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Biodegradation of endocrine disrupting compounds in harbour seawater and sediments
Science of the Total Environment 407 (2009) 5713–5718
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Science of the Total Environment j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / s c i t o t e n v
Biodegradation of endocrine disrupting compounds in harbour seawater and sediments Brian J. Robinson a,b, Jocelyne Hellou a,b,c,⁎ a b c
Fisheries and Oceans Canada, Bedford Institute of Oceanography, PO Box 1006, Dartmouth, Nova Scotia, Canada B2Y 4A2 Department of Oceanography, Dalhousie University, Halifax, Nova Scotia, Canada B3H 4J1 Department of Chemistry, Dalhousie University, Halifax, Nova Scotia, Canada B3H 4J1
a r t i c l e
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Article history: Received 18 March 2009 Received in revised form 1 July 2009 Accepted 5 July 2009 Available online 7 August 2009 Keywords: Biodegradation Halifax Harbour Endocrine Disrupting Compounds Metabolites
a b s t r a c t The biodegradation of three endocrine disrupting compounds was examined using samples of seawater and sediment collected from Halifax Harbour, Nova Scotia, Canada, an urbanized harbour impacted by over two centuries of anthropogenic contamination. Flask experiments, where the samples were mixed to form a slurry were used to monitor the aerobic biodegradation of the synthetic plasticizer bisphenol A (BPA), the natural hormone 17β-estradiol (E2), and the pharmaceutical and contraceptive ethinylestradiol (EE2). Degradation rates followed the order E2 N EE2 N BPA with half-lives of up to 1, 5 and 14 days in seawater, respectively. A rapid initial degradation rate for all three compounds with no apparent lag phase indicated the ability of the microbial community to readily catabolise the chemicals. The formation of unidentified non-persistent intermediate metabolites was observed during the E2 degradation experiments. These degradation rates are more rapid and complete than reported in previous studies, indicating the adaptation of native microbial communities to these contaminants. Crown Copyright © 2009 Published by Elsevier B.V. All rights reserved.
1. Introduction The occurrence of endocrine disrupting compounds (EDC) in the aquatic environment has generated worldwide interest because these chemicals can cause feminization of fish as well as interfere with the reproduction and development of other aquatic organisms (Purdom et al., 1994; Harries et al., 1996; Larsson et al., 1999). Most EDC are released through sewage effluents, and like all organic compounds their fate in the aquatic environment is multi-faceted, with as a general rule, non-ionic compounds with an octanol–water partition coefficient (log Kow) below 3.0 expected to reside preferentially in the water phase, where they can be dispersed, diluted, degraded and sometimes photooxidized. Lipophilic contaminants with a log Kow N 3.0 will be attracted to lipid enriched particles, further transported in the aquatic environment under consideration and deposited in sediments. This view comes from the basic definition of the partition coefficient where octanol represents a substitute for lipid in organisms or of the total organic carbon (TOC) content of sediments (Mackay, 1991). When comparing the concentration of contaminants in sediments, values are typically normalised to a TOC content of 1%, as discussed in Swartz (1999). In contrast, the dissolved organic carbon (DOC) content of seawater in a near-shore turbulent location is in the order of 0.0001% (Kepkay 2000) and even ⁎ Corresponding author. Ecosystem Research Division, Bedford Institute of Oceanography, PO Box 1006, Dartmouth, NS, Canada B2Y 4A2. Tel.: +1 902 426 7451; fax: +1 902 426 6695. E-mail address: [email protected] (J. Hellou).
then with less affinity for polar organic compounds. Once discharged in the environment, organic chemicals can also be altered through chemical, photo- and bio-degradation producing derivatives with longer or shorter half-lives. For the past fifteen years, concern has been directed towards the fate of endocrine disrupting compounds (EDC). The majority of research on the biodegradation of EDC has focused on degradation during the sewage treatment process (Ternes et al., 1999; Lee and Liu, 2002). Few studies have examined partitioning (Lai et al., 2000; Bowman et al., 2002; Robinson et al., 2009) or the biodegradation of EDC in field collected marine samples (Ying and Kookana, 2003). Halifax Harbour, located in Nova Scotia, Canada, is the largest marine port in Atlantic Canada and it represents a unique location to study the fate of EDC since the harbour and its surrounding waters have been used for the disposal of human and industrial wastes for over a century. However, over the last decade the municipality invested millions of dollars in the construction of sewage treatment plants to process the 180 million liters of sewage that is released into the harbour every day. Since the construction of three STP had not been completed prior to the start of this study, it was of interest to investigate the presence of EDC for future comparison. The goal of the present research was to use seawater and sediment from Halifax Harbour, Nova Scotia, Canada, to examine the ability of local microbial communities to biodegrade EDC deriving from three distinct input sources that are discharged in combined sewage effluents. Bisphenol A (BPA) is a synthetic plasticizer commonly used in polycarbonates and has been the subject of recent interest due to links to human health effects (Cao et al., 2008). The natural compound 17β-
0048-9697/$ – see front matter. Crown Copyright © 2009 Published by Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2009.07.003
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estradiol (E2) is produced by males and females alike but in varying proportion depending on the age and life cycle of individuals. The pharmaceutical product 17α-ethinylestradiol (EE2) represents the active ingredient in the birth control pill and is detected in amounts that depend on the use of various contraceptives by a population. The potential variation in the biodegradation rates of these three EDC was examined through experiments using seawater and sediment samples collected at different times of the year near a major sewage effluent in Halifax Harbour. Biodegradation trends were compared to the levels of bacteria in field samples, as well as total bacterial growth during flask experiments. Simple first-order reaction kinetics were used to determine degradation rates and half-lives. The possible identity of metabolites formed during one of the experiments is discussed, and the importance of other modes of degradation is summarized. 2. Materials and methods 2.1. Sample collection Seawater and sediment were collected on three occasions (July 2004, November 2004 and March 2005) from Halifax Harbour, Nova Scotia, Canada at a location near a major sewage effluent (44°40.28′ N 63°35.94′ W). This location has been documented as one of the most polluted sites in the harbour for organic contaminants (Site 7 in Hellou et al. 2002; Site C in Robinson et al. 2009). Seawater samples were collected at 10–15 cm below the water surface in 4 L glass bottles and stored at 5 °C, while the top 2–3 cm of sediments were collected in an Eckman grab sampler, passed through a 2 mm sieve to remove large debris and stored in mason jars at 5 °C prior to use. Sediment subsamples were set aside from all locations for total organic carbon (TOC) analysis. The organic carbon content of the sediments was determined using a Leco WR-112 carbon analyzer after treatment with 1 M hydrochloric acid to remove inorganic carbon (Sunderland et al. 2004). 