyzed by gel chromatography-mass spectrometry (GC-MS). The analytical standard (the active sample) was found to contain 0.5% of the ortho -isomer of bisphenol A. The inactive preparation of DTDP did not contain this chemical. A sample of o , p ´-bisphenol A was then obtained and its response in the yeast screen was compared with that of the active DTDP sample. Figure 3 shows that o , p ´-bisphenol A was about 100 times more potent than DTDP. Therefore, the presence of this chemical at just 0.5% in the DTDP sample would produce a response equivalent to that seen. Thus, it is likely that this chemical ( o , p ´-bisphenol A) was responsible for the weak activity observed in this phthalate sample (see Fig. 1); hence DTDP is not estrogenic.
Figure 3 . The activity of bisphenol A (rows A and B),o , p ´-bisphenol A (rows D and E), and DTDP (rows G and H) in the recombinant yeast screen. Bisphenol A and the ortho – para isomer of this chemical were serially diluted (left to right) from 10 -5 M. DTDP was serially diluted from 10 -3 M. Rows C and F are controls (10 µl ethanol added to each of these wells).
The results shown in Figure 2A and 2B show that most of the active phthalates were unable to produce a maximal response in the yeast assay; only DTDP did so. For example, the response to BBP (the most estrogenic phthalate) reached a plateau at approximately 50% of the maximum response achieved with 17ß-estradiol. To determine whether this means that most of the phthalates are only partial estrogen agonists or whether other explanations account for the submaximal responses observed, a yeast screen containing BBP was incubated for longer than usual and the response was monitored daily. The results (Fig. 4) show that on day 4 (the usual incubation time for our yeast assays) BBP produced a shallow dose-response curve. However, by day 6, the response was greater. By day 13 the highest concentration of BBP had produced a maximal response. Note also that the dose-response curve to 17ß-estradiol moved approximately fourfold to the left between days 4 and 13 (i.e., the yeast screen became more sensitive), but the dose-response curve for BBP moved considerably further. Thus, the potency of BBP increased somewhat with time. For this reason, all the other phthalate data shown in this paper was obtained from yeast assays incubated for 6 days.
Figure 4 . Development of the butyl benzyl phthalate (BBP) standard curve over time. The BBP standard curve can be seen to be developing to an almost maximal response in this yeast assay.
To assess whether the estrogenic responses observed in the yeast assay were reproducible in other estrogen assays, active phthalates (plus the major volume use phthalates, DEHP and DIDP) were also tested for their ability to stimulate proliferation of MCF-7 and ZR-75 cells. The results from these assays (Fig. 5 and Fig. 6), which are based on human breast cancer cell lines, are mostly comparable to those obtained from the yeast screen. However, DEP and DTDP failed to induce proliferation of ZR-75 cells at 10 -5 , 10 -6, or 10 -7 M (Fig. 5B) although they had been active in the yeast screen, albeit only at higher concentrations. Using the ZR-75 cells, DINP at 10 -5 ,10 -6 , and 10 -7 M induced proliferation to a significantly greater extent than the control, which is in contrast to our findings for this chemical using the yeast screen. Growth curves for all estrogenic phthalates (i.e., those active in the yeast assay) and for DEHP and DIDP were obtained using MCF-7 cells. The results (Fig. 6) showed that BBP was considerably more mitogenic than any of the other phthalates. DTDP, DIBP, and DBP were approximately equivalent in activity, and all the other phthalates tested showed relatively little activity. All these results are consistent with those obtained using the yeast assay.
Figure 5 . The proliferation of ZR-75 cells incubated with various phthalates and controls including time = 0 (t = 0), ethanol, and 17ß-estradiol (data obtained from three separate assays). A) Cells incubated with butyl benzyl phthalate (BBP), diisobutyl phthalate (DIBP), and dibutyl phthalate (DBP). B) Cells incubated with diisononyl phthalate; (DINP), diethyl phthalate (DEP), and ditridecyl phthalate (DTDP). C) Cells incubated with bis(2-ethylhexyl) phthalate (DEHP), diisodecyl phthalate (DIDP), and dihexyl phthalate (DHP). A simple ANOVA was performed on the data, followed by the Bonferroni/Dunn test for multiple comparisons. Cell numbers significantly greater than the control are denoted by * p <0.05; ** p <0.01; # p <0.001.
