Environ. Sci. Technol. 2004, 38, 2389-2396
Environmental Fate of Bisphenol A and Its Biological Metabolites in River Water and Their Xeno-estrogenic Activity T O S H I N A R I S U Z U K I , * ,† YOSHIO NAKAGAWA,‡ ICHIRO TAKANO,‡ KUMIKO YAGUCHI,‡ AND KAZUO YASUDA† Tama Branch Laboratory, Tokyo Metropolitan Institute of Public Health, 3-16-25, Shibazaki-cho, Tachikawa-shi, Tokyo 190-0023, Japan, and Tokyo Metropolitan Institute of Public Health, 3-24-1, Hyakunin-cho, Shinjuku-ku, Tokyo 169-0073, Japan
Monitoring of bisphenol A [BPA; 2,2-bis(4-hydroxyphenyl)propane] and its biological metabolites [4,4′-dihydroxy-Rmethylstilbene (DHMS), 2,2-bis(4-hydroxyphenyl)-1propanol (BPA-OH), 2,2-bis(4-hydroxyphenyl)propanoic acid (BPA-COOH), and 2-(3,4-dihydroxyphenyl)-2-(4hydroxyphenyl)propane (3-OH-BPA)] in river waters was performed by solid-phase extraction and GC/MS determination. The concentrations of BPA, BPA-COOH, BPAOH, and 3-OH-BPA in the river water ranged from 2 to 230 (8.8 × 10-12 to 1.0 × 10-9 M), from 5 to 75 (1.9 × 10-11 to 2.9 × 10-10 M), from 3 to 16 (1.2 × 10-11 to 6.6 × 10-11 M), and from 3 to 11 (1.2 × 10-11 to 4.5 × 10-11 M) ng L-1, respectively. DHMS, an intermediate in the main degradation pathway of BPA, was not detected in any water sample. Under the aerobic conditions in the river water, BPA disappeared within 8 d of incubation, but BPACOOH, BPA-OH, and tetraol remained in the supernatant after 14 d of incubation. For the xeno-estrogenic activity of BPA and the metabolites, their ability to bind to recombinant human estrogen receptor R in competition with fluorescence-labeled 17β-estradiol was measured. Fifty percent inhibitory concentrations (IC50) of BPA, DHMS, 3-OH-BPA, and BPA-OH were approximately 1 × 10-5, 1 × 10-6, 3 × 10-5, and 1 × 10-2 M, respectively. In human cultured MCF-7 breast cancer cells, BPA increased cell numbers in a dose-dependent manner at concentrations from 10-7 to 10-5 M. For the BPA metabolites, DHMS, 3-OH-BPA, and BPA-COOH caused the cells proliferation at concentrations from 10-9 to 10-6, from 10-7 to 10-6, and from 10-5 to 10-4 M, respectively. BPA-OH did not cause MCF-7 cells proliferation. These results indicate that BPA is mainly metabolized through oxidative rearrangement by bacteria in the river water, and intermediate bisphenols via minor metabolic pathways exist in river water. The presence of the bisphenols having the xeno-estrogenic effect suggests the necessity of monitoring those in river water, in the effluent waters from sewage plants, or in landfill leachate. * Corresponding author phone: +81 42 524 8749; fax: +81 42 524 5307; e-mail:
[email protected]. † Tama Branch Laboratory, Tokyo Metropolitan Institute of Public Health. ‡ Tokyo Metropolitan Institute of Public Health. 10.1021/es030576z CCC: $27.50 Published on Web 03/18/2004
2004 American Chemical Society
Introduction Bisphenol A [2,2-bis(4-hydroxyphenyl)propane or BPA; compound 1 in Figure 1] is an important monomer in the manufacture of chemical products, including epoxy and polycarbonate plastics and flame retardants. The epoxies are used as food-contact surface coatings for cans, automobile parts, and adhesives and as a coating for PVC water pipe walls. Polycarbonate plastics are used for the encapsulation of electrical parts in automotive parts, high-impact windows, household appliances, food packaging, and plastic bottles for water (1, 2). In Japan, the production of BPA amounted to 355 000 t in 1999 (3). Although the water solubility of BPA, 120-300 mg L-1 at 25 °C, is not high, the primary sources of BPA release to environmental water are expected to be effluents from facilities that manufacture epoxy and polycarbonate plastics and elution from the products containing BPA (2). In previous studies, BPA has been detected at the maximum concentration of 17.2 mg L-1 in hazardous waste landfill leachates (4), 0.71 µg L-1 in environmental waters in Japan (5), and 12 µg L-1 in water samples from streams in the United States (6). However, BPA concentrations in natural waters in Europe are mostly undetectable (7). Under laboratory conditions, biological degradation of BPA has been investigated by isolated bacteria (8-12), in environmental waters (13-15), and with activated sludge from sewage plants (14, 16). BPA-degrading bacterium, strain MV1 (a Gram-negative aerobic Bacillus), underwent 60% of the total carbon of BPA to CO2, 20% to the bacterial cells, and 20% to soluble organic compounds (8, 9); the proposed degradation pathways of BPA are shown in Figure 1. The other bacterium that could mineralize BPA [Pseudomonas sp. and Pseudomonas putida strain (10), Sphingomonas paucimobilis (11), and Pseudomanas paucimobilis FJ-4 (14)] were isolated from activated sludge in sewage plants or natural river waters. The metabolic pathways of BPA in these bacteria are the same as those of the strain MV1. From these results, the BPA-degrading bacteria seem to distribute widely in environmental water. BPA is not expected to be persistent in surface waters because the half-life of BPA was calculated to be less than or equal to 4 d in environmental waters (1315) under the laboratory experiments. Despite this, a study for environmental fate based on the monitoring of BPA and its metabolites has not been conducted in natural surface water. Recently, considerable attention has focused on BPA as