Science of the Total Environment 524–525 (2015) 23–31

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Perfluoroalkyl acids in municipal landfill leachates from China: Occurrence, fate during leachate treatment and potential impact on groundwater Hong Yan a, Ian T. Cousins b, Chaojie Zhang a,⁎, Qi Zhou a a b

State Key Laboratory of Pollution Control and Resources Reuse, College of Environmental Science and Engineering, Tongji University, Shanghai 200092, China Department of Applied Environmental Science (ITM), Stockholm University, Svante Arrhenius väg 8, Stockholm, Sweden

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• This is the first study on PFAA fate during treatment of landfill leachates in China. • Replacement substance PFBS was one of the most abundant PFAAs in Chinese leachate. • Mass flow of individual PFAAs was substance and treatment system specific. • Landfill leachate is a significant source of PFAAs to groundwater.

a r t i c l e

i n f o

Article history: Received 24 January 2015 Received in revised form 23 March 2015 Accepted 24 March 2015 Available online 15 April 2015 Editor: Adrian Covaci Keywords: Perfluoroalkyl acids Leachate Treatment Mass flow China

a b s t r a c t Raw and treated landfill leachate samples were collected from 5 municipal landfill sites in China to measure the concentrations and contamination profile of perfluoroalkyl acids (PFAAs) in leachate during different steps of treatment. The total concentration of PFAAs (∑PFAAs) ranged from 7280 to 292,000 ng L−1 in raw leachate and from 98.4 to 282,000 ng L−1 in treated leachate. The dominant compounds measured were PFOA (mean contribution 28.8% and 36.8% in raw and treated leachate, respectively) and PFBS (26.1% and 40.8% in raw and treated leachate, respectively). A calculation of mass flows during the leachate treatment processes showed that the fate of individual PFAAs was substance and treatment-specific. The Chinese national leakage of ∑PFAAs to groundwater from landfill leachate was estimated to be 3110 kg year−1, which is a significant environmental release that is potentially threatening the sustainable use of groundwater as a drinking water source. © 2015 Elsevier B.V. All rights reserved.

1. Introduction

⁎ Corresponding author. E-mail address: [email protected] (C. Zhang).

http://dx.doi.org/10.1016/j.scitotenv.2015.03.111 0048-9697/© 2015 Elsevier B.V. All rights reserved.

Perfluoroalkyl acids (PFAAs), which include perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkane sulfonic acids (PFSAs), are important commercial organic substances widely used since the 1950s

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H. Yan et al. / Science of the Total Environment 524–525 (2015) 23–31

