Article pubs.acs.org/est
Cite This: Environ. Sci. Technol. 2019, 53, 8767−8775
Inference of Organophosphate Ester Emission History from Marine Sediment Cores Impacted by Wastewater Effluents Jun Li,†,‡ Jie Wang,† Allison R. Taylor,† Zachary Cryder,† Daniel Schlenk,† and Jay Gan†,* †
Department of Environmental Sciences, University of California, Riverside, California 92521, United States School of the Earth Sciences and Resources, China University of Geosciences, Beijing 100083, China
‡
Downloaded via IMPERIAL COLLEGE LONDON on August 7, 2019 at 01:47:32 (UTC). See https://pubs.acs.org/sharingguidelines for options on how to legitimately share published articles.
S Supporting Information *
ABSTRACT: Organophosphate esters (OPEs) have been in use as flame retardants for many decades, with their actual usage varying over time. Knowledge of the emission history of OPEs is valuable for improving our prediction of their environmental loadings and associated risks. In this study, concentrations and compositions of 10 OPEs were measured in three dated sediment cores from the Palos Verdes Shelf (PVS) off the coast of Los Angeles, which has been impacted by wastewater treatment plant (WWTP) effluents for over a century. The total OPE concentrations varied from 0.68 to 1064 ng/g along the sediment profile, with two apparent peaks. The first peak occurred in the 1970s, coinciding with peak emissions from WWTPs. The second peak appeared in the 2000s and was possibly attributed to increased consumption of OPEs as replacement flame retardants. Since downward movement of OPEs in the PVS sediment bed was retarded by their slow desorption, the reconstructed history likely provided an accurate picture of OPE emissions in Southern California and North America. These findings suggest that the near-shore marine sediments affected by WWTP effluents could serve as an environmental proxy documenting history in OPE use and emissions.
1. INTRODUCTION Organophosphate esters (OPEs) are an important class of chemicals that have been widely used as flame retardants and plasticizers in various products, including plastics, hydraulic fluids, and floor polishes.1,2 As many OPEs are routinely added into these products, they can easily escape from such treated materials into the environment.1,3 Several studies have shown associations between human body burdens of OPEs and adverse health effects such as thyroid hormone abnormality, allergic symptoms, and low birthweight.1,4,5 Contrary to initial scientific assumptions, OPEs are relatively persistent in the environment and are susceptible to offsite transport, much like the legacy flame retardants that they were intended to replace.6−8 This fate and transport behavior is reflected in the fact that OPE concentrations in various environmental matrices emulate past usage patterns.6,9−11 Since this chemical class has been in use for over half a century, with different congeners introduced at different periods,2,12 it is highly valuable to delineate OPE use and emission history. Such knowledge would improve our estimates for their past environmental loads and prediction of exposure to humans and wildlife. Due to restrictions on flame retardants such as polybrominated diphenyl ethers (PBDEs),13 the production and usage of OPEs as the major replacement flame retardant class has increased rapidly over the last several decades. For example, © 2019 American Chemical Society
the annual U.S. production volumes for chlorinated OPEs, including tris(2-chloroethyl) phosphate (TCEP), tris(2,3dichloro-n-propyl) phosphate (TCIPP), and tris(1,3-dichloro2-propyl) phosphate (TDCIPP), increased from less than 14000 tons in the 1980s to 38000 tons in the 2010s,14 while the production of tris(phenyl) phosphate (TPHP) increased from 4500 to 22700 tons.1,10 In the 2010s, the global demand of OPEs reached several hundred thousand tons, accounting for 30% of all flame retardants.8 In the U.S., which is the largest consumer and manufacturer of OPEs in the world, it was estimated that 1.3−2.8% of manufactured OPEs eventually entered wastewater treatment plants (WWTPs).9,12 Due to incomplete removal,15 a significant fraction of OPEs entering WWTPs may be discharged into the downstream aquatic environments.9,16 The amount and composition of released OPEs are expected to have changed over time, due to differences in production volumes, usage patterns, and regulatory actions on specific congeners.12 However, few studies have considered the historical trends and patterns of OPEs.6 Many researchers have shown that sediment cores in some surface aquatic Received: Revised: Accepted: Published: 8767
March 20, 2019 July 11, 2019 July 14, 2019 July 15, 2019 DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
Article
Environmental Science & Technology systems could provide a wealth of information on the emission and deposition history of persistent organics.17,18 However, for OPEs, some of the polar analogues may move downward, resulting in fractionation of the mixtures over time.6 The downward shift depends closely on the rapid desorption fraction, which can be approximated by the rapid desorption fraction (Frapid) derived by Tenax extraction.19 Previous studies showed that contaminant Frapid of hydrophobic and hydrophilic organic contaminants was significantly minimized following wastewater treatment.20,21 Therefore, marine sediments impacted by WWTP effluents may provide an accurate record of the emission history of contaminants such as OPEs. In this study, we collected three sediment cores from an ocean floor off the coast of Los Angeles. The cores were sampled from 80 m below water, from a site known to have a long history of contamination from WWTP discharge.22,23 The specific objectives were to (1) characterize levels and compositions of OPEs along the vertical profile to infer OPE deposition patterns over time, (2) assess the postdeposition movement of OPEs and the influence of Frapid in this process, and (3) estimate the annual inventories of OPEs due to WWTP discharge in this region and extrapolate the role of WWTPs as a source for the mass loading of OPEs into oceans on a global scale.
Figure 1. Map showing the sampling sites in the Palos Verdes Shelf. Each grid represents 10 km. An overview map of California showing the location of the study area is shown in the bottom left insert.