well as other phenolic compounds as endocrine-disrupting chemicals. It has been reported that oral administration of BPA elicits adverse effects in reproductive organs of rats and mice (17, 18) and that BPA also binds to estrogen receptors (19, 20). Although BPA exhibited a proliferative effect on MCF-7 human breast cancer cells, which are estrogenresponsive (21-23), the estrogenic activity of BPA in in vitro and in vitro assays is extremely low, at least 100 times less than that of 17β-estradiol. As for the BPA metabolites, BPA glucronide (which is found in rodent urine, liver microsomes, and hepatocytes) did not interact with the estrogen receptors and did not affect estrogen activity in MCF-7 cells (24, 25). Estrogen activity of 2-(3,4-dihydroxyphenyl)-2-(4-hydroxyphenyl)propane (3-OH-BPA), a minor metabolite in liver microsomes of human and rats and rat hepatocytes was approximately 10 times less than that of BPA in MCF-7 cells (23) and in the coupled microsomal metabolism-yeast estrogenicity assay (26). On the other hand, there have been a few studies on the estrogenic activity of BPA metabolites produced by bacteria. Ike et al. reported that hydrxyacetophenone (HAP), except for p-hydroxybenzaldehyde (HBAL) VOL. 38, NO. 8, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2389
FIGURE 1. Metabolic pathways for BPA by bacterium, strain MV1, in human and rodents. and p-hydroxybenzoic acid (HBA), has a slight estrogenic activity on the yeast two-hybrid assay (12). Some of the urban river waters at downstream sites were treated and then supplied to people that live around there as tap water. Aquatic organisms properly inhabit the river. Therefore, it is necessary to examine the estrogenic activity of the bacterial BPA metabolites having bisphenol structure. In the present study, we first investigated the concentrations of BPA and its biological metabolites in river water and BPA degradation under laboratory conditions, and then assessed estrogenic activities of the bisphenol intermediates using a competitive binding assay of recombinant human estrogen receptor R (ERR) and a proliferative assay of MCF-7 cells.
Experimental Section Synthesis of BPA Metabolites. 4,4′-Dihydroxy-R-methylstilbene (DHMS), 2,2-bis(4-hydroxyphenyl)-1-propanol (BPAOH), and 2,2-bis(4-hydroxyphenyl)propanoic acid (BPACOOH) were synthesized according to the methods of Spivack et al. (9), purified with silica gel column chromatography, and then recrystallized with benzene. All compounds were identified by 1H and 13C NMR. Their purities were checked by LC/UV at wavelength 260 nm and by GC/MS after trimethylsililation by bis(trimethylsilyl)trifluoroacetamide (BSTFA; Wako Pure Chemicals Co., Osaka, Japan). BPA, 3-OHBPA, HBAL, HAP, and HBA were purchased from Wako Pure Chemicals Co. The deuterated BPA (BPA-d8) was obtained from Hayashi Pure Chemicals Industries (Osaka, Japan). The purities of all compounds mentioned above were more than 97%; however, DHMS consists of 85% trans(Z)-form and 15% cis(E)-form. Standard mixtures of BPA and the metabolites were made by dichloromethane for GC/MS and by 5% CH3CN in 0.1% acetic acid for LC/MS. 2390
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 38, NO. 8, 2004
Sampling Sites of River Waters. Water samples were collected monthly from May 1999 to February 2000 and from November 2002 to March 2003 at six sites along the Tama River, which flows through the Tokyo Metropolitan area as illustrated in Figure 2. The Aki and Asa Rivers flow into the Tama River near sampling sites 5 and 6, respectively. The total population around the six sampling sites is 3 841 000, and the populations around sampling site 1 plus site 5 and around site 6 are about 6% and 18%, respectively, of the total population in 1999. Effluent from sewage plants runs into the river at sampling sites 2-4 and 6. The water samples were taken using amber glass containers (2 L), precleaned with acetone and containing 1.0 g of sodium ascorbate, and were then brought back to the laboratory. The water qualities of the river waters at upstream sites are better than those at downstream sites as shown in Table 1. Solid-Phase Extraction and Derivatization of BPA and Its Biological Metabolites. Some attention was needed to prevent contamination of BPA during sample preparation. The solid-phase extraction cartridge, PS-2 (Waters, Milford, MA), was prewashed dichloromethane, acetone, methanol, and the purified water adjusted pH to 3 (with 5 mL of each) before use. Commercially available organic solvents, free of BPA, and purified water prepared by PURIC-MX for HPLC grade (Organo, Bunkyo-ku, Tokyo) were used in order to decrease BPA in blanks. Sampling bottles and all glassware were washed with acetone before use, and the use of plastic products was avoided. BPA and its biological metabolites in 500-mL river water samples, which were spiked with 5 µL of 10 mg L-1 BPA-d8 in acetone, were extracted by passing through a PS-2 cartridge with Sep-pack concentrator (Waters) at a flow rate of 20 mL min-1 after acidifying water samples to pH 2-3 with 3% HCl. The cartridges were dehydrated by passing air through them
FIGURE 2. Sampling locations of river water on the Tama River in Tokyo. Site 1, Chofu bridge; site 2, Hino bridge; site 3, Sekido bridge; site 4, Tamagawara bridge; site 5, Higashiakikawa bridge; site 6, Takahata bridge.