(Paul et al., 2008; Prevedouros et al., 2006). Due to their high stability, surface tension lowering properties and ability to create stable foams, PFAAs have had a widespread application in consumer products and industrial processes, including metal plating and cleaning, coating formulations, fire-fighting foams, polyurethane production, inks, varnishes, vinyl polymerization, lubricants, gasoline, and oil, and water repellents for leather, paper, and textiles (Paul et al., 2008; Prevedouros et al., 2006). In the last decade, PFAAs have been detected in abiotic, biota and human samples worldwide (Giesy and Kannan, 2001; Kannan et al., 2004; Stock et al., 2007; Yamashita et al., 2008; Yang et al., 2011). The presence of long-chain PFAAs (long-chain PFAAs are PFCAs with 7 or more perfluorinated carbons and PFSAs with 6 or more perfluorinated carbons (Buck et al., 2011)) in the environment is of concern due to their persistence, bioaccumulation and adverse effects in biota and humans (Brooke et al., 2004; Conder et al., 2008; Lau et al., 2004). Short-chain PFAAs are equally persistent as the long-chain PFAAs and are expected to be more mobile, but not as bioaccumulative (Scheringer et al., 2014). Municipal solid waste (MSW) is expected to contain PFAAs derived from residential, commercial and institutional sources (Eggen et al., 2010). In China, sanitary landfill is the dominant disposal option for MSW (Idris et al., 2004). Municipal sewage sludge, which contains PFAAs (Yan et al., 2012), is also landfilled. It is therefore probable that PFAA-containing MSW will be introduced into the landfills, and subsequently PFAAs released and leached into landfill leachate with the potential for PFAA contamination of surface and groundwater in the vicinities of dumping sites. Presently, several surveys have been conducted on PFAAs in landfill leachate from North America and Europe (3M, 2001; Bossi et al., 2008; Busch et al., 2010; Huset et al., 2011; Kallenborn et al., 2004; Oliaei et al., 2006; Woldegiorgis et al., 2006). In China, only one PFAA study in leachate was conducted in Beijing (Zhang et al., 2014b). Yet, not enough data have been available across China, where landfills are distributed widely. This is the first study to report the presence of PFAAs in landfill leachates at multiple landfill sites across China. Considering that the use history of PFAAs in China is different to Europe and North America (Wang et al., 2014a), one might expect different contamination levels and patterns in Chinese landfill leachates. Leachate handling typically involves recirculation of leachate back into the landfill or treatment either on- or off-site, but the extent to which these processes reduce or contain PFAAs is unknown. Efficient removal of PFAAs during leachate treatment is expected to be a challenging task; additional expensive treatment processes (e.g. black carbon adsorption) are probably required but mostly not implemented (Appleman et al., 2014; Zhang et al., 2014a). However, the effectiveness of commonly-applied leachate treatment processes for removing PFAAs is not well-known due to the limited studies reporting the fate and mass flow of PFAAs during leachate treatment processes. It is possible that if the leachate treatment is ineffective the leachate waters could contaminate the environment surrounding a landfill including groundwater (Edil, 2003; Mor et al., 2006). More than 1000 landfill facilities operate throughout China (Jing et al., 2006) and one third of them are located in the coastal cities where groundwater is one of many sources of drinking water. Landfills pose a high potential threat of PFAA migration into the subterranean environment. An estimation of the quantity of PFAA leakage is thus needed to assess the risks associated with contaminated groundwater and to determine if additional remediation measures are necessary. The objectives of this study were: (1) to examine the levels and contamination profiles of PFAAs in landfill leachate samples collected in China, thus providing unique insights into the use history of PFAAs in China; (2) to systematically investigate the mass flow and fate of PFAAs during each step of the leachate treatment process; and (3) to estimate the national leakage of PFAAs through leachate to better understand the role landfill plays in the release of PFAAs to the subterranean environment.

2. Materials and methods 2.1. Chemicals and standards In this study, 14 PFAAs were selected as target chemicals, namely, pentafluoropropionic acid (PFPrA), perfluorobutyric acid (PFBA), perfluoropentanoic acid (PFPeA), perfluorohexanoic acid (PFHxA), perfluoroheptanoic acid (PFHpA), perfluorooctanoic acid (PFOA), perfluorononanoic acid (PFNA), perfluorodecanoic acid (PFDA), perfluoroundecanoic acid (PFUnA), perfluorododecanoic acid (PFDoA), perfluorotetradecanoic acid (PFTA), perfluoro-1-butanesulfonic acid potassium salt (PFBS), perfluorohexanesulfonic acid potassium salt (PFHxS) and perfluorooctanesulfonic acid (PFOS). Perfluoro-n[1,2,3,4-13C4]octanoic acid (MPFOA) and sodium perfluoro-1[1,2,3,4-13C4]octanesulfonate (MPFOS) were selected as internal standards. All of the analytical standards were of ≥95% purity. The detailed descriptions of chemicals are available in the Supplementary material. 2.2. Landfill sampling Raw and treated leachate samples were collected from 5 municipal landfill sites in China during the spring of 2013. Raw leachate grab samples (unfiltered, 2 L) were taken from the leachate lift station before the leachate was pumped off-site for treatment. Leachate treatment facilities at the study sites employed a two-stage process that integrated an external membrane bioreactor (MBR) unit with a post-treatment reverse osmosis (RO) or nanofiltration (NF) unit. Treated leachate samples (unfiltered, approximately 2 L of each sample type) including the bioreactor mixture, ultrafiltration (UF) effluent, NF effluent and RO effluent were taken during the treatment process as 24 h composites. All samples were collected in polyethylene (PE) bottles pre-washed with methanol. Samples were then stored at 4 °C and extracted for analysis within four weeks of sampling. PTFE (polytetrafluoroethylene)based materials were avoided throughout the sampling and analysis to avoid potential sample contamination. A description of the landfill sites, including site location, estimated amount of leachate generated per day and the treatment process used to clean the leachate, is presented in Table 1. The physical–chemical properties of the raw leachate samples (see Table 1) including pH, total organic carbon (TOC), and electrical conductivity (EC) were determined according to the U.S. EPA Method. 2.3. Sample preparation and analysis The raw and treated leachate samples were centrifuged at 11,000 ×g for 15 min to remove large particles before extraction. The amount of PFAAs absorbed onto particles (except for the bioreactor mixture) was considered low because of the small amount of particles (less than 200 mg L−1) present in the leachate. After centrifugation, the supernatants of leachate samples were spiked with internal standards prior to extraction using a solid phase extraction (SPE) method. The supernatant and solid sludge from the bioreactor mixture were analyzed separately. Sludge samples were further air-dried and extracted according to a method developed in our previous study (Li et al., 2010), which includes sonication solvent extraction, SPE and dispersive carbon sorbent cleanup. Different subsample aliquots of leachate were used for SPE depending on the sample type: an aliquot of 50 mL of raw leachate and NF/RO concentrate, 100 mL of bioreactor supernatant and 500 mL of effluent, respectively. Sludge samples were extracted by sonication using acetic acid and methanol followed by SPE. SPE was performed using Oasis® WAX cartridge (Waters, Milford, MA, USA) preconditioned with 1% ammonium hydroxide in methanol, methanol and 1% formic acid. A flow rate of 1 drop/s was maintained through the cartridges. The cartridges were then washed with 2% formic acid and were dried completely under vacuum. The target compounds were eluted in 2 mL of methanol and 3 mL of 1% ammonium hydroxide

