1 Evidence for photochemical and microbial debromination of polybrominated 1 diphenyl ether flame retardants in San Francisco Bay sediment 2 3 Lisa A. Rodenburg 1* , Qingyu Meng 2 , Don Yee 3 , and Ben K. Greenfield 3,4 4 1 Department of Environmental Sciences, Rutgers University, 14 College Farm 5 Road, New Brunswick, NJ 08901, USA 6 2 School of Public Health, Rutgers University, Piscataway, New Jersey 08854, United States 7 3 San Francisco Estuary Institute, 4911 Central Avenue, Richmond, CA 94804 8 4 Current affiliation: Environmental Health Sciences Division, School of Public Health, 9 University of California, Berkeley, 50 University Hall #7360, Berkeley, CA 94720-7360 10 *Corresponding author. Phone 732-932-9800 x 6218, Fax 732-932-8644, email 11 [email protected]12 13
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Evidence for photochemical and microbial debromination of polybrominated diphenyl ether flame retardants in San Francisco Bay sediment
ABSTRACT Brominated diphenyl ethers (BDEs) are flame retardant compounds that have been classified as persistent organic pollutants under the Stockholm Convention and targeted for phase-out. Despite their classification as persistent, PBDEs undergo debromination in the environment, via both microbial and photochemical pathways. We examined concentrations of 24 PBDE congeners in 233 sediment samples from San Francisco Bay using Positive Matrix Factorization (PMF). PMF analysis revealed five factors, two of which contained high proportions of congeners with two or three bromines, indicating that they are related to debromination processes. One of the factors included PBDE 15 (4,4’-dibromo diphenyl ether, comprising 20% of the factor); the other included PBDE 7 (2,4-dibromo diphenyl ether; 12%) and PBDE 17 (2,2’,4-tribromo diphenyl ether; 16%). The debromination processes that produce these congeners are probably photochemical debromination and anaerobic microbial debromination, although other processes could also be responsible. Together, these two debromination factors represent about 8% of the mass and 13% of the moles of PBDEs in the data matrix, suggesting that PBDEs undergo measurable degradation in the environment.
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1
Evidence for photochemical and microbial debromination of polybrominated 1
diphenyl ether flame retardants in San Francisco Bay sediment 2
3
Lisa A. Rodenburg1*, Qingyu Meng2, Don Yee3, and Ben K. Greenfield3,4 4
1 Department of Environmental Sciences, Rutgers University, 14 College Farm 5
Road, New Brunswick, NJ 08901, USA 6
2School of Public Health, Rutgers University, Piscataway, New Jersey 08854, United States 7
3San Francisco Estuary Institute, 4911 Central Avenue, Richmond, CA 94804 8
4Current affiliation: Environmental Health Sciences Division, School of Public Health, 9
University of California, Berkeley, 50 University Hall #7360, Berkeley, CA 94720-7360 10
*Corresponding author. Phone 732-932-9800 x 6218, Fax 732-932-8644, email 11
PBDEs with distance from the probable site of release (Gulf of Lion, Mediterranean Sea). Zhao 323
et al. (2011) found that congeners that were believed to come from photolysis were the most 324
abundant congeners in the sediments of the Daliao River Estuary, China. 325
In a study similar to the present investigation, Zou et al. (2013) used PMF combined with 326
eigenspace projection to investigate PBDE congener patterns in sediments cores from the Great 327
Lakes. Using a data matrix of 10 congeners in 93 samples, they resolved five factors, thought to 328
represent the commercial penta, octa, and deca formulations, and two factors thought to represent 329
debromination products. The first of these was characterized by large contributions from PBDEs 330
66 and 85, while the other was characterized by a high proportion of PBDE 28. Unlike the 331
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present work, their study did not include congeners 7, 8, 12, 15, and 17, so it is difficult to 332
compare their factors with those resolved in the present study. Nevertheless, it is noteworthy 333
that two studies, one in a freshwater system and one in an estuary, identified two distinct 334
debromination signals in sediment. In the future, we recommend that PBDEs 7, 15, and 17 be 335
routinely monitored, since they appear to be markers for debromination processes. The present 336
analysis did not find an octa signal in San Francisco Bay. PBDE 183, the major congener in the 337
octa formulation, was not included in the data matrix because it was below detection in a 338
majority of samples. However, the maximum contribution of PBDE 183 to the sum of PBDEs 339
was 10%. PBDE 183 comprised more than 2% of the sum of PBDEs in only 11 of 344 samples. 340
In addition, the data matrix did include PBDEs 153, 154, and 155, which are prominent in the 341
octa formulation, yet PMF analysis did not resolve an octa factor. Thus we conclude that the 342
octa formulation is not a major source of PBDEs to San Francisco Bay, possibly due to lower 343
production of octa relative to penta and deca BDE formulations in the Americas (Birnbaum and 344
Staskal, 2004). Also, both the octa and penta BDE formulations have undergone a gradual 345
production bad in California, which was approved in 2003 and fully implemented in 2008. 346
Anaerobic microbial debromination, which we speculate is associated with Factor 3, may 347
have occurred in Bay sediments. Such a process has been shown to debrominate octa mixtures in 348
laboratory microcosms (Lee and He, 2010). It is instructive to consider the similar process of 349
bacterial dechlorination, since in both cases, anaerobic bacteria use the halogenated compound as 350
an electron acceptor. Bacteria capable of dechlorination have been isolated from Bay sediments 351
(Sun et al., 2000; He et al., 2006). Although some researchers have suggested that the threshold 352
concentration for PCB dechlorination is around 40 ppm (Cho et al., 2003), well above the 353
maximum of about 50 ppb reached in Bay sediments, others have shown evidence of PCB 354
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dechlorination with an enriched culture under substrate concentrations as low as 1 to 5 ppm 355