TABLE 1. Water Quality of the Sampling Sites in the Tama Rivera no.b
sampling site
1 4 5 6
Chofu bridge Tamagawara bridge Higashiakikawa brige Takahata Bridge
temp DO conductivity BOD (°C) (mg L-1) pH (µS cm-1) (mg L-1) 13.0 17.5 14.7 15.8
10.4 9.2 9.8 7.9
8.0 7.7 7.7 7.3
100 303 130 245
0.5 2.0 < 0.5 2.9
a
Values are the average from April 1999 to March 2000. DO, dissolved oxygen; BOD, biochemical oxygen demand. b Numbers refer to Figure 2.
for 30 min, and then BPA and the metabolites were eluted with 5 mL of dichloromethane. The dichloromethane solution was dehydrated by passing it through a small glass column packed with sodium sulfate anhydride, which was cleaned by passing it through dichlormethane before use, and then concentrated to 0.25 mL under a stream of nitrogen. BPA and its biological metabolites in the concentrated extract were trimethylsilylated by addition of 50 µL of BSTFA at room temperature for 20 min, and then the final solution was analyzed by GC/MS. Degradation of BPA in River Water under Laboratory Conditions. The degradation study was performed according to the method of Kang and Kondo (10). The river water (200 mL) collected at site 3 in Figure 2 was transferred into a 250-mL amber flask. BPA (20 mg mL-1 in acetonitorile, CH3CN) was added into the flask at the final concentrations of 1 and 10 mg L-1. The flask was incubated under aerobic conditions by bubbling with 50 mL of air min-1 that was first passed through an active-carbon column, at 15 °C in an incubator for 14 d in the dark. An aliquot of the water samples was taken out at incubation times of 0, 2, 3, 4, 5, 6, and 14 d and analyzed by LC/MS after centrifugation at 3000 rpm for 10 min. The experiment was performed in duplicate for each spiked level. Competitive Binding Assay of BPA and Its Biological Metabolites. Competitive binding assay between 17β-estradiol and various compounds was performed by using an estrogen-R(R) competitor screening kit (Wako Pure Chemicals Co.). The kit consists of recombinant human estrogen receptor (ERR) coated on the bottom of 96-well multiwell plates and fluorescein-labeled 17β-estradiol as the competitor assay. Diethylstilbesterol (DES) or the other compounds dissolved in dimethyl sulfoxide (DMSO) were added to a reaction solution containing fluorescein-labeled 17β-estradiol, and the mixture (100 µL) was added to each well. After 2 h of incubation at room temperature, the mixture, which contained free compounds or fluorescein-labeled
17β-estradiol unbound to ERR, in wells was aspirated and exchanged with the assay solution (100 µL). The concentration of fluorescein-labeled 17β-estradiol bound to ERR of the bottom was measured in a Cytofluor 4000 fluorescence plate reader (PerSeptive Biosystems Inc., Framingham, MA) with filters set for 485 nm excitation and 535 nm emission. The results are expressed as percentages of the fluorescence values for the reaction solution without samples. MCF-7 Cell Proliferation Assay of BPA and Its Biological Metabolites. MCF-7 cells (cultured human breast cancer cells) were purchased from the American Type Culture Collection (Manassas, VA). Cells were cultured in phenol red-free RPMI-1640 medium, supplemented with 5% fetal calf serum (FCS), 15 mM HEPES, 50 U mL-1 penicillin, 50 µg mL-1 streptomycin and 10 ng mL-1 insulin at 37 °C with 5% CO2 in the air at saturated humidity; cells were routinely passed at approximately 80% confluence. Prior to initiating the experiments, cells were seeded and attached in 96-well multiwell plates at 4 × 103 per well in 0.3 mL of PRMI-1640 medium, supplemented with 5% estrogen-free FCS. After 24 h of incubation, the medium was replaced with the same volume of the above medium containing 1 nM 17β-estradiol as a positive control, BPA, and the four degradation products from 1 nM to 100 µM. Estrogen-free FCS was prepared using the dextran-charcoal procedure (32). The cells were cultured for 5 d, and cell numbers in each well were determined by using a cell proliferation assay kit (cck-8; Dojindo Laboratories Co., Kumamoto, Japan). A 