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Table 1 Summary information for the selected landfill sites and physical–chemical properties of raw leachate. Landfill site

Location

Amount of leachate (m3 day−1)

Treatment processa

Status

NF/RO concentrate yield

pHb

Electrical conductivityb (μS cm−1)

Total organic carbonb (mg L−1)

CZ GZ NJ

Changzhou Guangzhou Nanjing

550 800 800

A/O/O-MBR + NF UASB + A/O-MBR + RO 2-stage A/O-MBR + NF

15% 20% 15%

7.33 7.95 8.63

18,690 22,200 20,400

4093 4566 2838

SH SZ

Shanghai Suzhou

800 950

A/O/O-MBR + RO 2-stage A/O-MBR + RO

Active Inactive, closed in 2004 Inactive old site, closed in 2010; active current site Active Inactive old site, closed in 2009; active current site

20% 15%

8.44 8.06

18,200 16,010

4340 2529

a b

MBR = membrane bioreactor, A/O/O = Anoxic/Oxic/Oxic, A/O = Anoxic/Oxic, UASB = Up-flow Anaerobic Sludge Blanket, RO = reverse osmosis, NF = nanofiltration. pH, electrical conductivity and total organic carbon were determined for raw leachate samples.

into a polypropylene (PP) tube. The SPE eluent was cleaned up using dispersive carbon sorbent to remove the co-eluted interfering compounds. The purified SPE eluents were then analyzed using high performance liquid chromatography-tandem mass spectrometer (HPLC-MS/MS). A detailed description of the instrumental analysis and quantitation is available in our previous study (Li et al., 2010). Briefly, aliquots of 10 μL of extracts were injected onto a 150 mm × 2.1 mm Hypersil Gold C18 column (3-μm pore size). A gradient mobile phase of methanol and 2 mM ammonium acetate aquatic solution was delivered at a flow rate of 250 μL min−1. Initial eluent conditions were 10% methanol and kept for 2 min, and the percent methanol was increased to 40% at 3 min, ramped to 100% at 12 min, held at 100% for 3 min, and then reverted to 10% at 15.5 min. For quantitative determination, the HPLC was interfaced to a Finnigan TSQ™ Quantum Access™ (Thermo Electron, San Jose, CA, USA) triple quadrupole mass spectrometer equipped with electrospray ionization (ESI) source. Electrospray negative ionization was used in the tandem mass spectrometer (MS/MS) ion source. Transitions for all ions were observed using selected reaction monitoring (SRM) mode. 2.4. Quality control and assurance Recovery experiments were performed to evaluate the precision and accuracy of the whole extraction method. Leachate samples were spiked with specific amounts of the target analytes (i.e. 500 ng L−1 for PFPrA, PFOA and PFBS, 50 ng L−1 for other PFAAs) prior to extraction. The detection limits and recoveries for the analytes in the different leachates are summarized in Table S1 in the Supplementary material. The recoveries of target compounds in real samples ranged from 66 to 111% for all types of leachate, and therefore reported concentrations were not corrected for recoveries. The precision of the entire method, represented by the relative standard deviation (RSD) of 1%–15% (n = 3), was considered acceptable. Sample blanks were routinely included in sample batches to eliminate any external source of contamination. Method blanks, prepared with Milli-Q water and treated with the same sample pre-treatment procedures as real leachate samples, were used to monitor potential contamination during sample extraction, cleanup and analysis. Method blanks were consistently below the instrumental detection limits (IDLs) for all monitored PFAAs. Instrumental blanks were used to monitor instrumental background during HPLC-MS/MS analysis. When the background contamination enhanced the signal-to-noise ratio (S/N) of instrumental blanks to N3 or reduced the S/N of the lowest calibration standard to b10, a solvent mixture consisting of 10% formic acid in 2-propanol (V/V) was run overnight through the system (Schultz et al., 2006). 3. Results and discussion 3.1. PFAA concentrations in landfill leachates The statistical summary of PFAA concentrations in raw and treated landfill leachates from China is shown in Table S2 in the Supplementary