Figure 6 . This figure depicts the proliferation of MCF-7 cells incubated with 10 -8 M 17ß-estradiol, 0.1% ethanol, or 10 -5 M of bis(2-ethylhexyl) phthalate (DEHP), diisodecyl phthalate (DIDP), diisobutyl phthalate (DIBP), diisononyl phthalate (DINP), butyl benzyl phthalate (BBP), dibutyl phthalate (DBP), diethyl phthalate (DEP), or ditridecyl phthalate (DTDP) over a period of 12 days.
Possible additive or synergistic effects between the most potent phthalates were investigated by incubating known concentrations of BBP, DBP, and 17ß-estradiol either individually or as simple mixtures in the yeast screen. The concentration of 17ß-estradiol used produced only a small response above background (Fig. 7), so that if additivity or synergism occurred, they could be observed within the range of the assay. Two concentrations of each of the most active phthalates (BBP and DBP) were tested alone and in combination with 17ß-estradiol. In all cases, the response obtained was very close to that expected if additivity had occurred (Fig. 7); in no case was the response significantly greater than predicted if additivity had occurred, that is, no evidence of synergism was observed.
Figure 7 . The activity observed in the yeast screen when yeast cells were incubated with single concentrations of butyl benzyl phthalate (BBP), dibutyl phthalate (DBP), and 17ß-estradiol, either individually or in simple mixtures. Abbreviations: A, 10 -11M 17ß-estradiol; B, 10 -4 M BBP; C, 10 -5 M BBP; D, 10 -4 M DBP; E, 10 -5 M DBP.
Actual absorbance represents the corrected absorbance figure (the absorbance read for the treated yeast minus that of the control). Theoretical absorbance is the corrected absorbance of the relevant individual treatments added together (the absorbance that would be expected if the two chemicals behaved in an additive manner).
The phthalate metabolites tested included 1) derivatives of the most abundant phthalate (DEHP), namely MEHP and metabolites V, VI, and IX ( 23 ); 2) MBzP and MBuP, which are primary metabolites of the most estrogenic phthalate (BBP); and 3) MHP, MnOP, and MPeP. All were serially diluted from 10 -3 M to 4.8 10 -7 M, and none showed any signs of estrogenic activity in the yeast screen (data not shown).
Table 2 summarizes the relative potency and the magnitude of the responses (compared to 17ß-estradiol) of all phthalates that were active in the yeast screen, together with their structures.
In this paper, we investigate the possible estrogenic behavior of a large number of phthalate estersin vitro . As far as we are aware, this is the first paper to address individual estrogenic potencies for such a comprehensive spectrum of this class of chemicals.
The phthalates studied are used by industry in variable amounts, the greatest of which is for DEHP, at up to 500,000 tons/annum in Western Europe. The worldwide production of another class of chemicals, the alkylphenol polyethoxylates, was 360,000 tons/annum in the late 1980s ( 27 ), which puts into perspective the large scale use of phthalate esters as industrial chemicals, as well as their potential environmental importance.
In terms of their estrogenic behavior, it seems that those phthalates requiring further scrutiny include 1) the shorter chain phthalates, namely BBP, DBP, and DIBP, which are used by industry in smaller quantities (Table 1), but are more estrogenically active; and 2) the longer chain phthalate DINP, which although extremely weakly estrogenic in vitro , is used in large quantities (up to 200,000 tons/annum in Europe). The estrogenic behavior of the phthalates in these assays compares favorably to that previously reported (9-11), where the potency of BBP (approximately 1 millionfold less potent than estradiol in the yeast screen) was similar to that reported by Soto et al. ( 10 , 11 ) in the E-SCREEN assay (3 millionfold less potent than estradiol) and the relative strengths of the phthalates reported to be estrogenic by Jobling et al. ( 9 ) correspond to that observed in the yeast screen (BBP>DBP). It must also be noted that, generally speaking, the activities of the phthalates in the recombinant yeast screen were reproduced in the mammalian assays, thus implying that these are real estrogenic effects, and not artifactual. There were occasional discrepancies between assays: DTDP and DEP were not found to be mitogenic in the ZR-75 cell line, but they had shown slight mitogenic activity in the MCF-7 assay and a positive response in the recombinant yeast screen. The yeast cells are more robust than mammalian cells and so could be exposed to higher concentrations of phthalates with no adverse effects, hence, the observation of activity at the higher concentrations applied in the yeast screen. The reasons for discrepancy between the two mammalian assays are unclear, but may be a result of the enhanced proliferation of the MCF-7 cell line in the presence of growth factors (the identity of which is not known), as compared to the ZR-75 cell line, which is more estrogen specific.