2.2. Experimental set-up The biodegradation of BPA, E2 and EE2 was measured through a series of experiments whereby seawater and sediment samples were mixed to form a slurry. A similar technique has been used in previous studies (Jurgens et al., 1999; Lai et al., 2000; Bowman et al., 2002; Jürgens et al., 2002; Ying and Kookana, 2003), although our experiments differed in that concentrations in both the aqueous and sediment phases were monitored over the course of the experiment. The biodegradation experiments were performed in 250 mL Erlenmeyer flasks wrapped with aluminum foil and incubated at 20 ± 1 °C. This temperature was chosen since it is the conventional temperature used for biodegradation experiments (OECD) making it easier to compare results between our experiments and published studies. Each flask was spiked with a standard containing 50 μg of each compound dissolved in methanol. Experiments with the July 2004 samples were conducted using individual flasks for each compound; for all remaining experiments the EDC were combined in one flask. Seawater (230 mL) and sediment (10 g wet weight) were combined in the flask and mixed with a Teflon magnetic stir-bar at a medium setting. At specified time points, subsamples were taken from each flask using a 20 mL graduated pipette to remove an aliquot of the seawater-sediment mixture while the stir-bar was still mixing. These samples were collected in pre-weighed Teflon tubes and centrifuged at 2000 rpm for 10 min. The water and sediment phases were then separated and the weight of each phase was recorded. On average, each sample contained approximately 19 mL of seawater and 1 g (wet weight) of sediment. Prior to extraction, each seawater and sediment sample was spiked with 1.25 μg of a surrogate standard to monitor the extraction efficiency. For the experiments using individual flasks per compound, one of the remaining EDC was used as a surrogate (eg. for the E2 experiment, BPA was used as a surrogate). For all remaining experiments with three EDC combined in the Erlenmeyer
flask, ethinylestradiol-3-methyl-ether was used as the surrogate. The seawater phase was extracted three times using liquid–liquid extraction with 10 mL of ethyl acetate in a separatory funnel. Sodium sulphate was used to remove any water remaining in the solvent prior to transfer to a round bottom flask. This solvent was evaporated on a rotary-evaporator, transferred to 1 mL of methanol, and stored in a 2 mL vial for analysis. The sediment phase was dried by adding 3 g of sodium sulphate and then extracted in the Teflon centrifuge tube with 3 × 10 mL of ethyl acetate using a sonic probe. The solvent was collected in a round bottom flask, evaporated on a rotary-evaporator, transferred to 1 mL of methanol and stored in a 2 mL vial for analysis. 2.3. Sterile control and reference experiments A series of sterile control and reference biodegradation experiments were conducted using E2 only. For the sterile control experiments, seawater and sediment samples from Halifax Harbour were autoclaved for 1 h at 121 °C. For the reference biodegradation experiments requiring samples from a pristine un-urbanised location, seawater and sediment were collected from Hantsport, Nova Scotia, Canada (Hellou et al., 2009). Both the sterile control and reference biodegradation experiments were performed using the methods described above. 2.4. HPLC analysis The methods for high performance liquid chromatography (HPLC) analysis are detailed elsewhere (Robinson, 2006; Robinson et al., 2009). Briefly, samples were analyzed on a Hewlett Packard (HP) 1090 HPLC equipped with an HP 1046A fluorescence detector. A C18 reversed phase column measuring 2.1 × 250 mm with a particle size of 5 μm (Vydac 201TP52) was used for separations with a 2.1 × 12 mm Vydac guard column. Flow rate was maintained at 0.2 mL/min, and the mobile phase consisted of water (A) and acetonitrile (B) with 0.1% trifluoroacetic acid, pH b 3. The column was kept at 40 °C and the sample injection volume was 3 μL. Chromatographic separation was achieved with the following gradient elution: 0 to 3 min 40% B, 3 to 12 min 40 to 75% B, 12 to 15 min 75 to 100% B, 15 to 17 min 100% B, 17 to 23 min 100 to 40% B, and 23 to 26 min 40% B. The excitation and emission wavelengths were set at 226 nm and 310 nm on the fluorescence detector. Quantification was performed using a five point external calibration curve with standards ranging from 0.1 ng/μL to 4 ng/μL. Quality control consisted of running periodical procedural blanks to ensure there was no background interference from the sample matrix. Table 1 contains a summary of compound specific retention times, calibration coefficients, percent relative standard deviations (%RSD) and average percent recoveries as determined during method development. 2.5. Bacterial analysis To confirm that the seawater and sediment at site C were impacted by sewage effluents, samples collected in July 2004 were analyzed using the most probable number (MPN) fermentation method for total and fecal coliforms (Horwitz 2000). All remaining samples collected for this study were analyzed for total bacterial numbers using a flow cytometer to count DNA-stained bacteria (Li and Dickie, Table 1 Summary of HPLC method performance including compound specific retention times, calibration coefficients, percent relative standard deviations (%RSD) and average percent recoveries. Compound
Retention time (min)
Calibration coefficient
%RSD
Average % recovery
BPA E2 EE2 EE2-3-Methyl-Ether
6.55 7.70 8.52 14.75
N0.999 N0.999 N0.999 N0.999
1.3 1.2 1.6 1.6
80 ± 12 91 ± 8 95 ± 7 101 ± 6
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2001). The preparation of samples for this procedure consisted of taking 2 mL of water and preserving the samples with 1% paraformaldehyde for 10 min at room temperature and then storing them at − 80 °C in cryogenic vials. Analysis was performed using a FACSort instrument equipped with a 488 nm argon laser. 3. Results and discussion 3.1. Biodegradation results and kinetics In all experiments, E2 was the most readily degraded of the three compounds with a rapid decrease in the observed aqueous and sediment phase concentrations over the first 48 h of experiments conducted at three points in time (Fig. 1). E2 was degraded to nondetectable levels after 14 days. BPA and EE2 were more resistant to biodegradation, with detectable levels still remaining in the flask after 14 days. Higher rates of degradation for the natural hormone versus the two synthetic compounds have been reported in previous studies (Ying and Kookana, 2003; Jurgens et al., 1999; Jürgens et al., 2002). It has been suggested that anthropogenic compounds often have substituents and structural features that differ from those in natural products, and indigenous micro-organisms may not have evolved the enzymes to deal with these chemical structures (Cerniglia and Heitcamp, 1989). At the beginning of the experiments, a high rate of sorption onto sediments was observed, with sorption following the order of EE2 N E2 N BPA. This trend is expected based on the hydrophobicity of the compounds as measured by the octanol–water partition coefficient (Table 2). However, as the experiment progressed there was proportionally more of the compound