material. Eleven of the 14 PFAA analytes were detected in raw and treated leachates with high detection frequency (36.7%–100%). PFUnA, PFDoA and PFTA were not detected in any of the samples. Extensive detection of PFAAs in leachate samples at levels substantially above those found in background surface waters is an evidence that PFAAcontaining municipal solid waste has been disposed in these landfills. Regarding raw leachate samples, the total concentration of 11 PFAAs (∑ PFAAs) ranged from 7280 ng L−1 (CZ) to 292,000 ng L−1 (SH), with a mean value of 82,100 ng L − 1 . The relative contribution of each individual PFAA to the overall PFAA contamination profiles in raw leachates is provided in Fig. 1. PFOA (mean contribution 28.8%), PFBS (26.1%) and PFPrA (15.9%) were the most abundant PFAAs in raw leachates, with concentrations ranging from 281 ng L− 1 (GZ) to 214,000 ng L− 1 (SH), from 1600 ng L− 1 (NJ) to 41,600 ng L− 1 (SH) and from 638 ng L− 1 (CZ) to 10,000 ng L− 1 (SH), respectively. The PFCAs with 9 or more perfluorinated carbons (C9–C14) were detected with less frequency and at lower concentrations than other PFAAs. PFNA and PFDA were only detected in 3 and 2 of 5 samples, respectively, and PFUnA, PFDoA and PFTA were not detected at all. Though PFOS was detected in all samples, its concentration (ranging from 1150 ng L−1 (CZ) to 6020 ng L−1 (SH)) was relatively low compared to PFOA, PFBS and PFPrA. PFOS thus contributed only 2.1–15.8% of ∑ PFAAs. It is tempting to conclude that the relatively lower abundance of PFOS and higher abundance of short-chain PFAAs (especially PFBS) are caused by the phase-out of PFOS-based chemistry by the 3M Company and subsequent production shift to products containing shorter-chain perfluoroalkyl substances, including PFBS-based products. Unexpectedly, GZ site which closed in 2004 didn't show the lowest PFBS concentration, indicating other potentical sources like biodegradation at this site. However, it should be noted that PFOS and the PFCAs with 9 or more carbons sorb considerably more strongly to organic solids than some of the shorter chain PFAAs (Guelfo and Higgins, 2013; Higgins and Luthy, 2006). Sorption of PFAAs is related to the length of the perfluoroalkyl chain (with each CF 2 moiety contributing 0.50–0.60 log units to the measured organic carbon–water distribution coefficients (Higgins and Luthy, 2006)) and thus PFBS will sorb much more weakly to solid organic matter in the landfill than PFOS. One would therefore expect that the abundance of long-chain PFAAs in leachate would be lower than short-chain PFAAs if the amounts added to the landfill in MSW were equivalent. Nevertheless, the high abundance of PFBS in leachate demonstrates that considerable amounts of PFBS-containing waste are now being disposed of in Chinese landfills. The presence of PFOS in the landfills indicates that the phase-out has not reduced PFOS levels in leachate entirely, which may be due to (1) the long use lifetimes of some PFOScontaining products (e.g. in carpets and textiles), (2) the long lifetime of PFOS-containing MSW in the landfill and/or (3) some on-going uses of PFOS-containing products in China (Wang et al., 2014a; Xie et al., 2013). We recommend that landfill leachates be monitored repeatedly over the coming years to assess the effectiveness of phase-outs of e.g. PFOS and long-chain PFCAs and to monitor the appearance of replacement substances such as PFBS.