All active chemicals, however potent, are said to be active because they cause a response above the baseline. However, for all active phthalates, only a partial dose response was observed after the usual incubation time. For example, for DINP, the most used of all the active phthalates, the maximum response was just 15% of the maximum response obtained with 17ß-estradiol. A possible explanation for results such as these, which suggest partial agonistic behavior of the phthalates, is that these chemicals were not fully solubilized in the water-based medium employed in these assays. This is a situation frequently encountered when applying highly organic compounds to in vitro assays and is entirely feasible since, generally speaking, the solubility values for phthalates are lower than the concentrations used in these trials. Thus, it is plausible that some of the phthalates tested are actually more potent than they appear to be. However, it must be noted that the chemical treatments were added to the medium of the mammalian cell assays in ethanol, thus leading to greater homogeneity throughout, and still only a partial response was observed. Conversely, contamination of a chemical with an estrogenic compound can imply a weak estrogenicity of the substance in question when it is, in fact, the contaminant that is generating the observed response and the chemical under investigation is not estrogenically active. This phenomenon was detected in the case of DTDP, where the weakly estrogenic preparation was found to be contaminated with the ortho -isomer of bisphenol A. Hence, caution must be applied when labeling a chemical a weak estrogen, particularly if the chemical is not pure (which is usually the case, especially with industrial chemicals).
It has been reported that there is a relationship between the structure of a chemical and its estrogenic behavior ( 28 ). Of the total number of phthalates tested in our study, five possessed a secondary ring structure (BBP, BCHP, DPhP, IHBP, DCHP); of these, the first four were all weakly estrogenic, albeit with varying potencies. However, by no means was this the key to estrogenicity. Of those considered to be estrogenically active, there were several that possessed alkyl side-chains, and of these, a greater maximum response was obtained with DBP, DIBP, and DEP than by those with a secondary ring structure. It appeared that the majority of the active phthalates were among the lower molecular weight species, but again there were inconsistencies with this observation, with many of the lighter phthalates being inactive in the recombinant yeast screen. It is therefore difficult to deduce, from their two-dimensional structures alone, which phthalate esters will elicit estrogenic responses.
If a chemical exhibits only weak estrogenic activity in vitro , it does not necessarily follow that the effect of the same chemical will be insignificant when applied to a whole organism. Unfortunately, results of the nature obtained here cannot be directly extrapolated to an in vivo situation. It is not known at present whether any phthalates are estrogenic in vivo , and it will be necessary to test these chemicals in vivo via different routes of exposure before reaching conclusions. Although in vitro assays give us an idea of the potential strength of a chemical as a xenoestrogen, they cannot simulate changes to the chemical within an organism and differences in the systems of individual organisms, which may influence the potency and/or bioavailability of the chemical. Metabolic processes will vary greatly, depending on the route of uptake and on the characteristics of both the chemical and the organism concerned.
Another difficulty in estimating the environmental hazard posed by phthalate esters is the lack of data documenting the exposure of humans or wildlife to these chemicals. The fact that phthalates are used in a wide variety of extensively used goods is indisputable. It is also known that they can exude from these products. For example, DBP has been found to leach from dentures ( 29 ), as has DINP from milk tubing ( 30 ). Furtmann ( 31 ) has suggested that the main source of phthalates are the consumer products themselves and that there is some justification in the inference that, following dumping or incineration of these products, there are considerable phthalate emissions into the environment. The estimated total loss of phthalate esters in Western Europe has been put at 7,740 tons/annum, or approximately 1% of total consumption ( 32 ). However, the use of such data in the analysis of environmental hazard assessment for individual chemicals is problematic because the data is generalized and estimates refer to total phthalates.