associated with the sediment phase than the water phase. For a compound to biodegrade, it must be freely bioavailable to micro-organisms (Dagley 1984; Schwarzenback et al., 2003) that attack the soluble compounds at the water sediment interface, transform them and eventually eliminate them through mineralization. As the aqueous phase concentrations decrease, the sediment phase will not be in equilibrium. To regain equilibrium in the system, some of the compound in the sediment phase must desorb and partition back into the aqueous phase. For many organic compounds the desorption process has been described as slow, resistant and not following equilibrium (Cornelissen et al., 2005). This may indicate that in our experiments, the rapid initial degradation rate is a result of aqueous phase degradation of the compound, while the decrease in degradation rates was caused by the slow desorption of the compound from sediment particles into the aqueous phase. To compare the differences in degradation rates for the three compounds, simple first-order reaction kinetics (Jürgens et al., 2002; Boethling and Alexander, 1979; Battersby 1990) were used to calculate compound specific half-lives (Table 2). These calculations show that there was variability in the degradation rates between the three compounds over time. For example, E2 had the shortest half-lives in seawater (1 day) over all three sampling periods. EE2 had half-lives ranging from 3 to 5 days, while BPA was the most resistant to degradation, with half-lives of 4 to 14 days. A limited number of other studies have used first-order rate models to describe biodegradation results. Ying et al. (2003) reported half-lives of 2 days for E2 and 81 days for EE2, while BPA concentrations remained unchanged after 70 days. The half-life of E2 in experiments using freshwater from rivers in England was 2 days or less in most samples, although in one case it was as high as 8 days (Jürgens et al., 2002), while EE2 had a half-life of 17 days (Jurgens et al., 1999). Since the biodegradation experiments were performed using samples collected at three different times of the year, there was some variability in the degradation rates as shown by the error bars in Fig. 1. These changes in the degradation rates over time may be due to variability in the abundance and makeup of the microbial communities in the natural environment. Numerous environmental factors can affect the growth and composition of bacterial communities
Fig. 1. Biodegradation of BPA (top), E2 (middle) and EE2 (lower) in samples collected from Halifax Harbour in July 2004, November 2004 and March 2005. Degradation results for both aqueous and sediment phases are shown.
including oxygen, pH, salinity, temperature and organic carbon (Schwarzenback et al., 2003; Robinson, 2006; Robinson et al, 2009). Two of these parameters, temperature and TOC were measured and found to vary between the three sampling times (Table 3). These differences are due to natural variability in the environment, as well as changes in discharges and disturbances near the sewage effluents. However, the rapid degradation of E2 was nearly identical in all experiments, suggesting this compound is readily degraded by a variety of micro-organisms. It is also possible that abiotic factors play a role in degrading the three EDC (Kang and Kondo, 2005). Results from the sterile control experiment suggest that this may be the case, with a 40% loss of E2 after 14 days (Fig. 2). This loss may be due to processes
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Table 2 Summary of first-order rate calculations Halifax Harbour samples and the Hantsport reference site.
1
(Jurgens et al., 1999; Lai et al., 2000; Robinson et al., 2009).
Table 3 Total organic carbon (TOC) content, microbiology results and water temperatures for samples used for the biodegradation experiments. Site
Date sampled
Sediment TOC %
Total SW bacteria #/mL
Total coliforms (SW/SED) #/100 mL
Fecal coliforms (SW/SED) #/100 mL
Water temp @ collection (°C)
Sediment grain size (percent b 100 µm)
Site C
07/04 11/04 03/05 07/04
6.76 8.26 18.90 0.69
– 2,163,043 3,673,881 –
1,600,000/1700 – – –
540,000/130 – – –
16 11 1 –
100% 100% 100% 80%
Hantsport
such as volitization and chemical oxidation (Cerniglia and Heitcamp, 1989), although it may also be due to difficulties encountered in effectively sterilizing sediment and autoclaving resistant microorganisms (eg. bacterial spores). 3.2. Effect of bacterial composition and growth
Fig. 2. Biodegradation of E2 in sterile control samples (top) and Hantsport reference site (bottom).
The observed rate of degradation was relatively fast, with half-lives of less than 2 weeks for the three compounds. Conversely, several other studies using samples collected from relatively pristine environments report half-lives in the range of weeks to months (Ying and Kookana, 2003; Jürgens et al., 2002; Ying and Kookana, 2003), where the initial rate of degradation was much slower, suggesting that acclimation was necessary for the micro-organisms to adapt before degradation began. In contrast, the study by Dorn et al. (1987) reported that using contaminated effluent from a plastics factory resulted in complete degradation of BPA in the mg/L range after four days of exposure. Presently, the lack of lag phase would demonstrate the adaptation of the microbial community to readily catabolise the three EDC. Results from biodegradation experiments using samples from the Hantsport reference site seem to confirm this hypothesis, with almost no degradation over the first three days for E2, the most readily degraded compound (Fig. 2). These differences in degradation rates between clean and contaminated environments would indicate that natural microbial assemblages can adapt and develop the capability to degrade organic compounds, while in media without previous exposure, degradation rates are much slower (Neilson 2000; Ying and Kookana, 2003). Halifax Harbour would represent an area where micro-organisms have adapted to anthropogenic contaminants, having been used for the disposal of human and industrial wastes for over two centuries. In the 1990s, over 180,000 m3 of raw untreated sewage was released into the harbour per day (Buckley et al., 1995). However, in 2005 the construction of three additional STP began, consolidating 44 sewer outfalls and 10
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growth rates was not significant since the total amount of compound that was spiked into the flask was very small (50 μg) compared to the large amount of other carbon sources in the system, as reflected by the TOC values. The rapid growth of total bacteria in all three flasks would be due to an increase in available oxygen due to stirring of the solution compared to what the organisms would be exposed to in the natural undisturbed environment. Rapid bacterial growth would also be due to the fact that the experiments were conducted at room temperature as opposed to lower temperatures that would be observed in the environment. It was reported that microbial growth was delayed up to 25 days when experiments were conducted at 4 °C (Kang and Kondo, 2005). The combination of both increased temperature and oxygen in our experiments would have contributed to the observed high rates of bacterial growth and subsequent chemical degradation. 3.3. Formation of metabolites
Fig. 3. Comparison of bacterial growth measured using flow cytrometry for BPA, E2 and EE2 in the contained flask system used during the biodegradation experiments.