30 µL of WST-8 solution, 5 mM 2-(2-methoxy-4-nitrophenyl)-3-(4-nitrophenyl)-5-(2,4-disulfophenyl)-2H-tetrazolium monosodium salt, and 0.2 mM 1-methoxy phenazium methyl sulfate dissolved in 150 mM KCl was added to wells containing 0.3 mL of medium with cells. After incubation for 60-90 min at 37 °C in a humidified 5% CO2 atmosphere, the absorbance of the well was read at 450 nm (reference wavelength at 650 nm) by a microculture plate reader (model 450, Bio-Rad Laboratories, New York). The assay, a modified tetrazolium colorimetric assay, is based on metabolic reduction of WST-8 to its corresponding formazan, and the coloration obtained is directly proportional to the cell number. The coloration was found to linear for up to 120 min after the addition of the dye-containing medium, with cell numbers more than 3 × 104. GC/MS Conditions. Analytical conditions for the trimethylsilyl (TMS) derivatives of BPA and its biological metabolites were as follows; GC model, HP-5890 series II (Hewlett-Packard, Palo Ato, CA); injector temperature, 220 °C; column head pressure, 80 kPa; carrier gas, helium; autoinjector, HP-7673 (Hewlett-Packard); sample size, 2 µL (splitless injection, purge on time for 1 min; glass wool was not inserted into the splitless insert); analytical column, HP5VOL. 38, NO. 8, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2391
TABLE 2. Monitor Ions and Retention Times of Bisphenol A and Its Metabolites on GC/MS and LC/MS GC/MSb
LC/MSc
no.a
compound
abbreviation
MW
monitor ions (m/z)
tR (min)
monitor ions (m/z)
tR (min)
1 3 3 4 5 6 7 8 13
2,2-bis(4-hydroxyphenyl)propane, bisphenol A cis-4,4′-dihydroxy-R-methylstilbene trans-4,4′-dihydroxy-R-methylstilbene 4-hydroxybenzaldehyde 4-hydroxyacetophenone 4-hydroxybenzoic acid 2,2-bis(4-hydroxyphenyl)-1-propanol 2,2-bis(4-hydroxyphenyl)propanoic acid 2-(3,4-dihydroxyphenyl)-2-(4-hydroxyphenyl)propane
BPA DHMS (cis) DHMS (trans) HBAL HAP HBA BPA-OH BPA-COOH 3-OH-BPA
228 226 226 122 136 138 244 258 244
357 370 370 357 357 445
21:53 21:58 25:22 24:18 24:49 23:40
227 225 225 121 135 137 243 213 243
21:46 23:22 20:50 4:52 4:96 3:42 6:64 8:40 17:94
a
Numbers refer to Figure 1.
b
Monitor ions and tR for trimethylsilyl derivatives; -, not investigated. c Monitor inos for negative ESI.
TABLE 3. Detection Limits and Recovery of BPA and Its Metabolites from River Water Samples
a
calibration curvec
no.a
compound
detection limitsb (µg L-1)
quantitation limits (µg L-1)
order
r2
recoveryd (%)
1 3 3 7 8 13
BPA DHMS (cis) DHMS (trans) BPA-OH BPA-COOH 3-OH-BPA
1 2 2 2 3 1
3 6 6 6 10 3
linear linear linear linear quadratic quadratic
0.9998 0.9985 0.9981 0.9997 0.9984 0.9980
114 ( 2 105 ( 8 107 ( 6 132 ( 12 102 ( 16 52 ( 5
Numbers refer to Figure 1.
b
2 µL injection, based on IUPAC methods (3δ). c 5-100 µg L-1.
MS, 0.25 mm i.d. × 30 m, film thickness 0.25 µm (J&W Scientific, Folsom, CA). The GC oven temperature was programmed as follows: held at 50 °C for 1 min, increased from 50 to 200 °C at 10 °C min-1 and from 200 to 300 °C at 6 °C min-1. MS model, Automass II mass spectrometer (JEOL, Akishima, Tokyo). ionization potential, 70 eV; ionization current, 300 µA; ion source temperature, 220 °C; temperature of transfer line between GC and MS, 250 °C. The TMS derivatives of the analytes were identified and quantified by single ion monitoring (SIM) using the base peaks listed in Table 2. LC/MS Conditions. BPA and its biological metabolites were measured under the following conditions. LC model, 2690 Separation Module (Waters); solvents, 20% CH3CN containing 0.1% acetic acid for 5 min, and then up to 70% CH3CN containing 1% acetic acid for 25 min; flow rate, 0.2 mL min-1; column, XTerra MSC18 (2.1 × 150 mm, 5 µm, Waters); column temperature, 40 °C; MS model, Micromass ZMD (Waters); mode, negative electron spray ionization (ESI-); capillary voltage, 3.18 kV; cone voltage, 30 V; source block temperature, 120 °C; desolvation temperature 250 °C. BPA and its biological metabolites were identified and quantified by SIM using the base peaks listed in Table 2.