By far the most frequently reported phthalate, and that found at highest concentrations in the environment, is DEHP. This is to be expected, considering its high usage and greater likelihood of persistence relative to the shorter chain phthalates. For this reason, one would also expect DIDP and DINP to be apparent in environmental samples, but reports concerning these phthalates are sparse. Other phthalates that have been regularly documented in food ( 19 , 20 ), air ( 33 , 34 ), sediments ( 31 , 35 ), and river water ( 36 , 37 ) include the lower molecular weight phthalates such as DMP, DEP, DBP, and BBP. These are less stable as plasticizers and are therefore liable to migrate from a polymer matrix, particularly when this material is subjected to elevated temperature or surrounded by a lipophilic medium. For this reason, despite lower consumption of these phthalates compared to the higher molecular weight species, it is perhaps not surprising for them to be commonly detected, albeit at very low concentrations, in environmental samples. The solubility and environmental persistence of individual phthalates is somewhat dependent upon the chain length of the phthalate concerned. [For a more detailed discussion of the behavior of phthalates in the aquatic environment, see Furtmann ( 31 )]. It must also be considered that these chemicals are not present in isolation in environmental systems. In any one system, various mixtures of toxic organic chemicals can be found. For example, a cocktail of trace organics was documented in alligator eggs in Lake Apopka, Florida ( 38 ). Phthalates themselves have been found in environmental samples alongside polychlorinated biphenyls, p , p ´-DDT, and p , p ´-DDE ( 34 ). Certain of the PCB congeners, for example 3,4,3´,4´-tetrachlorobiphenyl, have been identified as estrogen mimics ( 39), whereas p , p ´-DDT and p , p ´-DDE have both been reported to possess antiandrogenic properties ( 40 ). In addition, various combinations of phthalates have been found to be present in environmental samples ( 37 , 41 ). With the possibility that any contaminated environmental sample will contain more than one endocrine disrupting chemical, it seems necessary to investigate whether the effect of a combination of these chemicals will be additive, more than additive, or antagonistic. This issue was addressed by incubating simple combinations of 17ß-estradiol and two estrogenic phthalates (BBP and DBP) in the recombinant yeast screen. Jobling et al. ( 9 ) found DBP and BBP, in the presence of 17ß-estradiol, to have an agonistic, as opposed to antagonistic, effect on the stimulation of transcriptional activity in transfected MCF-7 cells. We demonstrate in this paper that the activity of combinations of two phthalates, DBP and BBP, at the concentrations shown (Fig. 7) are, in fact, slightly less than additive. When these chemicals were incubated in the presence of 17ß-estradiol (with BBP at a concentration that would induce a less than maximal response), the behavior of the combination was again additive rather than synergistic.
Another factor influencing the occurrence of phthalates in the environment is their potential for persisting and accumulating in organic matrices. This would be expected to be high because phthalates are hydrophobic chemicals; thus, it might be possible to predict their environmental fate pattern based on that of other man-made organic chemicals. For example, the polychlorinated biphenyls ( 42 ) and 4-nonylphenol ( 43 ) bioaccumulate in organisms that are exposed to these chemicals over a period of time, and they also biomagnify through the food chain. However, phthalates appear to be more readily metabolized than these persistent chemicals, particularly by enzymes in the gut ( 44 ) and in sewage treatment works, although their rate of degradation does appear to be influenced by the length of their side chains ( 45 , 46 ). It is not known whether the yeast strain employed in the assays shown in this paper is capable of metabolizing complex organic chemicals, although methoxychlor has shown a positive response in the recombinant yeast screen (47 ); and it has been reported that this chemical must be metabolized before it becomes estrogenically active ( 48 ), thus suggesting that the yeast strain is capable of degrading certain organic chemicals. A small number of phthalate metabolites were tested in the recombinant yeast screen, including monobutyl phthalate (the primary metabolite of DBP and DIBP) and monobenzyl phthalate (which, with monobutyl phthalate, are the primary metabolites of BBP). All metabolites tested were inactive in this assay, suggesting that it is the parent compounds which are estrogenic. This is significant in that, as previously discussed, the phthalates appear to be metabolized following oral exposure, and hence the monoesters are more likely to be the bioavailable form of phthalates.
It is conceivable that the route of exposure of an organism to phthalates is an important parameter when considering metabolism of these chemicals in vivo . It seems probable that phthalates are readily metabolized in the gut, such that oral exposure would not lead to accumulation of high concentrations of these chemicals. However, there is little data available on the metabolism of this group of chemicals following inhalation or dermal exposure. It is perhaps necessary to investigate the fate of phthalates within an organism following administration via these routes, judging by the presence of these chemicals in a wide array of contact media. In addition, uptake via the gills, hence directly into the blood system, as occurs in fish, may elicit responses that other routes of exposure would not.