fluvial drainage systems, the first of which became functional by the end of 2007. Another factor that may have influenced the rates of biodegradation during the experiments was microbial growth within the contained flask system. To monitor bacterial growth during an experiment, water samples were collected from each of the BPA, E2 and EE2 flasks and total bacterial numbers measured by flow cytometry (Fig. 3). These analyses revealed that rapid microbial growth in the flask occurred at the start of the experiment, with initial levels of 2 million/mL at time zero and growing to 15–20 million/mL after four days. An almost identical trend in bacterial growth was observed by Kang and Kondo (2005) where BPA degradation was conducted at 25 °C. In our experiments, the bacterial growth for all three compounds was very similar, although the E2 flask had more rapid growth during the first two days compared to BPA and EE2. Although this may indicate that bacteria were better able to utilize E2 in their metabolic pathways, it is more likely that this difference in
The biodegradation of a complex molecule does not always lead to the direct formation of carbon dioxide; intermediate molecules can be formed with parts of the structure of the original molecule remaining intact (Neilson 2000). The lifetime of intermediates or metabolites can vary depending on their stability as well as their susceptibility to further degradation. Some indirect observations were made during the analysis of several extracts from the biodegradation experiments. HPLC chromatograms of seawater extracts (Fig. 4) from one of the E2 biodegradation experiments show that over the course of one week, the degradation of E2 corresponded with the formation and further degradation of two unknown compounds. It is likely that these unknown compounds were degradation products of E2. Estradiol is known to be oxidized by bacteria to estrone (Johnson et al., 1998), but these unknown peaks did not match the retention time of a commercial estrone standard. Only one of the unknown metabolites was abundant enough to obtain a ultraviolet (UV) spectrum that could be used for a library search. Although the signal was weak, the UV spectrum indicates that this compound would have retained the basic conjugated unsaturations, i.e. the chormophore moiety of estrone in its structure (Fig. 4). Since this peak disappeared with time, the intermediate does not persist, once again pointing to the adaptation of the microbial community to the presence of diverse structures of
Fig. 4. Chromatograms obtained using HPLC-fluorescence detection showing the rapid degradation of estradiol in the aqueous phase, and the corresponding formation of two unknown metabolites over a 1 week time period. The UV spectrum (inset) of unknown metabolite#2 is compared to an estrone standard.
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anthropogenic compounds. The common moiety in the examined EDC that made it possible to analyse the compounds by fluorescence is due to the phenolic ring. This aromatic ring is also present in other compounds, such as nonylphenol ethoxylates and anti-oxidants such as butylated hydroxy toluene (BHT). These anthropogenic compounds are also discharged in sewage effluents and present in a harbour ecosystem, along with natural phenolic compounds associated with foods, although the formation of more polar compounds was not observed with the fluorescence detector set for sensitivity towards the presence of a monoaromatic phenol substituent. 4. Conclusion Studies conducted using anaerobic conditions have shown either significantly reduced rates of biodegradation of E2 (Ying and Kookana, 2003), or in some cases, no detectable biodegradation over the course of the entire experiment for BPA and EE2 (Ying and Kookana, 2003; Kang and Kondo 2005). Based on these findings, some authors have concluded that these compounds could persist and accumulate in anoxic environments (Ying and Kookana 2003), such as many surface sediments in Halifax Harbour. Indeed, BPA, E2 and EE2 were detected in seawater and sediments collected from seven sites in the harbour and at the highest concentrations in a secondary sewage treatment plant (Robinson et al., 2009). Regardless of the propensity of the studied chemicals to reach mineralization in either seawater or sediments of Halifax Harbour, the rate of discharge is such that they persist at trace levels. The results from this study indicate that native microbial communities can degrade EDC, and so additional research is needed to examine the role of additional sewage treatment plants constructed after 2007 relative to the fate of contaminants in this waterway that is now increasingly used for recreational purposes. Acknowledgements This project was supported by funding from the Natural Sciences and Engineering Research Council (NSERC), Dalhousie University and Fisheries and Oceans Canada. The authors would like to thank Dr. Bill Li (Fisheries and Oceans) for providing the flow cytrometry analysis, Diane Tremblay and Laura O'Connor (Environment Canada) for assisting with coliform analysis and William LeBlanc (Natural Resources Canada) for providing organic carbon measurements. We also thank the anonymous reviewers for their helpful comments. References Battersby NS. A review of biodegradation kinetics in the aquatic environment. Chemosphere 1990;21:1243–84. Boethling RS, Alexander M. Effect of concentration of organic chemicals on their biodegradation by natural microbial communities. Appl Environ Microbiol 1979;37: 1211–6. Bowman JC, Zhou JL, Readman JW. Sediment-water interactions of natural oestrogens under estuarine conditions. Mar Chem 2002;77:263–76. Buckley DE, Smith JN, Winters GV. Accumulation of contaminant metals in marine sediments of Halifax Harbour, Nova Scotia: environmental factors and historical trends. Appl Geochem 1995;10:175–95. Cao XL, Dufresne G, Belisle S, Clement G, Falicki M, Beraldin F, Rulibikiye A. Levels of bisphenol a in canned liquid infant formula products in Canada and dietary intake estimates. J Agric Food Chem 2008;56:7919–24.