Results and Discussion Analytical Conditions of BPA and Its Biological Metabolites in Water Samples. To analyze simultaneously BPA and its biological metabolites involving phenolic and alcoholic hydroxyl and carboxyl groups by GC/MS, BSTFA is used as a suitable silylation agent (27). However, when the removal of water in the concentrated extract by sodium sulfate anhydride is inadequate, trimethylsilylation of BPA-OH is incomplete, resulting in the appearance of the peak, which is identified as the disubstituted BPA-OH for the two phenolic hydroxyl groups by TMS group after the peak of trimethylsilylated BPA-COOH. The fragmentation of the TMS derivatives of BPA and its biological metabolites, except for DHMS, showed weak intensity of molecular ions less than 10% of the base peaks. Therefore, the monitor ions were defined as listed in Table 2. 2392
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 38, NO. 8, 2004
d
Spiked level, 50 ng L-1; mean ( SD (n ) 6).
Based on the signal-to-noise ratio, detection and quantification limits of BPA and its biological metabolites ranged from 1 to 3 and from 3 to 10 µg L-1, respectively, under the GC/MS-SIM conditions (Table 3). The correlation coefficients for the peak area ratios of trimethylsilylated analytes to surrogated compounds (BPA-d8) versus concentrations from 5 to 100 µg L-1 of the analytes were 0.9990-0.9999 for BPA, DHMS, and BPA-OH (linear order), 0.9992 for BPACOOH, and 0.9990 for 3-OH-BPA (quadratic order) as shown in Table 3. The recovery of BPA, DHMS, BPA-OH, and BPACOOH were 102-132%, although the recovery of 3-OH-BPA was approximately 50% at the spiked concentration of 50 ng L-1 (Table 3). A SIM chromatogram of the river water sample is presented in Figure 3. Each peak of the analytes on the chromatogram did not appear without TMS derivatization. Monitoring of BPA and Its Biological Metabolites in the Tama River. The annual monitoring data for BPA and its biological metabolites in the Tama River are shown in Table 4. BPA is detected at high frequencies at downstream sites 2, 3, 4, and 6, and the maximum concentration was 230 ng L-1 at site 3. On the other hand, at the upstream sites 1 and 5, detectable frequencies and concentrations of BPA are lower than those at downstream sites. The BPA metabolites BPA-OH, BPA-COOH, and 3-OH-BPA as well as BPA were detected at the downstream sites. However, DHMS, an intermediate of a main metabolic pathway in bacteria, was not detected in any water samples. Annual changes in the concentrations of BPA and the metabolites at site 3 are shown in Figure 4. The order of their concentrations was BPA > BPA-COOH > BPA-OH > 3-OH-BPA, and the concentrations in the summer were lower than those in the winter. In a previous study, behavior of the phthalic acid mono- and di-esters in this river was the same as that of BPA found in this study (28). The seasonal change in their concentrations might be due to the flow rate rather than discharge in the river water because the total amounts of BPA calculated from the concentration and the flow rate of the river water are almost same in every sampling time at site 3.
FIGURE 3. GC/MS-SIM chromatogram of the trimethysilylated extract from the river water sample. For further references, see Table 2.
TABLE 4. Monitoring of BPA and Its Metabolites in the Tama Rivera site 1 QLc
site 2
site 3
site 4
site 5
site 6
no.b
compound
(ng L-1)
dt/st
concn (ng L-1)d
dt/st
concn (ng L-1)d
dt/st
concn (ng L-1)d
dt/st
concn (ng L-1)d
dt/st
concn (ng L-1)d
dt/st
concn (ng L-1)d
1 3 3 7 8 13
BPA DHMS (cis) DHMS (trans) BPA-OH BPA-COOH 3-OH-BPA
2 3 3 3 5 2
1/7 0/7 0/7 0/7 0/7 0/7
ND-2 ND ND ND ND ND
7/7 0/7 0/7 4/7 6/7 4/7
4-230 ND ND ND-11 ND-75 ND-11
14/14 0/14 0/14 5/14 10/14 4/14
3-140 ND ND ND-11 ND-35 ND-9
7/7 0/7 0/7 2/7 4/7 3/7
3-200 ND ND ND-15 ND-30 ND-7
1/7 0/7 0/7 0/7 0/7 0/7
ND-3 ND ND ND ND ND
6/7 0/7 0/7 2/7 2/7 2/7
ND-200 ND ND ND-16 ND-13 ND-3
a ND, less than QL; dt, detection times; st, sampling times. b Numbers refere to Figure 1. c QL, quantitation limits. maximum.