The age/depth relationship for the sediment cores was established based upon the analysis of 210Pb by γ-ray spectrometry. A constant initial concentration model was used for chronological calculation.22 The sediment cores covered about six decades of sediment accumulation (1940s2000s), and the average sedimentation rates were determined to be 0.65−1.23 cm/yr. The detailed method for sediment dating may be found in Liao et al. (2017). 22 The sedimentation rates in this study were close to those in Santschi et al. (2001),25 in which the sediment rates were calculated to be 1.1−1.8 cm/yr, with little change over time. The historical data of WWTP discharge and population growth (Figure 2a) were obtained from Los Angeles County Sanitation Districts26 and the U.S. Census Bureau (https:// www.census.gov), respectively. Sediment Sample Extraction and Cleanup. The extraction and cleanup procedures of sediment sample for OPEs,7−9 OPE metabolites,15 and PBDEs7 were based on published methods. After freeze-drying, approximately 20 g (dry weight) of sediment sample was spiked with known amounts of tri-nbutyl phosphate-d27 (TNBP-d27), tris(1,3-dichloro-2-propyl) phosphate-d15 (TDCIPP-d15), BDE-77, and BDE-166 as surrogate standards. The sediment sample was then extracted with 30 mL of dichloromethane/acetone (1:1, v/v) for 20 min in an ultrasonic water bath. The sediment slurry was centrifuged for 20 min at 2700 rpm to obtain the supernatant. The same extraction was repeated for two additional times. The extracts were combined and concentrated, exchanged to hexane as solvent, and then fractionated on a column containing deactivated silica gel. The column was eluted with 25 mL of hexane and 25 mL of hexane/dichloromethane (1:1, v/v) as fraction 1 (F1), followed by 25 mL dichloromethane/ acetone (7:3, v/v) as fraction 2 (F2). Fraction 1 was collected for analysis of PBDEs, while F2 was collected for OPEs. After evaporation under a gentle stream of nitrogen, the two fractions were spiked with internal standards (perylene-d12 and BDE-181 for F1; tris(2-chloroethyl) phosphate-d12 (TCEPd12) and tris(phenyl) phosphate-d15 (TPHP-d15) for F2). The
2. MATERIALS AND METHODS Study Site and Sampling. The Palos Verdes Peninsula is a part of the Greater Los Angeles metropolitan area in Southern California that has been subject to intensive anthropogenic impacts from drastic population growth and urbanization over the last century. To the west of the peninsula, the continental margin has a narrow shelf and slope and has been the recipient of effluent discharged from many WWTPs dispersed throughout the Great Los Angeles area since the 1930s. Effluents from different treatment facilities are converged and transported through large pipes underwater before discharging at an outfall point about 1.5 km from the coast. The continuous discharge of large quantities of effluents has resulted in the accumulation of a large swath of organic matter-rich sediment layer on the ocean floor and the listing of the Palos Verdes Shelf as a Superfund site due to elevated levels of contaminants including DDT and PCBs.22,24 Three sediment cores were collected from the C1 (74 cm), C2 (76 cm), and C3 (44 cm) locations in 2007. The sampling sites are shown in Figure 1, and detailed information is provided in Table S1 of the Supporting Information (SI). In 2014, three surface sediment samples (0−5 cm) were collected at the same locations, which were used to infer more recent deposition of OPEs. At the same time, an additional 41 surface sediment samples were also collected from the PVS Superfund site surrounding the outfalls, covering a total of approximately 44 km2 (Figure S1). All samples were stored on dry ice and transferred to freezers at −20 °C in the laboratory.22 The cores were sliced at 2 cm intervals with a stainless steel saw. The individual sediment samples were transferred into glass jars and kept frozen until analysis. Detailed sampling methods are provided in SI. The impact of local urban emissions on the 44 sampling locations was assessed by their local population pressure, which was calculated by dividing the population of nearby cities over the distance between these cities and the sampling location. More details about the calculation is given in SI. 8768
DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
Article
Environmental Science & Technology
Figure 2. (a) Histories of WWTP discharge and population in California. (b−d) Temporal trends in OPE concentrations with the ratio of ∑ClOPEs/∑PBDEs in sediment cores C1, C2, and C3.
8%, 82 ± 15%, 77 ± 14%, 94 ± 6%, and 89 ± 10%, respectively. Sample concentrations were adjusted by the recoveries of surrogate standards. Measurement of Rapid Desorption Fraction. The rapid desorption fraction (Frapid) of OPEs in the historically contaminated sediments was measured by using 24 h Tenaxaided desorption.27,28 Briefly, dried sediment samples (5 g) were transferred to a 50 mL glass centrifuge tube and combined with 0.5 g Tenax beads and 20 mL of 0.2% NaN3 solution. The sample tube was closed and then mixed at 120 rpm for 24 h. After centrifugation at 3000 rpm for 20 min, the Tenax beads with the supernatant were transferred into a funnel lined with filter paper. The beads were collected on the filter paper and rinsed with deionized water, air-dried, and then transferred to a 20 mL glass scintillation vial. The Tenax beads were extracted by sonication in 10 mL of acetone/hexane (1:1, v/v) for 15 min. The extraction procedure was repeated three times, and the extracts were then combined. The extract was concentrated under nitrogen to near dryness and then reconstituted in 1.0 mL hexane, followed by addition of known amounts of internal standards.27 The amounts of OPEs desorbed from the sediment over the 24 h were divided over the initial amounts of OPEs in the sediment sample to yield the rapid desorption fraction Frapid.