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Science of the Total Environment j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / s c i t o t e n v
Biodegradation of endocrine disrupting compounds in harbour seawater and sediments Brian J. Robinson a,b, Jocelyne Hellou a,b,c,⁎ a b c
Fisheries and Oceans Canada, Bedford Institute of Oceanography, PO Box 1006, Dartmouth, Nova Scotia, Canada B2Y 4A2 Department of Oceanography, Dalhousie University, Halifax, Nova Scotia, Canada B3H 4J1 Department of Chemistry, Dalhousie University, Halifax, Nova Scotia, Canada B3H 4J1
a r t i c l e
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Article history: Received 18 March 2009 Received in revised form 1 July 2009 Accepted 5 July 2009 Available online 7 August 2009 Keywords: Biodegradation Halifax Harbour Endocrine Disrupting Compounds Metabolites
a b s t r a c t The biodegradation of three endocrine disrupting compounds was examined using samples of seawater and sediment collected from Halifax Harbour, Nova Scotia, Canada, an urbanized harbour impacted by over two centuries of anthropogenic contamination. Flask experiments, where the samples were mixed to form a slurry were used to monitor the aerobic biodegradation of the synthetic plasticizer bisphenol A (BPA), the natural hormone 17β-estradiol (E2), and the pharmaceutical and contraceptive ethinylestradiol (EE2). Degradation rates followed the order E2 N EE2 N BPA with half-lives of up to 1, 5 and 14 days in seawater, respectively. A rapid initial degradation rate for all three compounds with no apparent lag phase indicated the ability of the microbial community to readily catabolise the chemicals. The formation of unidentified non-persistent intermediate metabolites was observed during the E2 degradation experiments. These degradation rates are more rapid and complete than reported in previous studies, indicating the adaptation of native microbial communities to these contaminants. Crown Copyright © 2009 Published by Elsevier B.V. All rights reserved.
1. Introduction The occurrence of endocrine disrupting compounds (EDC) in the aquatic environment has generated worldwide interest because these chemicals can cause feminization of fish as well as interfere with the reproduction and development of other aquatic organisms (Purdom et al., 1994; Harries et al., 1996; Larsson et al., 1999). Most EDC are released through sewage effluents, and like all organic compounds their fate in the aquatic environment is multi-faceted, with as a general rule, non-ionic compounds with an octanol–water partition coefficient (log Kow) below 3.0 expected to reside preferentially in the water phase, where they can be dispersed, diluted, degraded and sometimes photooxidized. Lipophilic contaminants with a log Kow N 3.0 will be attracted to lipid enriched particles, further transported in the aquatic environment under consideration and deposited in sediments. This view comes from the basic definition of the partition coefficient where octanol represents a substitute for lipid in organisms or of the total organic carbon (TOC) content of sediments (Mackay, 1991). When comparing the concentration of contaminants in sediments, values are typically normalised to a TOC content of 1%, as discussed in Swartz (1999). In contrast, the dissolved organic carbon (DOC) content of seawater in a near-shore turbulent location is in the order of 0.0001% (Kepkay 2000) and even ⁎ Corresponding author. Ecosystem Research Division, Bedford Institute of Oceanography, PO Box 1006, Dartmouth, NS, Canada B2Y 4A2. Tel.: +1 902 426 7451; fax: +1 902 426 6695. E-mail address: [email protected] (J. Hellou).
then with less affinity for polar organic compounds. Once discharged in the environment, organic chemicals can also be altered through chemical, photo- and bio-degradation producing derivatives with longer or shorter half-lives. For the past fifteen years, concern has been directed towards the fate of endocrine disrupting compounds (EDC). The majority of research on the biodegradation of EDC has focused on degradation during the sewage treatment process (Ternes et al., 1999; Lee and Liu, 2002). Few studies have examined partitioning (Lai et al., 2000; Bowman et al., 2002; Robinson et al., 2009) or the biodegradation of EDC in field collected marine samples (Ying and Kookana, 2003). Halifax Harbour, located in Nova Scotia, Canada, is the largest marine port in Atlantic Canada and it represents a unique location to study the fate of EDC since the harbour and its surrounding waters have been used for the disposal of human and industrial wastes for over a century. However, over the last decade the municipality invested millions of dollars in the construction of sewage treatment plants to process the 180 million liters of sewage that is released into the harbour every day. Since the construction of three STP had not been completed prior to the start of this study, it was of interest to investigate the presence of EDC for future comparison. The goal of the present research was to use seawater and sediment from Halifax Harbour, Nova Scotia, Canada, to examine the ability of local microbial communities to biodegrade EDC deriving from three distinct input sources that are discharged in combined sewage effluents. Bisphenol A (BPA) is a synthetic plasticizer commonly used in polycarbonates and has been the subject of recent interest due to links to human health effects (Cao et al., 2008). The natural compound 17β-
0048-9697/$ – see front matter. Crown Copyright © 2009 Published by Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2009.07.003
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estradiol (E2) is produced by males and females alike but in varying proportion depending on the age and life cycle of individuals. The pharmaceutical product 17α-ethinylestradiol (EE2) represents the active ingredient in the birth control pill and is detected in amounts that depend on the use of various contraceptives by a population. The potential variation in the biodegradation rates of these three EDC was examined through experiments using seawater and sediment samples collected at different times of the year near a major sewage effluent in Halifax Harbour. Biodegradation trends were compared to the levels of bacteria in field samples, as well as total bacterial growth during flask experiments. Simple first-order reaction kinetics were used to determine degradation rates and half-lives. The possible identity of metabolites formed during one of the experiments is discussed, and the importance of other modes of degradation is summarized. 2. Materials and methods 2.1. Sample collection Seawater and sediment were collected on three occasions (July 2004, November 2004 and March 2005) from Halifax Harbour, Nova Scotia, Canada at a location near a major sewage effluent (44°40.28′ N 63°35.94′ W). This location has been documented as one of the most polluted sites in the harbour for organic contaminants (Site 7 in Hellou et al. 2002; Site C in Robinson et al. 2009). Seawater samples were collected at 10–15 cm below the water surface in 4 L glass bottles and stored at 5 °C, while the top 2–3 cm of sediments were collected in an Eckman grab sampler, passed through a 2 mm sieve to remove large debris and stored in mason jars at 5 °C prior to use. Sediment subsamples were set aside from all locations for total organic carbon (TOC) analysis. The organic carbon content of the sediments was determined using a Leco WR-112 carbon analyzer after treatment with 1 M hydrochloric acid to remove inorganic carbon (Sunderland et al. 2004). 