FIGURE 4. Seasonal changes in BPA and its biological metabolites in river waters at sampling site 4. In parentheses, the concentrations of BPA-COOH and BPA-OH are lower than quantification limits but higher than detection limits. Tokyo Metropolitan Government has conducted monitoring of BPA in the influent and effluent waters at the six sewage plants, which were located around sampling sites 2-4 in 1999. The concentrations of BPA in the influents and the effluents ranged from 230 to 1700 and from 10 to 90 ng L-1, and the averages were 600 and 35 ng L-1. Overall reduction of BPA at the six sewage plants was estimated at more than 95%, which is the same result reported for the other sewage plants (29). The average flow rate of site 4 on the Tama River in 1999 was about 1 300 000 m3 d-1, and the sewage effluent contributes 40-50% of the river water at site 4. On the basis of these results, it seems that the principal
d
Ranges are minimum to
source of BPA in this river water is the effluents from the sewage plants. Biodegradation of BPA in River Water under Laboratory Conditions. Biodegradation of BPA by river water collected at site 3 was investigated under an aerobic system (Table 5). After a lag period of 2 or 3 d, BPA in the water samples decreased rapidly to less than 10% of the initial concentration within 6 d of incubation at both 1 and 10 mg L-1. Optical densities in all samples increased after 2 d of incubation. On the other hand, the water samples autoclaved at 120 °C for 15 min did not cause BPA degradation, and also the optical density up to 14 d of incubation was unchanged under the same conditions. These results suggest that BPA is rapidly degraded with the proliferation of bacteria in the water samples. The half-lives of BPA based on the first-order kinetics were 0.4 and 1.1 d at the initial concentrations of 1 and 10 mg L-1, respectively. The degradation rate of BPA and the distribution of the BPA-degrading bacteria in the river water are the same as in previously published reports (10-16). The main biological metabolites of BPA, HAP, BPA-OH, and BPA-COOH appeared in the supernatant after 3 d of incubation (Table 5 and Figure 5). The concentrations of HAP, BPA-OH, and BPA-COOH were less than 0.4, 3.9, and 4.2%, respectively, of the initial BPA concentration (1 mg L-1). In the study with the river water samples spiked with 100 mg of BPA L-1 under this degradation conditions, a metabolite in addition to HAP, BPA-OH, and BPA-COOH was detected in the supernatant at the retention time 3.67 min on LC/MS-ESI-. The peak is tetraol (9 in Figure 1) because its fragment ions appeared at m/z 135 (80%), 241 (40%), 259 (100%), and 260 (15%). Spivack et al. identified tetraol as a minor intermediate of BPA in the culture medium VOL. 38, NO. 8, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2393
TABLE 5. Biodegradtion of BPA in the River Water under Aerobic Conditionsa initial BPA concn (mg/L) 1
10
a
compound
unit
QLb
0
2
3
OD 600 nm BPA DHMS (cis) DHMS (trans) HBAL HAP HBA BPA-OH BPA-COOH 3-OH-BPA tetraol OD 600 nm BPA DHMS (cis) DHMS (trans) HBAL HAP HBA BPA-OH BPA-COOH 3-OH-BPA tetraol
× 10-3 mg/L µg/L µg/L µg/L µg/L µg/L µg/L µg/L µg/L area × 10-3 mg/L µg/L µg/L µg/L µg/L µg/L µg/L µg/L µg/L area
1 0.02 10 10 5 2 10 5 5 5 100 1 0.02 10 10 5 2 10 5 5 5 100
4 1.00 ND ND ND ND ND ND ND ND ND 3 10.0 ND ND ND 2 ND ND ND ND ND
3 1.00 ND ND ND ND ND ND ND ND ND 5 10.0 ND ND ND ND ND ND ND ND ND
5 0.97 ND ND ND ND ND 12 ND ND ND 15 6.72 ND 20 ND 34 ND 391 25 ND 213
ND, less than QL.
b
incubation time (d) 4 5 11 0.33 ND ND ND 4 ND 39 30 ND 390 58 3.94 ND ND ND 84 ND 679 92 ND 1030
18 0.02 ND ND ND ND ND ND 42 ND 540 28 1.75 ND ND ND 172 ND 594 177 ND 2320
6
14
19 ND ND ND ND ND ND ND 37 ND 760 36 0.78 ND ND ND 258 ND 543 178 ND 3170
8 ND ND ND ND ND ND ND 29 ND 370 18 ND ND ND ND 2 ND 525 218 ND 2940
QL, quantitation limits.