pretreatment procedure for OPE metabolites was carried out separately.15 Briefly, sediment samples (20 g dry weight) were spiked with diphenyl phosphate-d10 (DPHP-d10) and then extracted with 20 mL of methanol/dichloromethane (4:1, v/v) three times. The extract was cleaned by passing it through a preconditioned Florisil cartridge, evaporated to near dryness, and redissolved in 0.5 mL methanol. More details of these methods are given in SI. Instrumental Analysis. The analysis of OPEs was carried out on a Varian 3800 gas chromatograph coupled to a Varian 1200 triple quadrupole mass spectrometer operated in the electron ionization mode.7 The target analytes included tris(methylphenyl) phosphate (TMPP), tris(ethyl) phosphate (TEP), TPHP, TCEP, tris(2-butoxyethyl) phosphate (TBOEP), TCIPP, TDCIPP, tris(2-ethylhexyl) phosphate (TEHP), tris(isobutyl) phosphate (TIBP), and TNBP. The quantification of two OPE metabolites, bis(2-butoxyethyl) phosphate (BBOEP) and bis(1,3-dichloro-2-propyl) phosphate (BDCIPP), was performed on a Waters ACQUITY ultraperformance liquid chromatography-tandem quadrupole mass spectrometer.15 Details regarding the instrumental analyses for OPEs, OPE metabolites, and PBDEs, chemicals used in this study, sediment characteristics, and statistical methods are given in SI. Quality Control and Quality Assurance. The procedural blanks (n = 6) and spike recovery samples (n = 8) were sediment samples from the 70−80 cm depth of a sediment core collected in 2016 from the Salton Sea in Southern California. A procedural blank or a recovery sample was included in every batch of 10 samples. The only compound detected in the procedural blanks was TPHP, and its method detection limit (MDL) was calculated as the average value of the blanks plus three times the standard deviation of the blanks. For the analytes not detected in the blanks, MDLs were calculated as three times the signal-to-noise ratio. The MDLs for OPEs and PBDEs range from 0.1 to 1.0 ng/g and 0.02 to 1.25 ng/g, respectively. Concentrations lower than MDLs were defined as nondetected (N.D.). Each recovery sample (10 g) was freeze-dried and spiked with 10 OPEs at 1 μg each and 8 PBDEs at 100 ng each, and these were then equilibrated for 2 weeks before extraction. The spiked standard recoveries for OPEs and PBDEs were 79 ± 18% and 87 ± 15%, respectively. In all field samples, the surrogate recoveries of TNBP-d27, TDCIPP-d15, DPHP-d10, BDE-77, and BDE-166 were 74 ±
3. RESULTS AND DISCUSSION OPE Levels and Patterns in Sediment Cores. All the target OPE compounds in sediment cores from the PVS were detected at frequencies ranging from 70.1% to 97.9% (Table S2 in SI). The total concentration of the 10 target OPEs varied from 0.68 to 1064 ng/g (dry weight), with a mean of 197 ± 252 ng/g. These values were an order of magnitude higher than those reported for the sediments from the Great Lakes,6 where the main input of OPEs was assumed to be from nonpoint source urban emissions. The PVS is located adjacent to a highly urbanized and populated region in the Southern California, and it receives approximately 946 million liters of treated effluents daily through large underwater pipes.26 Thus, the high level of OPEs in the PVS sediments was likely associated with WWTP discharge and local urban emissions. Since OPEs were used as flame retardants in modest amounts as far back as the 1960s,2 the study area has likely been exposed to OPE contamination for many decades. Although a similar use history was reported for the classical flame retardant 8769
DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
Article
Environmental Science & Technology
Figure 3. Temporal trends derived from sediment core C1 showing (a) TIBP concentration, (b) TDCIPP concentration, (c) TCIPP and TCEP concentrations, (d) TPHP concentration, and (e) the relative abundance of (TMPP + TEHP).
PBDEs,17 they were found at significantly lower levels (from N.D. to 32.5 ng/g, Table S2) in the same sediment cores than OPEs. The difference in the sediment concentrations between OPEs and PBDEs may be attributed to the less efficient removal of OPEs as compared to PBDEs at WWTPs. A recent study at a WWTP in New York showed that TMPP, TBOEP, and TEHP levels were reduced by approximately 69%, while the removal efficiencies for the other OPE congeners were less than 38%.12 In contrast, North (2004)29 reported that 96% of the PBDEs entering a WWTP in Palo Alto, California, was removed. The large variability in removal efficiencies between OPEs and PBDEs may be attributed to their different physicochemical properties. For organic contaminants entering WWTPs, sorption to the solid phase is an important mechanism of removal, and the removal efficiencies are therefore often proportional to their hydrophobicity (log Koc).12,29 In addition, environmental stability of OPEs was known to depend on the ester linkage as well,12,30 which may further influence their removal at WWTPs. A summary of individual compound concentrations is presented in Table S2, and their composition profiles are shown in Figure 2b−d. With a close proximity to the effluent outfall, the sediment profile of Core C1 may closely reflect the compositions of OPEs released from WWTP effluents from the Great Los Angeles region. In Core C1, the most abundant congeners, TMPP and TPHP, collectively contributed to 55.3% of the ΣOPEs, followed by TEP, TEHP, and TBOEP, with contributions ranging from 11.0% to 14.2%. Three ClOPEs (TDCIPP, TCIPP, and TCEP) and TIBP had almost negligible contributions ( 0.05) among their individual concentrations in sediment. This was possibly attributed to the different levels of these congeners in waste products and congener-specific processes during wastewater treatment. In the case of TPHP, it is commonly incorporated as an additive flame retardant in electronics, and its production and use may have increased due to the phase-out of legacy flame retardant PBDEs.38 This possibly resulted in the TPHP concentration (7360 ng/g) in house dust in the U.S. that was as high as the PBDE concentration and also similar distribution patterns in house dust.35 In North America, about 36.0% of the PBDE stock was used in electrical and electronic equipment.3 Like PBDEs, the high levels of TPHP (up to 2.0%) were also detected in some electronic products and related wastes, possibly leading to a significant emission of TPHP after wastewater treatment.38,39 Recent technological advancement and rapid demand changes likely contributed to more frequent acquisition and discard of electronic products. In the U.S., the sales of electronic products (including cell phones, TVs, and computers) increased about 5 times or from 450 to 2270 thousand tons in weight, from 1980 to 2010. Due to the everincreasing replacement of TVs and computers, their wastes may be estimated to have increased about 8 and 9-fold between 1990 and 2010, respectively (Figure 3d).40 Thus, the recent increase in TPHP levels in the PVS sediments may be partially related to the waste emissions from electrical and electronic equipment, although there is no direct evidence for this connection. On the basis of the trends of human exposure levels in the U.S., Hoffman et al. (2017)41 suggested the possibility