2.2. Experimental set-up The biodegradation of BPA, E2 and EE2 was measured through a series of experiments whereby seawater and sediment samples were mixed to form a slurry. A similar technique has been used in previous studies (Jurgens et al., 1999; Lai et al., 2000; Bowman et al., 2002; Jürgens et al., 2002; Ying and Kookana, 2003), although our experiments differed in that concentrations in both the aqueous and sediment phases were monitored over the course of the experiment. The biodegradation experiments were performed in 250 mL Erlenmeyer flasks wrapped with aluminum foil and incubated at 20 ± 1 °C. This temperature was chosen since it is the conventional temperature used for biodegradation experiments (OECD) making it easier to compare results between our experiments and published studies. Each flask was spiked with a standard containing 50 μg of each compound dissolved in methanol. Experiments with the July 2004 samples were conducted using individual flasks for each compound; for all remaining experiments the EDC were combined in one flask. Seawater (230 mL) and sediment (10 g wet weight) were combined in the flask and mixed with a Teflon magnetic stir-bar at a medium setting. At specified time points, subsamples were taken from each flask using a 20 mL graduated pipette to remove an aliquot of the seawater-sediment mixture while the stir-bar was still mixing. These samples were collected in pre-weighed Teflon tubes and centrifuged at 2000 rpm for 10 min. The water and sediment phases were then separated and the weight of each phase was recorded. On average, each sample contained approximately 19 mL of seawater and 1 g (wet weight) of sediment. Prior to extraction, each seawater and sediment sample was spiked with 1.25 μg of a surrogate standard to monitor the extraction efficiency. For the experiments using individual flasks per compound, one of the remaining EDC was used as a surrogate (eg. for the E2 experiment, BPA was used as a surrogate). For all remaining experiments with three EDC combined in the Erlenmeyer
flask, ethinylestradiol-3-methyl-ether was used as the surrogate. The seawater phase was extracted three times using liquid–liquid extraction with 10 mL of ethyl acetate in a separatory funnel. Sodium sulphate was used to remove any water remaining in the solvent prior to transfer to a round bottom flask. This solvent was evaporated on a rotary-evaporator, transferred to 1 mL of methanol, and stored in a 2 mL vial for analysis. The sediment phase was dried by adding 3 g of sodium sulphate and then extracted in the Teflon centrifuge tube with 3 × 10 mL of ethyl acetate using a sonic probe. The solvent was collected in a round bottom flask, evaporated on a rotary-evaporator, transferred to 1 mL of methanol and stored in a 2 mL vial for analysis. 2.3. Sterile control and reference experiments A series of sterile control and reference biodegradation experiments were conducted using E2 only. For the sterile control experiments, seawater and sediment samples from Halifax Harbour were autoclaved for 1 h at 121 °C. For the reference biodegradation experiments requiring samples from a pristine un-urbanised location, seawater and sediment were collected from Hantsport, Nova Scotia, Canada (Hellou et al., 2009). Both the sterile control and reference biodegradation experiments were performed using the methods described above. 2.4. HPLC analysis The methods for high performance liquid chromatography (HPLC) analysis are detailed elsewhere (Robinson, 2006; Robinson et al., 2009). Briefly, samples were analyzed on a Hewlett Packard (HP) 1090 HPLC equipped with an HP 1046A fluorescence detector. A C18 reversed phase column measuring 2.1 × 250 mm with a particle size of 5 μm (Vydac 201TP52) was used for separations with a 2.1 × 12 mm Vydac guard column. Flow rate was maintained at 0.2 mL/min, and the mobile phase consisted of water (A) and acetonitrile (B) with 0.1% trifluoroacetic acid, pH b 3. The column was kept at 40 °C and the sample injection volume was 3 μL. Chromatographic separation was achieved with the following gradient elution: 0 to 3 min 40% B, 3 to 12 min 40 to 75% B, 12 to 15 min 75 to 100% B, 15 to 17 min 100% B, 17 to 23 min 100 to 40% B, and 23 to 26 min 40% B. The excitation and emission wavelengths were set at 226 nm and 310 nm on the fluorescence detector. Quantification was performed using a five point external calibration curve with standards ranging from 0.1 ng/μL to 4 ng/μL. Quality control consisted of running periodical procedural blanks to ensure there was no background interference from the sample matrix. Table 1 contains a summary of compound specific retention times, calibration coefficients, percent relative standard deviations (%RSD) and average percent recoveries as determined during method development. 2.5. Bacterial analysis To confirm that the seawater and sediment at site C were impacted by sewage effluents, samples collected in July 2004 were analyzed using the most probable number (MPN) fermentation method for total and fecal coliforms (Horwitz 2000). All remaining samples collected for this study were analyzed for total bacterial numbers using a flow cytometer to count DNA-stained bacteria (Li and Dickie, Table 1 Summary of HPLC method performance including compound specific retention times, calibration coefficients, percent relative standard deviations (%RSD) and average percent recoveries. Compound
Retention time (min)
Calibration coefficient
%RSD
Average % recovery
BPA E2 EE2 EE2-3-Methyl-Ether
6.55 7.70 8.52 14.75
N0.999 N0.999 N0.999 N0.999
1.3 1.2 1.6 1.6
80 ± 12 91 ± 8 95 ± 7 101 ± 6
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2001). The preparation of samples for this procedure consisted of taking 2 mL of water and preserving the samples with 1% paraformaldehyde for 10 min at room temperature and then storing them at − 80 °C in cryogenic vials. Analysis was performed using a FACSort instrument equipped with a 488 nm argon laser. 3. Results and discussion 3.1. Biodegradation results and kinetics In all experiments, E2 was the most readily degraded of the three compounds with a rapid decrease in the observed aqueous and sediment phase concentrations over the first 48 h of experiments conducted at three points in time (Fig. 1). E2 was degraded to nondetectable levels after 14 days. BPA and EE2 were more resistant to biodegradation, with detectable levels still remaining in the flask after 14 days. Higher rates of degradation for the natural hormone versus the two synthetic compounds have been reported in previous studies (Ying and Kookana, 2003; Jurgens et al., 1999; Jürgens et al., 2002). It has been suggested that anthropogenic compounds often have substituents and structural features that differ from those in natural products, and indigenous micro-organisms may not have evolved the enzymes to deal with these chemical structures (Cerniglia and Heitcamp, 1989). At the beginning of the experiments, a high rate of sorption onto sediments was observed, with sorption following the order of EE2 N E2 N BPA. This trend is expected based on the hydrophobicity of the compounds as measured by the octanol–water partition coefficient (Table 2). However, as the experiment progressed there was proportionally more of the compound associated with the sediment phase than the water phase. For a compound