In summary, a small number of commercially available phthalate esters (BBP, DIBP, DBP, DEP, DINP) are capable of acting as extremely weak estrogens in vitro . How this is relevant to the environment cannot yet be directly estimated, partly because comprehensive data concerning the environmental fate and behavior of these individual phthalates is not available and partly due to the impracticalities involved with extrapolating in vitro data to a whole animal situation. The phthalate most widely used by the plastic industry, and that reported on with greatest frequency, is DEHP. This phthalate did not show estrogenic activity in the assays employed in this paper. Laboratory biodegradation studies, particularly of the shorter chain phthalates (that is, those which are the more potent xenoestrogens), might imply that concentrations in the environment as a whole, and within an organism, would not reach values high enough to be of significant danger. Although the potential exists for the above-mentioned chemicals to generate adverse effects when present within the system of an organism, the concentrations and the conditions of exposure required to do so are unknown. Also note that this paper has investigated one mechanism of action only, that is, the ability of phthalates to act as estrogen agonists. This may be just one of many pathways that might lead to adverse reproductive effects in animals exposed to these chemicals. The results of in vivo experiments, such as those reported by Sharpe et al. ( 12 ) and Wine et al. ( 13 ), may not be due solely to the weak estrogenic activities of the particular phthalates administered, but may involve other, and possibly more important, mechanisms of action. For example, DEHP has been recognized for many years to be a reproductive toxicant ( 49 – 52 ), yet this particular phthalate demonstrated no estrogenic behavior in the assays employed in this study. It may also transpire that it is not simply a matter of the response of a parent organism to the chemical concerned, whether exposure is acute or chronic, but that any effect may not be detected until subsequent generations. This possibility has been very clearly demonstrated by Wine et al. ( 13 ), who found that adverse reproductive effects induced by DBP in Sprague-Dawley rats were most pronounced in the second generation although the mechanisms generating these responses are unknown.
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Last Update: August 26, 1997
Phthalate leaching study of the X-TEX fabric
Phthalates are a group of chemicals used in the manufacture of plastics. They often are called plasticizers. Phthalates can prolong the life span or durability of plastics and increase the flexibility of some plastics. In addition, phthalates have been used as solvents for other materials. They are used in hundreds of products, including vinyl flooring; adhesives; detergents; lubricating oils; automotive plastics; plastic clothing, such as raincoats; and personal-care products, such as soap, shampoo, hair spray, and textiles.
The problem with phthalates is that they have been found to belong to a group of chemicals which can affect the delicate balance of the hormonal system. Their effects have been long-established in animals and are also thought to be endocrine disrupters in humans as well.
The problem is that the chemicals do not chemically bond and can therefore ‘escape’ from the product and enter the environment. Because of their widespread use for over 50 years, they are now very prevalent in the environment – found in homes, rivers, groundwater etc – and have also been found in humans. In addition, rather than biodegrading, phthalates ‘bio accumulate’ because they do not break down easily. So our exposure to them is increasing as the numbers gradually build up.
Xextex the manufacturer of X-Tex, an environmental oil sorbent recycled fabric, determined it was necessary to determine the products contribution to the phthalate problem. A comparison study was performed using the X-Tex fabric and a similar polypropylene fabric. The fabrics were submitted to Test America in Tacoma, Washington, for a TCLP leachate study per EPA method 1311, and EPA method 8270. The study was to determine if phthalates might leach from the material and be released into the environment.
Laboratory results from the testing of both products revealed that neither product produced phthalate levels above the reporting limits of the test method. However analysis of the extract concentrations produced results for comparison of the two products. The results are listed below;
Discussion: The X-Tex and the polypropylene fabric was extracted 18 hours per EPA Method 1311. This test subjects the fabrics to a glacial acetic buffered solution extraction which accelerates the conditions found in a landfill. It is designed to leach out organics and metals that are found in waste materials for profiling. The results from this test confirmed that non-detectable levels were obtained that were below the methods reporting level based on the instrumental method detection limits. However the extract values obtained from this analysis the X-Tex had extracted levels of phthalates that were below the level of polypropylene as illustrated in the above graph. X-Tex gives the environmental community a stellar re-cycled product that is part of the solution and not the problem to our water systems.