that TPHP use decreased in recent years. This may explain the relative low levels of TPHP in the surficial sediments. The decrease in the relative abundance of TEHP and TMPP over time may be a result of improved wastewater treatment (Figures 3e and S5);26,42 WWTPs were found to exhibit the highest removals for TEHP and TMPP among OPEs.12 Organophosphate triester metabolism by microorganisms was recently found to occur in solid matrices.12,15 It is likely that this process also took place in the PVS sediments. In sediment Core C1, two potential diester degradation products, i.e., BBOEP and BDCIPP, were quantified, with detection frequencies of 62.2% and 35.1%, respectively. Within the layers between 10 and 30 cm (representing 8−23 years of sedimentation), the values of BBOEP/(BBOEP+TBOEP) increased from 8.4% to 24.1% along the sediment depth. This highlights the likelihood that TBOEP was gradually degraded to BBOEP over time (Figure 4). The ratios of BDCIPP/(BDCIPP+TDCIPP) remained in a narrow range between 15 and 25%, with no clear temporal trend. One possible reason for this observation was the significantly longer half-lives of TDCIPP and BDCIPP in relation to the time
WWTP pathway. However, the levels of ∑OPEs increased toward the upper sediment layers after the 1990s, even though that the treatment capacity was continuously improved during the same period. The second peak in concentration seen in Core C1 was 1064 ng/g (Figure 2b), which was about four times higher than levels recorded in the 1980s. The ∑OPEs measured in the 2014 surface sediments were similar or even higher than the topmost sediment layer of the sediment cores, which validated the increased deposition of OPEs in the PVS in recent years. With the gradual restrictions of PBDE technical products,31 the recent increase in ∑OPEs may have been spurred by population growth and associated high demand for alternative flame retardants.16,32 Studies analyzing flame retardants in consumer products also showed a significant increase in Cl-OPEs following the phaseout of PBDEs,9,33,34 which was consistent with the tripled production volume of Cl-OPEs from the 1980s to 2012.32 The temporal variation in the ratio of ∑Cl-OPEs and ∑PBDEs was further calculated for the sediment cores in this study (Figure 2b−d). In the early 1970s, PBDEs were just introduced into the market while Cl-OPEs had already been used for over a decade,17,35 and this use history likely resulted in the relatively high ratios of ∑Cl-OPEs/∑PBDEs, with the peak value of 3.94 in Core C1. From 1970s to 1990s, a decrease in the ratio of ∑Cl-OPEs/∑PBDEs and an increase in ∑PBDEs (Figure S2) were observed, which coincided with the fact that PBDEs became the most widely used flame retardant additives in the world.18 Subsequently, likely due to the leveling off of domestic consumption of PBDEs,17,18 ∑PBDEs in the PVS sediments remained stable or even decreased slightly from 1990s to 2010s (Figure S2). Correspondingly, the ratio of ∑Cl-OPEs/∑PBDEs started to increase after the minima in the mid-1990s. In Core C1, the ratio of ∑Cl-OPEs/∑PBDEs increased from 0.62 in 1993 to 1.44 in 2014. This trend likely reflected the shift in the flame retardant market from PBDEs to OPEs in developed countries such as the U.S. Congener-specific temporal trends could also be observed in these cores. The congeners that showed considerable associations with urban sources, i.e., TIBP, TDCIPP, and TCIPP, generally displayed exponential increases in concentrations over time (Figures 3a,b and S3). These trends were consistent with population growth in the region. A nationwide survey of OPEs in the U.S. showed that the WWTPs serving greater populations generally contained higher levels of OPEs in their effluents.36 For the emerging contaminant perfluorooctanoate, concentrations in WWTP discharge were also found to be correlated closely with population density.37 Southern California, particularly the greater Los Angeles metropolitan area, is an extremely urbanized region. The overall population in California increased from 20 million in 1970 to 37.4 million in 2010 (Figure 2a). The demand for OPE-containing products may be expected to increase proportionally with the local resident population as reflected in the types and quantities of consumer products that each individual acquires and consumes. Because TIBP, TDCIPP, and TCIPP have been used as plasticizer and flame retardant additives in construction materials and furniture,1,14,34 their diffusion from host materials may result in continuous environmental release in urbanized regions. Likely due to contributions of urban emissions to TIBP, TDCIPP, and TCIPP in the PVS sediment, the temporal trends of these three congeners exhibited continuous increases over the last 8771
DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
Article
Environmental Science & Technology
reflecting the rapid or labile fraction. Accordingly, the sorbed OPEs may be characterized by slow (Fslow) and rapid desorption fractions (Frapid). It was assumed that only Frapid participates in phase equilibrium with sediment porewater,44 suggesting a correlation between Frapid and diffusion in porewater. The contribution from Fslow should be negligibly small.45 Thus, in addition to the rapid sedimentation rate, the reduced Frapid values of sediment OPEs following wastewater treatment may be responsible for their limited vertical migration in sediment. In this study, Tenax-aided desorption was used to estimate the Frapid of TMPP, TEP, TPHP, TBOEP, TDCIPP, and TEHP in Core C1 (Table S4). The highest and lowest Frapid values were observed for TDCIPP and TBOEP, respectively, which may be associated with their chemical structures,46 such as the ester linkage (Table S4). Subsequently, these Frapid values were applied to predict the fraction of dissolved OPEs (f iw) and their downward transport.45 Following a previous method,47 f iw was calculated using a modified equation combining the solid-to-water ratio of sediments (rsw) and the solid−water partition coefficient (Kd) f iw = (1 + rsw × Kd/ Frapid)−1. The Kd value was calculated by multiplying Koc with the organic carbon fraction of the sediment (foc). As described by White et al.,48 the diffusivity in water (Diw) may be estimated based on the size of the target compound and the solution viscosity at a given temperature. The effective diffusivity coefficient (Dieff) was calculated by multiplying f iw with Diw and dividing by the tortuosity. The Dieff values varied among the sedimentation periods. Figure 5a shows the Dieff values before and after calibration by Frapid values in Core C1, and the differences suggested that the calibrated diffusion values were greatly reduced in comparison to what was expected. While the peak emissions from WWTPs in the 1970s likely resulted in the first peak of total OPE concentrations in the PVS sediments, the first peaks in TEP and TDCIPP were skewed downward as compared to the layer dated to 1970 (Figure 5b). This may be attributed to their downward
Figure 4. Relative abundances of detected diester metabolites normalized to the sum of diesters and the parent compound in sediment core C1. Gray and purple shadows represent 95% confidence intervals.