FIGURE 5. LC/MS-SIM chromatogram of the river water for BPA degradation under the laboratory conditions. For further references, see Table 2. of BPA-degrading bacteria MV-1 (9). It seems that the metabolic fate of tetraol was similar to that of BPA-COOH, although its concentration was not determined (Table 5). BPA-OH, BPA-COOH, and tetraol were observed after 14 d of incubation. These metabolites were persistent under these laboratory conditions. On the other hand, DHMS (trans) was detected only after 3 d of incubation at the spiked level of 10 mg L-1 in this conditions. These results suggest that BPAOH and BPA-COOH detected in river waters are derived from degradation of BPA at the sewage plants and in river waters. However, the cause of 3-OH-BPA appearance in the river water is unclear because it was not detected in the degradation study under this laboratory conditions. Under UV irradiation at 365 nm, BPA was converted to phenol, 2394
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 38, NO. 8, 2004
4-propylophenol, and 4-isopropylenephenol (semiquinone derivatives; 30), but 3-OH-BPA was not produced. Sajiki and Yonekubo reported that BPA was degraded to 3-OH-BPA in the presence of radical oxygen species at room temperature for 20 min (31). Further investigation is necessary to clarify the source of 3-OH-BPA in the river waters. Xeno-estrogenic Effects of BPA and Its Biological Metabolites. The ability of BPA and the metabolites to bind to ERR in competition with fluorescence-labeled 17β-estradiol was measured (Figure 6). BPA, DHMS, 3-OH-BPA, and BPAOH displaced the 17β-estradiol bound to the receptor in a competitive manner; the competitive potency of these compounds was 50 times less than that of DES. IC50 values of BPA, DHMS, 3-OH-BPA, and BPA-OH in this assay were
TABLE 6. Effects of BPA and Its Metabolites on the Growth of MCF-7 Cellsa relative proliferation of MCF-7 cells (%) at concentration (M) compound
10-9
10-8
10-7
10-6
10-5
10-4
17β-estradiol BPA DHMS (trans, cis) BPA-OH BPA-COOH 3-OH-BPA
201 ( 13* 109 ( 12 137 ( 4* 104 ( 9 108 ( 9 110 ( 15
112 ( 6 147 ( 15* 113 ( 16 111 ( 12 112 ( 6
134 ( 11* 155 ( 10* 118 ( 10 103 ( 8 121 ( 2*
176 ( 4* 131 ( 11* 110 ( 9 99 ( 8 121 ( 7*
151 ( 23* 90 ( 9 116 ( 10 122 ( 2* 62 ( 2*
33 ( 7* 3 ( 3* 95 ( 8 129 ( 4* 34 ( 3*
a
% of control, 100 ( 6. An asterisk (*) indicates significances between the control group and treated groups (p < 0.05).
because DHMS was detected under higher concentration in this degradation study. All of the BPA metabolites were also detected at the concentrations from nanograms to micrograms per liter in the influent and effluent water from the sewage plants, which were stored in the polycarbonate bottle for 24 h at room temperature. BPA has been detected at the concentration more than 10 mg L-1 in landfill leachates in Japan (4). Therefore, we think that the monitoring of the bisphenols such as DHMS and 3-OH-BPA having estrogenic activity is necessary for the river water containing a higher concentration of BPA and for treatment water from the plastics industry or landfill leachate.
Acknowledgments We thank Dr. Miyaoka (Tokyo University of Pharmacy and Life Science, Hachioji-shi, Tokyo) for advice on the synthesis of biological metabolites of BPA.
Literature Cited
FIGURE 6. Competitive binding assay of BPA and its biological metabolites to the estrogen receptor (ERr). (b), DES; (O), DHMS; (2), BPA; (4), 3-OH-BPA; (9), BPA-OH; (0), BPA-COOH. approximately 1 × 10-5, 1 × 10-6, 3 × 10-5, and 1 × 10-2 M, respectively. The effects of BPA and the metabolites on the proliferation of MCF-7 cells are presented in Table 6. The results were compared with those from an untreated control. The number of MCF-7 cells on day 6 was approximately 1.2 × 104 cells per well. The proliferation by 17β-estradiol (positive control) in this study is consistent with previous reports (32, 33). BPA increased cell numbers in a dose-dependent manner from 10-7 to 10-5 M. 3-OH-BPA at the concentration of 10-7-10-6 M caused a slight increase in cell numbers, while the proliferation rate was lower than that of BPA. These results are the same as previous results (23, 26). The proliferation potency of DHMS was the strongest of the metabolites of BPA; the compound increased significantly at the lowest concentration, 10-9 M. BPA-COOH also increased slightly in cell numbers at 10-5 and 10-4 M. BPA-OH did not affect on MCF-7 cells proliferation. BPA, DHMS, and 3-OH-BPA caused a decrease in MCF-7 cell numbers at concentrations more than 10-4, 10-4, and 10-5 M, respectively. This indicates that these compounds were cytotoxic in MCF-7 cells at higher concentrations. In previous studies, estrogenic activities of the BPA metabolites by bacteria (HBAL, BAL, and HAP) were 100 times less than that of BPA (12). Ohtani et al. reported that estrogenic activity of the supernatant from a cultured medium with isolated bacteria and BPA increased in the course of BPA degradation, and this is presumed to be due to the formation of DHMS (11). In this study, DHMS was not detected in the river water samples. It might be due to the low concentrations of BPA