Wednesday March 12, 2008
Lynnea Dally | The Bottom Line
Phthalates, a family of chemical rubber softeners used in many sex toys, might be dangerous to your health. Phthalates are generally used in a variety of things, from plastic shower curtains to blood bags and car dashboards. Typically anything that smells â€œplasticyâ€ has phthalates. While people are regularly exposed to phthalates, we interact with few of these other products in quite the same, intimate way as sex toys.
Phthalates are controversial because they are considered a â€œprobable human carcinogen,â€ by FDA. In high doses they cause cancer in rats and in low doses still cause serious problems. Exposure to the chemical interfered with rat genital development and fetal development, which produces stillborn rats. In humans, there has been some evidence that phthalates interfere with sperm production and possibly infant genital development. While human studies are not currently as comprehensive as the animal studies, safety groups such as the FDA and Greenpeace warn that these chemicals present a possible health risk to humans.
In 2004 the European Union banned phthalates in products intended for infants, and in America there are patches of legislation protecting infants against phthalates, such as Californiaâ€™s toxic toy ban bill passed by Governor Schwarzenegger this past year.
According to Greenpeace representative Bart van Opzeeland â€œItâ€™s incredible that this substance can still be used in toys for adults. The last few years weâ€™ve tested a lot of products but never before did we find such high concentrations.â€ So if kidâ€™s toys are being protected, why arenâ€™t adult toys protected? After all, kids put their toys in their mouths, and adults put their toys in more places than that. Thereâ€™s a few reasons. First of all, the sex toy industry is loosely regulated. Sex toys are legally labeled as â€œnoveltyâ€ toys, meaning that theyâ€™re intended as a gag gift and not actual use. This means thereâ€™s no government-sponsored research into whether the products are safe for human use. Thanks to the scads of embarrassed people buying these presents â€œas giftsâ€ the industry continues gets away with easy-to-use but dangerous materials.
Furthermore, few customers, interest groups, and legislators want to get labeled as a â€œhealthy sex toy advocateâ€ so few people step up. Finally, the overall problem is not well known in America. People never suspected that their dildo, vibrator, cock ring or fake vagina could have been toxic.
So what can you do about it? Check to make sure that your products are phthalate-free, especially before you purchase. Look for a label that says phthalate-free and avoid rubber jelly products that do not list their contents. Hard plastic products are probably safe and silicone, glass, metal and wood are completely phthalate-free. There are many online sex toy stores that offer phthalate-free sections of their store. Look over the toys you have now: do they jiggle? Smell like plastic? You might want to consider trading up. Get rid of your old toys in an environmentally safe way by using a sex toy recycling program.
A: The U.S. Centers for Disease Control (CDC) recently expanded their long-running annual body burden survey of U.S. citizens from two chemicals (lead and cotinine from passive tobacco smoke) to 27 classes of chemicals and heavy metals. CDC reported the results of this survey in 2001: the levels of several toxic chemicals and metals were much higher than expected. The CDC found that levels of lead in blood continue to fall, although mercury, another heavy metal that is a potent neurotoxin for fetuses, infants and children, were high. High mercury levels in women of childbearing age were of particular concern because mercury, like lead, crosses the placenta during pregnancy and can affect the brain development of the fetus. After birth, babies and toddlers remain more susceptible to mercury and lead because their brains and nervous systems are still developing. For more information on the health effects of mercury and lead, visit theWashington Toxics Coalition website.
Phthalates were also found at much higher concentrations than expected in women of childbearing age. Phthalates are found in many beauty products, such as skin lotion, shampoo and nail polish, and are added to plastics to make them more pliable, such as infant feeding bottles, soft plastic toys for children and pets and some medical devices. They can be absorbed through the skin, inhaled as fumes, ingested when children bite or suck on toys, or directly administered during medical care. Phthalates have been shown to cause organ damage and severe reproductive and developmental problems in animal studies. The high phthalate levels found in people were enough of a concern that CDC has prioritized phthalates for further investigation. See Chemical Case Studies for more information on phthalates.
The CDC also looked for and found a number of widely used organophosphate pesticides in its study. Organophosphates pass through the body relatively quickly, so this means that the people studied were recently exposed to these pesticides. Organophosphate pesticides are used mostly as insecticides, both in agriculture and in household products. Short-term exposure can have a number of serious health effects including cancer and endocrine disruption (see www.pesticideinfo.org).