interval of sedimentation (Figure 4).15,30 Although degradation of OPEs was perceptible in sediments, pronounced influences were not observed for the concentration profiles. In Cores C2 and C3, BBOEP and BDCIPP were not detected, possibly due to the relatively low levels of their parent compounds at these locations. Influence of Downward Diffusion. Some OPEs may have migrated downward within the sediment profile via diffusion of the dissolved and colloidal phases in porewater, which may affect the apparent sedimentary record.6,43 In this case, there was no significant fractionation observed for most of the target compounds considered in this study, except for TEP (Figure S6). Prior to being available for downward diffusion within the sediment profile, OPE congeners must desorb from the sediment phase into the sediment porewater. However, because some OPEs were trapped in micropores or adsorbed in the dense region of organic matter, desorption of OPEs from a sediment may be progressively slower, with Frapid
Figure 5. (a) Comparison of effective diffusions in sediment core C1 before and after calibration with Frapid. (b) General observations of the measured data and estimated diffusion distances for the first peaks in TEP and TDCIPP concentrations (arrows). 8772
DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
Article
Environmental Science & Technology
transport of sediment-associated OPEs due to, e.g., advective movement caused by ocean currents. In conclusion, on the basis of distribution and congener profiles of OPEs in the dated sediment cores from the PVS as well as surficial sediment samples collected in 2014, this study presented a highly time-resolved historical record of OPEs. The covered time span and monitoring data enabled an illustration of the important roles of WWTP effluent discharge, regulatory actions, and population growth on the contamination intensity of OPEs. In a previous study,50 the PVS sediment was identified as a contaminant source of chlorinated pesticides to the Southern California Bight. With the repeated processes of sediment resuspension and current advection, the coastal marine sediments impacted by WWTP emissions serve as an important source for the regional loads of such flame retardants as well. As there are a great number of urban centers along coastlines in the world, the finding that WWTP emissions are a significant source of OPEs in oceans may be globally valid. However, knowledge regarding transport of OPEs away from the near-shore regions and attenuation processes in the marine environment is very limited. Therefore, additional research on the transport processes and abiotic/ biotic transformations of OPEs and other flame retardants in marine environments is warranted.
diffusion over the decades-long interval from 1970 to 2007. By applying the value of Dieff, the distance of downward diffusion (x) for the first concentration peak over a known time interval (t) may be estimated using the following equation, assuming that the peak follows a Gaussian distribution43 x = (0.5Dieff × t)1/2. Interestingly, the estimated distances of downward movement for TEP (6.2 cm) and TDCIPP (4.0 cm) were in close agreement with the observations reflected in the measured data (Figure 5b). For TMPP, TPHP, TBOEP, and TEHP, the estimated peak skewing was less than the sampling interval thickness (2 cm). Therefore, diffusion possibly accounted for some of the penetration of TEP and TDCIPP profiles to depths older than the actual emission. It is likely that OPEs were introduced to the sediment as attachment to organic matter-rich solids; this would imply reduced mobility and long persistence of OPEs in the sediment once deposited onto the ocean floor. The above observations imply that the concentration record of OPEs in marine sediment with WWTP effluents as the primary source may offer a generally accurate emission history of OPEs in the region. WWTP Input to OPE Inventory. Since OPE levels in the PVS sediments were observed to increase rapidly over the past decade, it is important to estimate the annual inventory of OPEs from WWTP emissions. Using the average sedimentation rate, and the OPE concentrations in the top three sediment layers in Core C1, the OPE inventory in the sediment in the area surrounding the wastewater effluent outfall point (4.5 km2) was estimated, ranging from 59.2 to 112 kg/yr, with a mean of 89.5 kg/yr. A detailed calculation method is provided in the SI. With the same method, the annual increase in the sediment inventory was also estimated for a 4.5 km2 area surrounding Core C3, which is far from the outfall and experiences highly local population pressure as well. Its inventory was calculated to be 2.01 kg/yr, which may serve as a reference for the contribution of urban emissions to sediment OPEs. The comparison of OPE inventories between C1 and C3 strongly suggested that the sediment inventory near the outfall derived dominantly from the WWTP emissions. The annual discharge amount of wastewater from Southern California into the ocean (1.78 trillion liters) was five times higher than that to the PVS (0.35 trillion liters).49 Assuming the released OPEs from California as a whole are at the same level as that seen at the PVS, the estimated OPE inventory in the marine sediment along the Southern California coast would be about 455 kg/yr from WWTP emissions. A recent study by Wang et al. (2019)36 estimated the total emissions of OPEs in the U.S. through WWTPs, which could be extrapolated to annual per capital emission for 10 target OPEs (261 mg/person/yr). By multiplying the population in Southern California (23.8 million) with the per capital emissions in Wang et al.,36 the annual loading of 10 OPEs in this region was calculated to be 6.22 tons. From the annual increment of sediment OPEs (455 kg), it appeared that about 7.3% of total loading may have deposited in the coastal sediment. Of the congeners, it is evident that the three ClOPEs exhibited a low ratio between the estimated sediment inventories and WWTP emissions, e.g., 1.4% for TCEP and 0.5% for both TCIPP and TDCIPP. This may be attributed to the relatively low hydrophobicity of these congeners (log Koc < 3) and their high Frapid in the PVS sediments. However, these ratios were likely an underestimate for the PVS, because the entire PVS area with effluent-impacted sediment is approximately 44 km2, and the estimation did not consider off-site
■
ASSOCIATED CONTENT
S Supporting Information *
The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.9b01713. Chemicals; instrumental analysis; sediment characteristics; statistical methods; population pressure assessment; estimation for annual OPE inventory; information on sampling sites; the descriptive statistics of OPE and PBDE compounds; correlations among concentrations in surface sediments and environmental parameters; mean values of Frapid for selected OPEs in each decade; spatial distribution of 10 OPE in surface sediments in 2014; temporal trends in total PBDE concentration; temporal trends in TDCIPP and TIBP concentrations; temporal trends in TCEP and TCIPP concentrations; relative abundance of (TMPP + TEHP) in sediment cores; relative abundance of TEP in sediment cores (PDF)
■
AUTHOR INFORMATION
Corresponding Author
*E-mail:
[email protected]. ORCID
Jay Gan: 0000-0002-7137-4988 Notes
The authors declare no competing financial interest.