to biodegrade, it must be freely bioavailable to micro-organisms (Dagley 1984; Schwarzenback et al., 2003) that attack the soluble compounds at the water sediment interface, transform them and eventually eliminate them through mineralization. As the aqueous phase concentrations decrease, the sediment phase will not be in equilibrium. To regain equilibrium in the system, some of the compound in the sediment phase must desorb and partition back into the aqueous phase. For many organic compounds the desorption process has been described as slow, resistant and not following equilibrium (Cornelissen et al., 2005). This may indicate that in our experiments, the rapid initial degradation rate is a result of aqueous phase degradation of the compound, while the decrease in degradation rates was caused by the slow desorption of the compound from sediment particles into the aqueous phase. To compare the differences in degradation rates for the three compounds, simple first-order reaction kinetics (Jürgens et al., 2002; Boethling and Alexander, 1979; Battersby 1990) were used to calculate compound specific half-lives (Table 2). These calculations show that there was variability in the degradation rates between the three compounds over time. For example, E2 had the shortest half-lives in seawater (1 day) over all three sampling periods. EE2 had half-lives ranging from 3 to 5 days, while BPA was the most resistant to degradation, with half-lives of 4 to 14 days. A limited number of other studies have used first-order rate models to describe biodegradation results. Ying et al. (2003) reported half-lives of 2 days for E2 and 81 days for EE2, while BPA concentrations remained unchanged after 70 days. The half-life of E2 in experiments using freshwater from rivers in England was 2 days or less in most samples, although in one case it was as high as 8 days (Jürgens et al., 2002), while EE2 had a half-life of 17 days (Jurgens et al., 1999). Since the biodegradation experiments were performed using samples collected at three different times of the year, there was some variability in the degradation rates as shown by the error bars in Fig. 1. These changes in the degradation rates over time may be due to variability in the abundance and makeup of the microbial communities in the natural environment. Numerous environmental factors can affect the growth and composition of bacterial communities
Fig. 1. Biodegradation of BPA (top), E2 (middle) and EE2 (lower) in samples collected from Halifax Harbour in July 2004, November 2004 and March 2005. Degradation results for both aqueous and sediment phases are shown.
including oxygen, pH, salinity, temperature and organic carbon (Schwarzenback et al., 2003; Robinson, 2006; Robinson et al, 2009). Two of these parameters, temperature and TOC were measured and found to vary between the three sampling times (Table 3). These differences are due to natural variability in the environment, as well as changes in discharges and disturbances near the sewage effluents. However, the rapid degradation of E2 was nearly identical in all experiments, suggesting this compound is readily degraded by a variety of micro-organisms. It is also possible that abiotic factors play a role in degrading the three EDC (Kang and Kondo, 2005). Results from the sterile control experiment suggest that this may be the case, with a 40% loss of E2 after 14 days (Fig. 2). This loss may be due to processes
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Table 2 Summary of first-order rate calculations Halifax Harbour samples and the Hantsport reference site.
1
(Jurgens et al., 1999; Lai et al., 2000; Robinson et al., 2009).
Table 3 Total organic carbon (TOC) content, microbiology results and water temperatures for samples used for the biodegradation experiments. Site
Date sampled
Sediment TOC %
Total SW bacteria #/mL
Total coliforms (SW/SED) #/100 mL
Fecal coliforms (SW/SED) #/100 mL
Water temp @ collection (°C)
Sediment grain size (percent b 100 µm)
Site C
07/04 11/04 03/05 07/04
6.76 8.26 18.90 0.69
– 2,163,043 3,673,881 –
1,600,000/1700 – – –
540,000/130 – – –
16 11 1 –
100% 100% 100% 80%
Hantsport
such as volitization and chemical oxidation (Cerniglia and Heitcamp, 1989), although it may also be due to difficulties encountered in effectively sterilizing sediment and autoclaving resistant microorganisms (eg. bacterial spores). 3.2. Effect of bacterial composition and growth
Fig. 2. Biodegradation of E2 in sterile control samples (top) and Hantsport reference site (bottom).
The observed rate of degradation was relatively fast, with half-lives of less than 2 weeks for the three compounds. Conversely, several other studies using samples collected from relatively pristine environments report half-lives in the range of weeks to months (Ying and Kookana, 2003; Jürgens et al., 2002; Ying and Kookana, 2003), where the initial rate of degradation was much slower, suggesting that acclimation was necessary for the micro-organisms to adapt before degradation began. In contrast, the study by Dorn et al. (1987) reported that using contaminated effluent from a plastics factory resulted in complete degradation of BPA in the mg/L range after four days of exposure. Presently, the lack of lag phase would demonstrate the adaptation of the microbial community to readily catabolise the three EDC. Results from biodegradation experiments using samples from the Hantsport reference site seem to confirm this hypothesis, with almost no degradation over the first three days for E2, the most readily degraded compound (Fig. 2). These differences in degradation rates between clean and contaminated environments would indicate that natural microbial assemblages can adapt and develop the capability to degrade organic compounds, while in media without previous exposure, degradation rates are much slower (Neilson 2000; Ying and Kookana, 2003). Halifax Harbour would represent an area where micro-organisms have adapted to anthropogenic contaminants, having been used for the disposal of human and industrial wastes for over two centuries. In the 1990s, over 180,000 m3 of raw untreated sewage was released into the harbour per day (Buckley et al., 1995). However, in 2005 the construction of three additional STP began, consolidating 44 sewer outfalls and 10
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growth rates was not significant since the total amount of compound that was spiked into the flask was very small (50 μg) compared to the large amount of other carbon sources in the system, as reflected by the TOC values. The rapid growth of total bacteria in all three flasks would be due to an increase in available oxygen due to stirring of the solution compared to what the organisms would be exposed to in the natural undisturbed environment. Rapid bacterial growth would also be due to the fact that the experiments were conducted at room temperature as opposed to lower temperatures that would be observed in the environment. It was reported that microbial growth was delayed up to 25 days when experiments were conducted at 4 °C (Kang and Kondo, 2005). The combination of both increased temperature and oxygen in our experiments would have contributed to the observed high rates of bacterial growth and subsequent chemical degradation. 3.3. Formation of metabolites
Fig. 3. Comparison of bacterial growth measured using flow cytrometry for BPA, E2 and EE2 in the contained flask system used during the biodegradation experiments.