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300

260

-1

Concentration (10 ng L )

280

25

3

240 20

220 70 60 50

15

PFOS PFHxS PFBS PFDA PFNA PFOA PFHpA PFHxA PFPeA PFBA PFPrA

10

5

40 30

0

20 10 0

CZ

GZ

NJ

SH

SZ

Fig. 1. PFAA concentrations and relative compositions in raw leachate samples.

Concentrations and compositional profiles varied widely among the five landfill sites. ∑ PFAA concentrations in raw leachates were found in the following order: SH (292,000 ng L − 1 ) N SZ (76,800 ng L − 1 ) N GZ (22,100 ng L − 1 ) N NJ (12,200) N CZ (7280 ng L − 1). The concentration differences of individual PFAAs, such as PFHxA, PFHpA, PFOA, PFNA, PFBS and PFHxS, were 1–3 orders of magnitude between high (at sites SH and SZ) and low (at sites GZ, NJ and CZ) values. However, the concentration differences of PFBA, PFPeA and PFOS were less than a factor of 8 among the five sites. PFOA was predominant at SH and SZ, accounting for 73.3% and 37.4% of ∑ PFAAs, while PFBS at CZ and GZ accounted for 33.6% and 41.8% of ∑ PFAAs, and PFPrA at NJ accounted for 28.5% of ∑ PFAAs. These different PFAA contamination profiles between landfills could indicate either differences in the PFAA contamination profile in MSW disposed in the landfills between cities or differences in the cells that were sampled at the different landfills (e.g. duration of operation, leachate properties, etc.). The highest concentration of ∑PFAAs was observed at the SH site, where the concentrations of individual PFAAs were also high. The ∑ PFAA concentration at site SH was 4–40 times higher than those at the other four sites. Landfill SH is located in Shanghai, one of the most industrialized and urbanized regions in China, reflecting a potential pollution source of PFAAs, such as textiles, carpets, and other PFAA-containing consumer products. Considerable variability has been reported in PFAA concentrations and profiles in leachates from North American and European landfills (Benskin et al., 2012; Bossi et al., 2008; Busch et al., 2010; Huset et al., 2011; Kallenborn et al., 2004; Li et al., 2012; Oliaei et al., 2006). The concentrations of PFAAs obtained for sites SZ, GZ, NJ and CZ fell in the high range of previously reported levels for leachates from landfills that did not receive fluorochemical manufacturing or related wastes (3M, 2001; Benskin et al., 2012; Bossi et al., 2008; Huset et al., 2011; Kallenborn et al., 2004; Woldegiorgis et al., 2006), but were lower than those from landfills associated with the disposal of fluorochemical manufacturing wastes from industrial fluorochemical applications (textile, carpet and paper production) (3M, 2001; Oliaei et al., 2006). The concentrations of PFAAs found at site SH were even higher than those from landfills receiving industrial MSW. Generally, PFCAs accounted for the majority of ∑ PFAAs (50.6%–83.7%) in this study, which is consistent with data for selected US, Nordic, German, and Danish leachates (Bossi et al., 2008; Busch et al., 2010; Huset et al., 2011; Kallenborn et al., 2004; Oliaei et al., 2006). A few studies in the

US and Denmark reported greater PFSA concentrations in leachates than PFCAs (3M, 2001; Bossi et al., 2008). In addition, short-chain PFCAs (bC7) were more abundant in our study than long-chain PFCAs, which is consistent with many previous studies. For example, Li et al. (2012) observed C4–C8 PFCAs that accounted for 73% of ∑ PFAAs in leachates collected from landfills and dumpsites across Canada. Busch et al. (2010) also reported profiles dominated by C4–C8 PFCAs in leachates from landfills in Germany. The dominance of short-chain PFCAs (bC7) is probably due to their preferential release and leaching from municipal solid waste, which is consistent with the higher aqueous solubilities (Higgins and Luthy, 2007) and lower organic carbon–water partition coefficients (Guelfo and Higgins, 2013; Higgins et al., 2005) of the short-chain homologues relative to the long-chain homologues. The PFBS concentrations detected in this study were 1–3 orders of magnitude higher in comparison to the reported values in leachates from Sweden (b0.5–112 ng L − 1) (Kallenborn et al., 2004; Woldegiorgis et al., 2006), and equal to those reported in Germany (b0.39–1356 ng L − 1 ) (Busch et al., 2010) and the US (280–2300 ng L − 1 ) (Huset et al., 2011). PFBS was more abundant than PFOS, which is consistent with the German study (Busch et al., 2010) and the US study (Huset et al., 2011). This study thus provides evidence that China appears to be in line with the global use pattern of PFAAs with a transition from the historically used PFOS-containing products to the newer PFBS-containing products. Despite the phase-out of PFOS by the major producer (3M Company), however, this study shows that there are still large amounts of PFOS-containing MSW in Chinese landfills that will continue to be source of PFOS to the subterranean environment. Treated leachates were analyzed at all sampling sites at different points of the treatment process. All PFAA homologues detected in raw leachates were detected in treated samples. The concentrations of individual PFAAs in treated leachates and the variation of ∑ PFAA concentration relative to raw leachate are shown in Fig. 2. ∑ PFAA concentrations ranged from 98.4 ng L − 1 (RO effluent at SZ) to 282,000 ng L − 1 (RO concentrate at SH), with a mean value of 40,000 ng L−1 (for details see Table S2 in the Supplementary material). PFOA (31.9–206,000 ng L− 1) and PFBS (22.4–55,300 ng L− 1) were again the dominant PFAAs, accounting for, on average, 36.8% and 40.8% of ∑ PFAAs, respectively. NF/RO concentrate samples showed the highest ∑PFAA concentrations (7580–282,000 ng L−1), followed by the bioreactor mixture (4570–111,000 ng L− 1) and UF effluent (2130–79,000 ng L−1). PFAAs were mostly below their MQL in effluent