■
ACKNOWLEDGMENTS This study was supported by the Superfund Research Program of the National Institute of Environmental Health Science via grant 5R01ES020921. J. Li was a visiting Ph.D. student with scholarship support from the Chinese Scholarship Council. The authors were grateful to Los Angeles County Sanitation District for providing the archived sediment samples. 8773
DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
Article
Environmental Science & Technology
■
(13) UNEP. Technical Review of the Implications of Recycling Commercial Pentabromodiphenyl Ether and Commercial Octabromodiphenyl Ether. 2010. (14) USEPA. The Chemical Data Reporting under the Toxic Substances Control Act (TSCA) Contains Chemical Physical Description and Chemical Use Categories. http://www.epa.gov/ chemical-data-reporting. (15) Fu, L. F.; Du, B. B.; Wang, F.; Lam, J. C. W.; Zeng, L. X.; Zeng, E. Y. Organophosphate Triesters and Diester Degradation Products in Municipal Sludge from Wastewater Treatment Plants in China: Spatial Patterns and Ecological Implications. Environ. Sci. Technol. 2017, 51 (23), 13614−13623. (16) Guo, J. H.; Romanak, K.; Westenbroek, S.; Hites, R. A.; Venier, M. Current-Use Flame Retardants in the Water of Lake Michigan Tributaries. Environ. Sci. Technol. 2017, 51 (17), 9960−9969. (17) Hale, R. C.; La Guardia, M. J.; Harvey, E.; Chen, D.; Mainor, T. M.; Luellen, D. R.; Hundal, L. S. Polybrominated Diphenyl Ethers in U.S. Sewage Sludges and Biosolids: Temporal and Geographical Trends and Uptake by Corn Following Land Application. Environ. Sci. Technol. 2012, 46 (4), 2055−2063. (18) Wei, H.; Aziz-Schwanbeck, A. C.; Zou, Y. H.; Corcoran, M. B.; Poghosyan, A.; Li, A.; Rockne, K. J.; Christensen, E. R.; Sturchio, N. C. Polybromodiphenyl Ethers and Decabromodiphenyl Ethane in Aquatic Sediments from Southern and Eastern Arkansas, United States. Environ. Sci. Technol. 2012, 46 (15), 8017−8024. (19) Cui, X. Y.; Mayer, P.; Gan, J. Methods to Assess Bioavailability of Hydrophobic Organic Contaminants: Principles, Operations, and Limitations. Environ. Pollut. 2013, 172, 223−234. (20) Meng, X. Z.; Xiang, N.; Yu, L. H.; Zhang, J. Y.; Chen, L.; Dai, X. H. Exploring the Bioaccessibility of Polybrominated Diphenyl Ethers (PBDEs) in Sewage Sludge. Environ. Pollut. 2015, 207, 1−5. (21) McClellan, K.; Halden, R. U. Pharmaceuticals and Personal Care Products in Archived US Biosolids from the 2001 Epa National Sewage Sludge Survey. Water Res. 2010, 44 (2), 658−668. (22) Liao, C. Y.; Taylor, A. R.; Kenney, W. F.; Schlenk, D.; Gan, J. Historical Record and Fluxes of DDTs at the Palos Verdes Shelf Superfund Site, California. Sci. Total Environ. 2017, 581, 697−704. (23) Fu, Q. G.; Wu, X. Q.; Ye, Q. F.; Ernst, F.; Gan, J. Biosolids Inhibit Bioavailability and Plant Uptake of Triclosan and Triclocarban. Water Res. 2016, 102, 117−124. (24) Fernandez, L. A.; Lao, W. J.; Maruya, K. A.; Burgess, R. M. Calculating the Diffusive Flux of Persistent Organic Pollutants between Sediments and the Water Column on the Palos Verdes Shelf Superfund Site Using Polymeric Passive Samplers. Environ. Sci. Technol. 2014, 48 (7), 3925−3934. (25) Santschi, P. H.; Guo, L. D.; Asbill, S.; Allison, M.; Kepple, A. B.; Wen, L. S. Accumulation Rates and Sources of Sediments and Organic Carbon on the Palos Verdes Shelf Based on Radioisotopic Tracers (Cs-137, Pu-239,Pu-240, Pb-210, Th-234, U-238 and C-14). Mar. Chem. 2001, 73 (2), 125−152. (26) LACSD. Palos Verdes Ocean Monitoring Annual Report. https://www.lacsd.org. (27) Jia, F.; Liao, C. Y.; Xue, J. Y.; Taylor, A.; Gan, J. Comparing Different Methods for Assessing Contaminant Bioavailability During Sediment Remediation. Sci. Total Environ. 2016, 573, 270−277. (28) He, H.; Gao, Z. Q.; Zhu, D. L.; Guo, J. H.; Yang, S. G.; Li, S. Y.; Zhang, L. M.; Sun, C. Assessing Bioaccessibility and Bioavailability of Chlorinated Organophosphorus Flame Retardants in Sediments. Chemosphere 2017, 189, 239−246. (29) North, K. D. Tracking Polybrominated Diphenyl Ether Releases in a Wastewater Treatment Plant Effluent, Palo Alto, California. Environ. Sci. Technol. 2004, 38 (17), 4484−4488. (30) Su, G. Y.; Letcher, R. J.; Yu, H. X. Organophosphate Flame Retardants and Plasticizers in Aqueous Solution: pH-Dependent Hydrolysis, Kinetics, and Pathways. Environ. Sci. Technol. 2016, 50 (15), 8103−8111. (31) UNEP. Proposal to List Decabromodiphenyl Ether (Commercial Mixture) in Annexes A,B and/or C to the Stockholm Convention on Persistent Organic Pollutants. 2013.