(Royal et al., 2003; Krumins et al., 2009; Payne et al., 2011). It is also possible that the threshold 356
concentration for PBDE debromination is lower than the threshold for PCBs. Previous studies 357
indicate microbial dechlorination in sewers (Rodenburg et al., 2010; Rodenburg et al., 2012), 358
raising the possibility of microbial debromination in San Francisco Bay area sewers, with 359
debromination products emitted to the Bay via stormwater or treated wastewater outfalls. As 360
noted above, several researchers have seen evidence of PBDE debromination in sewage sludge 361
and near sewage outfalls (La Guardia et al., 2007; Davis et al., 2012). The location of 362
degradation is important. If microbial debromination occurs in Bay sediments, then PBDEs will 363
be less persistent in the long run and will have less tendency to accumulate in sediments. In 364
contrast, if microbial debromination occurs primarily in sewers (i.e., prior to discharge), then 365
both the parent compounds and the debromination products would accumulate in sediments and 366
become problematic in the long term. Based on the increased percent contribution of factor 3 367
from Lower South Bay to South Bay, we hypothesize that the majority of microbial 368
debromination occurs in Bay sediments. Examination of congener ratios in wastewater and 369
treatment plant sludge would aid in confirming this. 370
Regardless of the location of dehalogenation, the results indicate that PBDEs undergo 371
measurable debromination in the environment. We cannot rule out the possibility that 372
debromination leads to the fully debrominated diphenyl ether or to bromophenols (Bendig and 373
Vetter, 2013), which were not measured in this data set. Thus our estimate that about 13% of the 374
moles of PBDEs in the Bay have undergone debromination is a lower bound. This is a relatively 375
large degree of transformation. By comparison, POPs such as PCBs and PCDD/Fs show no 376
evidence of degradation in most aquatic systems. Rodenburg et al. (2010) demonstrate that as 377
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much as 19% of the PCBs emitted by permitted dischargers in the Delaware River, USA, basin 378
were subject to dehalogenation, but the dechlorination products (primarily PCBs 4 and 19) are 379
barely detectable in Delaware River sediments (Praipipat et al., 2013). These lower chlorinated 380
congeners may be prevented from accumulating in sediments in part due to volatilization, 381
aerobic degradation, or dissolution and advective export. PCBs 4 and 19 have lower octanol-382
water partition coefficients (log Kow = 4.84 for PCB 4 and 5.16 for PCB 19) (Hansen et al., 1999) 383
than PBDE 17 (ranging 5.4 to 6.6) or PBDE 15 (reported at 5.48) (Wania and Dugani, 2003). 384
Thus, dehalogenated PBDE congeners are more likely to accumulate in sediments than the more 385
hydrophilic dehalogenated PCB congeners. The relative toxicity of PBDEs 7, 15, and 17 vs. 386
parent compounds is poorly characterized. If the debromination products have equal or greater 387
toxicity, this would be compounded by the greater environmental mobility and potential 388
exposure to aquatic life associated with lower hydrophobicities (Arnot and Gobas, 2003). For 389
example, PBDE 47 has a greater biota-sediment accumulation factor (BSAF) than PBDE 209 (La 390
Guardia et al., 2012), so debromination of PBDE 209 to PBDE 47 will result in greater 391
bioaccumulation of PBDEs in some organisms. This work has demonstrated that debromination 392
is an important process affecting the fate of PBDE formulations, and that the debromination 393
products accumulate in sediments. Further research is needed to determine the toxic impacts of 394
the debromination products. 395
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6 ACKNOWLEDGMENTS 397
Sediment PBDE collection and analysis was performed by Applied Marine Sciences, SFEI, the 398
East Bay Municipal Utility District, and Axys Analytical Services, and supported by the 399
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Regional Monitoring Program for Water Quality in the San Francisco Estuary. BG is supported 400
by a USEPA STAR Fellowship. 401
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Figure 1. Map of the San Francisco Bay showing municipal wastewater outfalls.
Table S-1. Pearson’s correlation coefficients (r) between the 24 congeners included in the data matrix for PMF analysis.
Determination of the correct number of factors
The number of factors was determined by a weight of evidence approach based on several criteria. First, the relative standard deviation of nine runs of the G matrix using different seed values increased from 1.1% when five factors were requested to 56% when six factors were requested. This is an indication that the PMF2 software cannot find a reproducible (stable) solution when six factors are requested. To determine whether rotation of the five factor solution was necessary, the F peak value was adjusted from -0.3 to 0.3 in 0.1 increments. The resulting Q values changed by less than 0.1%. An F peak value of 0.1 yielded the lowest Q value, but the factors were nearly identical to those generated when F Peak = 0 and the change in the Q value was negligible: Q equaled 3378.40 at F Peak = 0 and 3378.29 when F Peak = 0.1. Furthermore, G-space plots indicated that the five factors were independent of each other (Figure S-1). Thus rotation of the factors was not necessary. In contrast, two of the factors derived when six factors were requested were not independent of each other, as indicated by G-space plots.
This solution provided a good fit to 17 of the 24 congeners, based on an R2 value for the measured versus modeled concentrations of 0.7 or greater, with no more than one outlier discarded per congener. The congeners that were not well described by the five-factor solution were BDE 12 (R2 = 0.57), BDE 35 (0.41), BDE 37 (0.43), BDE 71 (0.44), BDE 85 (0.56), BDE 155 (0.49), and BDE 208 (0.69). BDEs 12, 35, 37, 71, and 155 were below detection limits in between 25 and 59 samples, partially explaining the discrepancy between the measured and modeled concentrations. None of these congeners was crucial to the interpretation of the factors.