A: The Natural Resources Defense Council has compiled dozens of studies on its website addressing questions about chemicals in breast milk. These studies show the results of breast milk testing which has been done for years in different parts of the world. In some cases, such as in Sweden, breast milk has been systematically tested for so many years that the impact of public policies (such as banning DDT) can be seen in the test results.
Studies of chemicals in the human body and how they affect our health continue to be conducted and released in countries around the world. One of the best sources for tracking new body burden studies is the "Our Stolen Future" website , which is operated by the authors of Our Stolen Future, Dr. John Peterson Myers, Dr. Theo Colburn and Dianne Dumanoski.
In the spring of 2000, the Center for Health and Environmental Justice collected dozens of studies from around the world documenting industrial chemicals and pesticides in blood, adipose tissue and breast milk. Pesticide Action Network has compiled these studies into a database that is searchable by chemical or country, so you can find a list of all studies that have been conducted in Denmark, or a list of countries in which studies have been conducted on DDT body burdens. While this database is not comprehensive, it provides a glimpse of the types of studies that have been conducted for many years around the world. It will be available on-line soon atwww.panna.org .
A: Some chemicals are called "persistent" because they last for a long time – in some cases decades – in the environment. Persistent organic pollutants or "POPs" also build up in the food chain, can travel around the world in global air and water currents, and are linked to serious health problems in humans and other species. Many organizations around the world are working to eliminate this class of POP chemicals. The international community recently recognized that the POP chemicals did not respect national borders, and an international treaty, the Stockholm Convention, was developed. The Convention calls for global elimination of an initial list of 12 POP chemicals, with more to be added once the treaty takes effect. For more information about the Stockholm Convention and the many groups working to eliminate POPs, visit the International POPs Elimination Network website.
Dr. Sandra Steingraber’s recent book, Having Faith: An Ecologist’s Journey to Motherhood provides a more personal perspective on the issue. Dr. Steingraber chronicles her pregnancy month by month, at each stage examining the potential effects of environmental contaminants on the growing fetus, including persistent chemicals that she carries in her body. She also describes the birth and breastfeeding of her daughter Faith – presenting both the wonder of the process and how those wonders are being threatened and diminished by pollutants. Dr. Steingraber’s website has excerpts from the book, information about chemicals and extensive links to organizations and information sources.
A: Some chemicals break down relatively quickly in the body, so the fact that they have been found in the body’s blood and urine means that the people tested were recently exposed to these chemicals. Chemicals that pass quickly through our bodies can still have damaging long-term effects. In fact, the body’s process of eliminating foreign compounds often makes them more reactive, and these reactive molecules can damage delicate proteins including DNA.
When a fetus is exposed to chemicals at particular stages of development, there are often serious, irreversible effects. You can learn more about the health effects of pesticides at www.pesticideinfo.org, and about some of the links between chemical exposure and developmental effects at the "Our Stolen Future" web site and in the reports Generations at Risk and In Harms Way.
A: While it would be impossible to find out exactly which chemicals you are exposed to in your house and neighborhood, there are some "right-to-know" resources that provide basic information on industrial chemical releases and pesticide applications in your state or neighborhood.
It is important to note that these resources do not capture the chemicals you are exposed to through the everyday use of many household products, pesticide residues on food, industrial by-products, and persistent pollutants that are pervasive in our environment.
www.checnet.org is a gateway to practical and accurate information for parents on how to prevent their children from being exposed to hazards in their homes.
Schettler, Ted and others. In Harm’s Way: Toxic Threats to Child Development (Greater Boston Physicians for Social Responsibility) 2000. http://www.igc.org/psr/
Solomon, Gina and Ted Schettler Generations at Risk: Reproductive Health and the Environment (MIT Press) July 1999. http://www.igc.org/psr/
Steingraber, Sandra Having Faith: An Ecologist’s Journey to Motherhood (Perseus Publishing, Cambridge, Massachusetts) 2001. www.steingraber.com
Trade Secrets home page –http://www.pbs.org/tradesecrets/problem/bodyburden.html
McDonough, William and Micheal Braungart. Cradle to Cradle/Remaking the Way We Make Things. (North Point Press) 2002.
Thornton, Joe, Michael McCally, and Jeff Howard, "Body burdens of industrial chemicals in the general population." In Life Support: The Environment and Human Health, ed. Michael McCally (MIT Press) 2002, 163-200.