ABBREVIATIONS PVS Palos Verdes Shelf WWTP wastewater treatment plant OPEs organophosphate esters TMPP tris(methylphenyl) phosphate TEP tris(ethyl) phosphate TPHP tris(phenyl) phosphate TCEP tris(2-chloroethyl) phosphate TBOEP tris(2-butoxyethyl) phosphate TCIPP tris(2,3-dichloro-n-propyl) phosphate TDCIPP tris(1,3-dichloro-2-propyl) phosphate TEHP tris(2-ethylhexyl) phosphate TIBP tris(isobutyl) phosphate TNBP tri-n-butyl phosphate BBOEP bis(2-butoxyethyl) phosphate BDCIPP bis(1,3-dichloro-2-propyl) phosphate PBDE polybrominated diphenyl ether MDL method detection limit
■
REFERENCES
(1) van der Veen, I.; de Boer, J. Phosphorus Flame Retardants: Properties, Production, Environmental Occurrence, Toxicity and Analysis. Chemosphere 2012, 88 (10), 1119−1153. (2) Greaves, A. K.; Letcher, R. J. A Review of Organophosphate Esters in the Environment from Biological Effects to Distribution and Fate. Bull. Environ. Contam. Toxicol. 2017, 98 (1), 2−7. (3) Abbasi, G.; Buser, A. M.; Soehl, A.; Murray, M. W.; Diamond, M. L. Stocks and Flows of PBDEs in Products from Use to Waste in the US and Canada from 1970 to 2020. Environ. Sci. Technol. 2015, 49 (3), 1521−1528. (4) Hoffman, K.; Stapleton, H. M.; Lorenzo, A.; Butt, C. M.; Adair, L.; Herring, A. H.; Daniels, J. L. Prenatal Exposure to Organophosphates and Associations with Birthweight and Gestational Length. Environ. Int. 2018, 116, 248−254. (5) Meeker, J. D.; Stapleton, H. M. House Dust Concentrations of Organophosphate Flame Retardants in Relation to Hormone Levels and Semen Quality Parameters. Environ. Health Perspect. 2010, 118 (3), 318−323. (6) Cao, D. D.; Guo, J. H.; Wang, Y. W.; Li, Z. N.; Liang, K.; Corcoran, M. B.; Hosseini, S.; Bonina, S. M. C.; Rockne, K. J.; Sturchio, N. C.; Giesy, J. P.; Liu, J. F.; Li, A.; Jiang, G. B. Organophosphate Esters in Sediment of the Great Lakes. Environ. Sci. Technol. 2017, 51 (3), 1441−1449. (7) Salamova, A.; Hermanson, M. H.; Hites, R. A. Organophosphate and Halogenated Flame Retardants in Atmospheric Particles from a European Arctic Site. Environ. Sci. Technol. 2014, 48 (11), 6133− 6140. (8) Ma, Y. X.; Xie, Z. Y.; Lohmann, R.; Mi, W. Y.; Gao, G. P. Organophosphate Ester Flame Retardants and Plasticizers in Ocean Sediments from the North Pacific to the Arctic Ocean. Environ. Sci. Technol. 2017, 51 (7), 3809−3815. (9) Schreder, E. D.; La Guardia, M. J. Flame Retardant Transfers from U.S. Households (Dust and Laundry Wastewater) to the Aquatic Environment. Environ. Sci. Technol. 2014, 48 (19), 11575− 11583. (10) Salamova, A.; Peverly, A. A.; Venier, M.; Hites, R. A. Spatial and Temporal Trends of Particle Phase Organophosphate Ester Concentrations in the Atmosphere of the Great Lakes. Environ. Sci. Technol. 2016, 50 (24), 13249−13255. (11) Cristale, J.; Vazquez, A. G.; Barata, C.; Lacorte, S. Priority and Emerging Flame Retardants in Rivers: Occurrence in Water and Sediment, Daphnia Magna Toxicity and Risk Assessment. Environ. Int. 2013, 59, 232−243. (12) Kim, U. J.; Oh, J. K.; Kannan, K. Occurrence, Removal, and Environmental Emission of Organophosphate Flame Retardants/ Plasticizers in a Wastewater Treatment Plant in New York State. Environ. Sci. Technol. 2017, 51 (14), 7872−7880. 8774
DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
Article
Environmental Science & Technology (32) Schreder, E. D.; Uding, N.; La Guardia, M. J. Inhalation a Significant Exposure Route for Chlorinated Organophosphate Flame Retardants. Chemosphere 2016, 150, 499−504. (33) Stapleton, H. M.; Sharma, S.; Getzinger, G.; Ferguson, P. L.; Gabriel, M.; Webster, T. F.; Blum, A. Novel and High Volume Use Flame Retardants in US Couches Reflective of the 2005 Pentabde Phase Out. Environ. Sci. Technol. 2012, 46 (24), 13432−13439. (34) Dodson, R. E.; Perovich, L. J.; Covaci, A.; Van den Eede, N.; Ionas, A. C.; Dirtu, A. C.; Brody, J. G.; Rudel, R. A. After the PBDE Phase-Out: A Broad Suite of Flame Retardants in Repeat House Dust Samples from California. Environ. Sci. Technol. 