H. Yan et al. / Science of the Total Environment 524–525 (2015) 23–31

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300 PFOS PFHxS PFBS PFDA PFNA PFOA PFHpA PFHxA PFPeA PFBA PFPrA

250

-1

140

3

Concentration (10 ng L )

200

120

20 18 16

100

14 12

80

10 8

60 40

6 4 2 0

20

CZ

GZ

NJ

SH

RO effluent

RO concentrate

Bioreactor

UF effluent

RO effluent

RO concentrate

Bioreactor

UF effluent

NF effluent

NF concentrate

Bioreactor

UF effluent

RO effluent

RO concentrate

Bioreactor

UF effluent

NF effluent

NF concentrate

Bioreactor

UF effluent

0

SZ

(a)

200%

PFAAs

180% 160% 140% 120% 100% 80% 60% 40% 20%

NJ

Raw leachate Bioreactor UF effluent RO effluent RO concentrate

GZ

Raw leachate Bioreactor UF effluent RO effluent RO concentrate

CZ

Raw leachate Bioreactor UF effluent RO effluent RO concentrate -Raw leachate Bioreactor UF effluent NF effluent NF concentrate

0%

Raw leachate Bioreactor UF effluent NF effluent NF concentrate

PFAAs percentage relavite to raw leachate (%)

220%

SH

SZ

(b) Fig. 2. (a) PFAA concentrations and relative compositions in treated leachate samples. (b) Percentages of ∑PFAA concentrations in treated leachate samples to raw leachate samples.

samples after NF/RO treatment, except for low concentrations of PFOA, PFBS and PFOS at all sites, and low concentrations of PFBA, PFNA and PFHxS at one or two sites. Little information exists about PFAAs in treated leachates and detailed information on the treatment system was seldom reported in previous studies. Only in a survey of leachates collected from 22 landfills in Germany, Busch et al. (2010) reported ∑ PFAA concentrations of 3.97–8060 ng L− 1 in treated leachates, with a mean value of 1336 ng L−1. The ∑PFAAs were in the range of 15–129 ng L−1 (average 42 ng L − 1 ) in RO treated leachates (n = 8), 621 and 1257 ng L − 1 (n = 2) in NF treated leachates, 1992 and 4610 ng L− 1 in wet air oxidation treated leachates, and 4023 and 8059 ng L−1 in biologically treated leachates. When compared to the values for individual PFAAs in this previous study, our data show a relatively similar pattern of

occurrence but higher concentration levels especially in NF/RO concentrate samples.