(1) Staples, C. A.; Dorn, P. B.; Klecka, G. M.; O’Block, S. T.; Harris, L. R. Cemosphere 1998, 36, 2149-2173. (2) Howard, P. H. Handbook of Environmental Fate and Exposure Data for Organic Chemicals; Lewis: Chelsea, MI, 1989; Vol. I, pp 95-100. (3) Research and Statistics Department Minister’s Secretariat. Yearbook of Chemical Industries Statistics in 1999; Ministry of International Trade and Industry: Tokyo, 2000. (4) Yamamoto, T.; Yasuhara, A.; Shiraishi, H.; Nakasugi, O. Chemosphere 2001, 42, 415-418. (5) Water Quality Management Division, Environment Agency in Japan. Surveillance of Endocrine Disruptors at Public Water Area; October 1999. (6) Kolpin, D. W.; Furlong, E. T.; Meyer, M. T.; Thurman, E. M.; Zaugg, S. D.; Barber, L. B.; Buxton, H. T. Environ. Sci. Technol. 2002, 36, 1202-1211. (7) Hendriks, A. J.; Maas-Diepeveen, J. L.; Noordsij, A.; Van der Gaag, M. A. Water Res. 1994, 28, 581-598. (8) Lobos, J. H.; Leib, T. K.; Su, T. Appl. Environ. Microbiol. 1992, 58, 1823-1831. (9) Spivack, J.; Leib, T. K.; Lobos, J. H. J. Biol. Chem. 1994, 269, 7323-7329. (10) Kang, J.-H.; Kondo, F. Arch. Environ. Contam. Toxicol. 2002, 43, 265-269. (11) Ohtani, Y.; Fujinami, H.; Shimada, Y. J. Environ. Lab. Assoc. (Zenkokukankyokenkaishi, in Japanese) 2001, 26, 176-184. (12) Ike, M.; Chen, M.-Y.; Jin, C.-S.; Fujita, M. Environ. Toxicol. 2002, 17, 457-461. (13) Jin, C.-S.; Tokuhiro, K.; Ike, M.; Furukawa, K.; Fujita, M. J. Jpn. Soc. Water Environ. 1996, 19, 878-884. (14) Ike, M.; Jin, C.-S.; Fujita, M. Water Sci. Technol. 2000, 42, 31-38. (15) Klecka, G. M.; Gonsior, S. J.; West, R. J.; Goodwin, P. A.; Markham, D. A. Environ. Toxicol. Chem. 2001, 20, 2725-2735. (16) West, R. J.; Goodwin, P. A.; Klecka, G. M. Bull. Environ. Contam. Toxicol. 2001, 67, 106-112. (17) Bond, G. P.; McGinnis, P. M.; Cheever, K. L.; Harris, S. J.; Plantnick, H. B.; Neimeier, R. W. Proc. Soc. Toxicol. 1980, 19, A23. (18) Morrissey, R. E.; George, J. D.; Price, C. T.; Tyl, R. W.; Marr, M. C.; Kimmel, C. A. Fundam. Appl. Toxicol. 1987, 8, 571-582. VOL. 38, NO. 8, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
2395
(19) Krishnan, A. V.; Stathis, P.; Permuth, S. F.; Tokes, L.; Feldman, D. Endocrinology 1993, 132, 2279-2286. (20) Kuiper, G. G. J. M.; Carlsson, B.; Grandien, K.; Enmark, E.; Haggblad, J.; Nilsson, S.; Gustafsson, J.-A. Endocrinology 1997, 138, 863-870. (21) Nagel, S. C.; vom Saal, F. S.; Thayer, K. A.; Dhar, M. G.; Boechler, M.; Welshons, W. V. Environ. Health Perspect. 1997, 105, 7076. (22) Perez, P.; Pulgar, R.; Olea-Serrano, F.; Villalobs, M.; Rivas, A.; Metzler, M.; Pedraza, V.; Olea, N. Environ. Health Perspect. 1998, 106, 167-174. (23) Nakagawa, Y.; Suzuki, T. Xenobiotica 2001, 31, 113-123. (24) Snyder, R. W.; Maness, S. C.; Gaido, K. W.; Welsch, F.; Sumner, S. C.; Fennell, T. R. Toxicol. Appl. Pharmacol. 2000, 168, 225234. (25) Matthews, J. B.; Twomey, K.; Zacharewski, T. R. Chem. Res. Toxicol. 2001, 14, 149-157. (26) Elsby, R.; Maggs, J. L.; Ashby, J.; Park, B. K. J. Pharmacol. Exp. Ther. 2001, 297, 103-113.
2396
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 38, NO. 8, 2004
(27) Blau, K.; Halket, J. Handbook of Derivatizatives for Chromatography, 2nd ed.; John Wiley & Sons: New York, 1993. (28) Suzuki, T.; Yaguchi, K.; Suzuki, K.; Suga, T. Environ. Sci. Technol. 2001, 35, 3757-3763. (29) Kuribayashi, S.; Goto, M. J. Inst. Waste Eng. Technol. 2000, 42, 21-28. (30) Peltonen, K.; Zitting, A.; Koskinen, H.; Itkonen, A. Photochem. Photobiol. 1986, 43, 481-484. (31) Sajiki, J.; Yonekubo, J. Chemosphere 2003, 51, 55-62. (32) Soto, A. M.; Sonnenscheen, C.; Chung, K. L.; Fernandez, M. F.; Olea, N.; Serrano, F. O. Environ. Health Perspect. 1995, 103, 113-122. (33) Breinholt, V.; Larsen, J. C. Chem. Res. Toxicol. 1998, 11, 622629.
Received for review July 31, 2003. Revised manuscript received January 15, 2004. Accepted February 3, 2004. ES030576Z