fluvial drainage systems, the first of which became functional by the end of 2007. Another factor that may have influenced the rates of biodegradation during the experiments was microbial growth within the contained flask system. To monitor bacterial growth during an experiment, water samples were collected from each of the BPA, E2 and EE2 flasks and total bacterial numbers measured by flow cytometry (Fig. 3). These analyses revealed that rapid microbial growth in the flask occurred at the start of the experiment, with initial levels of 2 million/mL at time zero and growing to 15–20 million/mL after four days. An almost identical trend in bacterial growth was observed by Kang and Kondo (2005) where BPA degradation was conducted at 25 °C. In our experiments, the bacterial growth for all three compounds was very similar, although the E2 flask had more rapid growth during the first two days compared to BPA and EE2. Although this may indicate that bacteria were better able to utilize E2 in their metabolic pathways, it is more likely that this difference in
The biodegradation of a complex molecule does not always lead to the direct formation of carbon dioxide; intermediate molecules can be formed with parts of the structure of the original molecule remaining intact (Neilson 2000). The lifetime of intermediates or metabolites can vary depending on their stability as well as their susceptibility to further degradation. Some indirect observations were made during the analysis of several extracts from the biodegradation experiments. HPLC chromatograms of seawater extracts (Fig. 4) from one of the E2 biodegradation experiments show that over the course of one week, the degradation of E2 corresponded with the formation and further degradation of two unknown compounds. It is likely that these unknown compounds were degradation products of E2. Estradiol is known to be oxidized by bacteria to estrone (Johnson et al., 1998), but these unknown peaks did not match the retention time of a commercial estrone standard. Only one of the unknown metabolites was abundant enough to obtain a ultraviolet (UV) spectrum that could be used for a library search. Although the signal was weak, the UV spectrum indicates that this compound would have retained the basic conjugated unsaturations, i.e. the chormophore moiety of estrone in its structure (Fig. 4). Since this peak disappeared with time, the intermediate does not persist, once again pointing to the adaptation of the microbial community to the presence of diverse structures of
Fig. 4. Chromatograms obtained using HPLC-fluorescence detection showing the rapid degradation of estradiol in the aqueous phase, and the corresponding formation of two unknown metabolites over a 1 week time period. The UV spectrum (inset) of unknown metabolite#2 is compared to an estrone standard.
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anthropogenic compounds. The common moiety in the examined EDC that made it possible to analyse the compounds by fluorescence is due to the phenolic ring. This aromatic ring is also present in other compounds, such as nonylphenol ethoxylates and anti-oxidants such as butylated hydroxy toluene (BHT). These anthropogenic compounds are also discharged in sewage effluents and present in a harbour ecosystem, along with natural phenolic compounds associated with foods, although the formation of more polar compounds was not observed with the fluorescence detector set for sensitivity towards the presence of a monoaromatic phenol substituent. 4. Conclusion Studies conducted using anaerobic conditions have shown either significantly reduced rates of biodegradation of E2 (Ying and Kookana, 2003), or in some cases, no detectable biodegradation over the course of the entire experiment for BPA and EE2 (Ying and Kookana, 2003; Kang and Kondo 2005). Based on these findings, some authors have concluded that these compounds could persist and accumulate in anoxic environments (Ying and Kookana 2003), such as many surface sediments in Halifax Harbour. Indeed, BPA, E2 and EE2 were detected in seawater and sediments collected from seven sites in the harbour and at the highest concentrations in a secondary sewage treatment plant (Robinson et al., 2009). Regardless of the propensity of the studied chemicals to reach mineralization in either seawater or sediments of Halifax Harbour, the rate of discharge is such that they persist at trace levels. The results from this study indicate that native microbial communities can degrade EDC, and so additional research is needed to examine the role of additional sewage treatment plants constructed after 2007 relative to the fate of contaminants in this waterway that is now increasingly used for recreational purposes. Acknowledgements This project was supported by funding from the Natural Sciences and Engineering Research Council (NSERC), Dalhousie University and Fisheries and Oceans Canada. The authors would like to thank Dr. Bill Li (Fisheries and Oceans) for providing the flow cytrometry analysis, Diane Tremblay and Laura O'Connor (Environment Canada) for assisting with coliform analysis and William LeBlanc (Natural Resources Canada) for providing organic carbon measurements. We also thank the anonymous reviewers for their helpful comments. References Battersby NS. A review of biodegradation kinetics in the aquatic environment. Chemosphere 1990;21:1243–84. Boethling RS, Alexander M. Effect of concentration of organic chemicals on their biodegradation by natural microbial communities. Appl Environ Microbiol 1979;37: 1211–6. Bowman JC, Zhou JL, Readman JW. Sediment-water interactions of natural oestrogens under estuarine conditions. Mar Chem 2002;77:263–76. Buckley DE, Smith JN, Winters GV. Accumulation of contaminant metals in marine sediments of Halifax Harbour, Nova Scotia: environmental factors and historical trends. Appl Geochem 1995;10:175–95. Cao XL, Dufresne G, Belisle S, Clement G, Falicki M, Beraldin F, Rulibikiye A. Levels of bisphenol a in canned liquid infant formula products in Canada and dietary intake estimates. J Agric Food Chem 2008;56:7919–24.
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