3.2. Mass flows of PFASs during the treatment process Membrane bioreactor (MBR) treatment has been demonstrated to be an effective alternative for leachate treatment. MBR can provide stable performance while accommodating large variations in the composition of influents and other operation conditions (Ahmed and Lan, 2012). Thus all the 5 leachate treatment facilities in this study employ a two-stage process that integrates a MBR unit with a post-treatment reverse osmosis (RO) or nanofiltration (NF) unit. The MBR utilized in this study uses an external configuration,

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whereby the ultrafiltration (UF) membrane module is outside the bioreactor and the sludge is recirculated to the aeration tank. To assess the fate of PFAAs during each step of different leachate treatment processes, the average mass flow of each PFAA was determined by multiplying the PFAA concentration by the average flow rate of each treatment unit. The flow rate was constant through influent, bioreactor and UF unit, while it was calculated based on the concentrate yield for NF/RO concentrate and effluent. The mass flow of sludge was calculated separately since it is internally recycled between the UF module and bioreactor. Approximately 95% of sludge is recycled while the remaining 5% is waste sludge. The average mass flows of each PFAA for each landfill site are shown in Fig. 3 and Fig. S1 in the Supplementary material. After biological treatment, the changes in mass flows between raw leachate and bioreactor were substance and treatment system specific. Mass flows of PFPrA, PFHxA, PFHpA, PFBS and PFOS decreased in the aqueous phase after the biological treatment process at all sampling

sites, whereas those of PFBA, PFDA and PFHxS at SH, PFBA at NJ, PFNA and PFHxS at CZ, PFPeA, PFOA and PFDA at SZ, PFBA, PFPeA PFOA and PFHxS at GZ increased significantly. The significant reduction of some PFAAs during biological treatment cannot be attributed to biodegradation since PFAAs are not susceptible to biodegradation (Higgins et al., 2007). The removal may be possibly due to the sorption onto the activated sludge, which was consistent with the considerable amounts of PFAAs detected in the sludge phase. The mass flows associated with the sum of the aqueous and sludge phase after biotreatment showed net decreases compared to raw leachate for most PFAAs. Losses due to volatilization are considered unlikely due to the negligible vapor pressure of PFAAs when fully dissociated (Vierke et al., 2013). This net mass decrease could be due to water-to-air transfer of PFAAs by bursting bubbles from the aeration tank (Reth et al., 2011). PFAAs are strong surfactants and are known to be enriched at the air/water interface (Psillakis et al., 2009). The PFAAs in the aeration tank, where bubbles are generated in large amounts, can be ejected on spray

Landfill CZ

Landfill GZ

Landfill NJ

Fig. 3. Average mass flows (g d−1) of dominant PFAAs for each landfill site. UF = ultrafiltration, RO = reverse osmosis, NF = nanofiltration, RS = recycled sludge, WS = waste sludge, n.d. = not detected.

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Landfill SH

Landfill SZ

Fig. 3 (continued).

aerosols and transported to the atmosphere. In some cases, the combined mass flows of the aqueous and sludge phases increased compared to raw leachate, indicating a net formation of these compounds. For example, the combined mass flows increased by 24%, 16%, 55%, 258%, 112%, 216%, and 100% for PFBA, PFPeA, PFHpA, PFOA, PFNA, PFDA and PFHxS at GZ, respectively. This increase may be due to the degradation of precursors such as fluorotelomer alcohols (FTOHs), perfluoroalkyl phosphates (PAPS), or fluorotelomer sulfonates (FTSs) in the biological process, as reported in the literature (Schultz et al., 2006; Yu et al., 2009). Similar results were reported by Busch et al. (2010) showing higher PFAA levels in leachates after biological treatment. The UF membrane module is responsible for the separation of the treated water from biosolids or microorganisms (Cicek, 2003). The PFAA effluent mass flows in the bioreactor decreased significantly (by 5%–100%) after the UF unit for all compounds at all sites. This efficient removal of PFAAs after UF could be explained by the removal of suspended particulates and or colloidal material, to which PFAAs are sorbed. NF and RO separate the leachate into a clean permeate and a contaminated residue. Both NF and RO showed efficient removal of PFAAs in effluent. At site GZ, SH and SZ, the mass flows of individual PFAAs decreased by 90%–100% in the RO effluent compared to the UF effluent, while it decreased by 5%–100% in the NF effluent at sites CZ and NJ. Previous studies also reported that PFAAs were effectively rejected after RO treatment (Appleman et al., 2014; Tang et al., 2006; Thompson et al., 2011). The membrane pores of RO are smaller than NF therefore more effective for PFAA removal than NF. Despite RO's effectiveness, RO is more costly than NF treatment (Appleman et al., 2014). Within the error of the measurements, the mass flows associated with the sum of the NF/RO effluent and concentrate were in agreement with or slightly lower than that of the UF effluent. The decrease of