2012, 46 (24), 13056− 13066. (35) Stapleton, H. M.; Klosterhaus, S.; Eagle, S.; Fuh, J.; Meeker, J. D.; Blum, A.; Webster, T. F. Detection of Organophosphate Flame Retardants in Furniture Foam and US House Dust. Environ. Sci. Technol. 2009, 43 (19), 7490−7495. (36) Wang, Y.; Kannan, P.; Halden, R. U.; Kannan, K. A Nationwide Survey of 31 Organophosphate Esters in Sewage Sludge from the United States. Sci. Total Environ. 2019, 655, 446−453. (37) Zhang, W.; Zhang, Y. T.; Taniyasu, S.; Yeung, L. W. Y.; Lam, P. K. S.; Wang, J. S.; Li, X. H.; Yamashita, N.; Dai, J. Y. Distribution and Fate of Perfluoroalkyl Substances in Municipal Wastewater Treatment Plants in Economically Developed Areas of China. Environ. Pollut. 2013, 176, 10−17. (38) Iqbal, M.; Syed, J. H.; Breivik, K.; Chaudhry, M. J. I.; Li, J.; Zhang, G.; Malik, R. N. E-Waste Driven Pollution in Pakistan: The First Evidence of Environmental and Human Exposure to Flame Retardants (FRs) in Karachi City. Environ. Sci. Technol. 2017, 51 (23), 13895−13905. (39) Ballesteros-Gomez, A.; Brandsma, S. H.; de Boer, J.; Leonards, P. E. G. Analysis of Two Alternative Organophosphorus Flame Retardants in Electronic and Plastic Consumer Products: Resorcinol Bis-(Diphenylphosphate) (PBDPP) and Bisphenol A Bis (Diphenylphosphate) (BPA-BDPP). Chemosphere 2014, 116, 10−14. (40) USEPA. Office of Resource Conservation and Recovery. Electronics Waste Management in the United States through 2009. 2011. (41) Hoffman, K.; Butt, C. M.; Webster, T. F.; Preston, E. V.; Hammel, S. C.; Makey, C.; Lorenzo, A. M.; Cooper, E. M.; Carignan, C.; Meeker, J. D.; Hauser, R.; Soubry, A.; Murphy, S. K.; Price, T. M.; Hoyo, C.; Mendelsohn, E.; Congleton, J.; Daniels, J. L.; Stapleton, H. M. Temporal Trends in Exposure to Organophosphate Flame Retardants in the United States. Environ. Sci. Technol. Lett. 2017, 4 (3), 112−118. (42) Stein, E. D.; Cadien, D. B. Ecosystem Response to Regulatory and Management Actions: The Southern California Experience in Long-Term Monitoring. Mar. Pollut. Bull. 2009, 59 (4−7), 91−100. (43) Wong, C. S.; Sanders, G.; Engstrom, D. R.; Long, D. T.; Swackhamer, D. L.; Eisenreich, S. J. Accumulation, Inventory, and Diagenesis of Chlorinated Hydrocarbons in Lake-Ontario Sediments. Environ. Sci. Technol. 1995, 29 (10), 2661−2672. (44) You, J.; Pehkonen, S.; Landrum, P. F.; Lydy, M. J. Desorption of Hydrophobic Compounds from Laboratory-Spiked Sediments Measured by Tenax Absorbent and Matrix Solid-Phase Microextraction. Environ. Sci. Environ. Sci. Technol. 2007, 41 (16), 5672− 5678. (45) Kraaij, R.; Mayer, P.; Busser, F. J. M.; van het Bolscher, M.; Seinen, W.; Tolls, J.; Belfroid, A. C. Measured Pore-Water Concentrations Make Equilibrium Partitioning Work - A Data Analysis. Environ. Sci. Technol. 2003, 37 (2), 268−274. (46) Kukkonen, J. V. K.; Landrum, P. F.; Mitra, S.; Gossiaux, D. C.; Gunnarsson, J.; Weston, D. Sediment Characteristics Affecting Desorption Kinetics of Select PAH and Pcb Congeners for Seven Laboratory Spiked Sediments. Environ. Sci. Technol. 2003, 37 (20), 4656−4663. (47) Gschwend, P. M.; Wu, S. C. On the Constancy of Sediment Water Partition-Coefficients of Hydrophobic Organic Pollutants. Environ. Sci. Technol. 1985, 19 (1), 90−96.
(48) White, H. K.; Xu, L.; Lima, A. L. C.; Eglinton, T. I.; Reddy, C. M. Abundance, Composition, and Vertical Transport of PAHs in Marsh Sediments. Environ. Sci. Technol. 2005, 39 (21), 8273−8280. (49) Rodman, K.; Cervania, A.; Budig-Markin, V.; Schermesser, C.; Rogers, O.; Martinez, J.; King, J.; Hassett, P.; Burns, J.; Gonzales, M.; Folkerts, A.; Duin, P.; Virgil, A.; Aldrete, M.; Lagasca, A.; InfanzonMarin, A.; Aitchison, J.; White, D.; Boutros, B.; Ortega, S.; Davis, B.; Tran, V.; Achilli, A. Coastal California Wastewater Effluent as a Resource for Seawater Desalination Brine Commingling. Water 2018, 10 (3), 322. (50) Zeng, E. Y.; Tsukada, D.; Diehl, D. W.; Peng, J.; Schiff, K.; Noblet, J. A.; Maruya, K. A. Distribution and Mass Inventory of Total Dichlorodiphenyldichlorothylene in the Water Column of the Southern California Bight. Environ. Sci. Technol. 2005, 39 (21), 8170−8176.
8775
DOI: 10.1021/acs.est.9b01713 Environ. Sci. Technol. 2019, 53, 8767−8775