net flows may be due to the adsorption onto membrane materials. After post-treatment, most PFAAs in the final effluent were removed completely, with the exception of few cases of extremely high concentrations in raw leachate. These results indicate that the MBR-NF/RO system is effective for the removal of PFAAs from the aqueous phase. It is noteworthy that the daily outflows of ∑PFAAs were estimated at 0.63–45.2 g d−1 in the NF/RO concentrate. This NF/RO concentrate is usually recirculated back into the landfill or disposed by incineration. It could be a significant source of PFAAs, which poses a high risk of secondary pollution to the environment during disposal. 3.3. PFAA concentration dependence on leachate physical–chemical properties PFAAs are subjected to a long-term leaching process during the lifetime of a landfill site. Meanwhile, PFAA precursors can also migrate or be leached to aqueous media which are then biotransformed to PFAAs (Stock et al., 2004). These processes are affected not only by physical–chemical properties of PFAAs, but also potentially by the leachate physical–chemical properties, which impact the mass transfer of PFAAs and precursors from solid waste to aqueous media and the biotransformation of precursors. Notably, the leachates examined in the present study contain combined contributions from throughout the landfill (i.e. multiple cells at various stages of stabilization), thus it was not possible to examine the effect of the leachate properties on PFAA levels for a specific stabilization phase. However, we did analyze the correlation coefficients between individual PFAAs in raw leachates and a few leachate properties (pH, EC and TOC, see details in Table 1), to identify the factors contributing most to the PFAA levels. The coefficients obtained in this study suggested a non-statistically significant but moderate or weak impact of leachate physical–chemical

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H. Yan et al. / Science of the Total Environment 524–525 (2015) 23–31

parameters on the occurrence of PFAAs. Negative correlations were found between EC (a measure of the concentration of ions in solution) and concentrations of PFPrA (r = − 0.418), PFBA (r = − 0.706), PFHxA (r = − 0.561), PFOA (r = − 0.320), PFNA (r = − 0.730), PFDA (r = − 0.737), PFBS (r = − 0.472), PFHxS (r = − 0.823) and PFOS (r = − 0.483). This is consistent with the decrease in mobility of PFASs with increasing ionic strength (Higgins and Luthy, 2006; You et al., 2010). Moderate associations were observed between concentrations of PFPrA (r = 0.509), PFPeA (r = 0.514), PFHxA (r = 0.387), PFHpA (r = 0.413), PFOA (r = 0.409), PFBS (r = 0.364) and PFOS (r = 0.489) and increasing pH. Benskin et al. (2012) reported significant correlation between increasing pH and PFAA concentrations in a landfill. These results are consistent with several studies (Higgins and Luthy, 2006; Wang and Shih, 2011), demonstrating that PFAA mobility is enhanced with increasing pH. However the pH range of leachate in this study was narrow and several log units above the pKa of the PFAAs. We therefore have no plausible, mechanistic explanation for the effect of pH on PFAA leachate concentrations. It is well-known that hydrophobic organic contaminants are closely associated with organic carbon/matter (Nam et al., 2008). TOC in leachate could potentially enhance the leachability of PFAAs in a landfill. Contrary to our hypothesis, no significant relationships were observed between individual PFAA concentrations and leachate TOC. This is likely partly due to the heterogeneity and complex chemical nature of landfill leachate as well as the relatively low sorption of PFAAs to organic matter (Higgins and Luthy, 2006). In addition to the leachate properties investigated here, there are many other parameters that could influence leachability, including age and composition of waste, temperature of the landfill, degree of waste compaction, stage of waste decomposition, waste filling procedures, physical state of waste, moisture content, and rate of water movement (Edil, 2003; Mor et al., 2006; Qasim and Chiang, 1994). 3.4. PFAA leakage estimation into the subterranean environment As demonstrated in Section 3.2, MBR-NF/RO system is effective for the removal of PFAAs from raw leachate. Most PFAAs present in raw leachate were removed from the final effluent by NF or RO to concentrations below detection limits. Consequently release of the final effluent could not be considered as one of the principal sources of PFAAs to surface water. However, the most serious environmental impact of waste disposal landfill is the contamination of local groundwater (Edil, 2003; Mor et al., 2006), thus there is a high potential threat of PFAA migration into the subterranean environment through landfill leachate. In China, most landfills are not adequately lined to prevent leachate migration. Even for the standard landfills with a composite liner, there is a risk of leakage due to inevitable defects introduced during installation. For landfills which are not adequately lined, leachate containing PFAAs generated from municipal solid waste can permeate into the subterranean environment, thus contaminating groundwater. The Chinese national leakage rate of PFAAs (kg3 year−1) for inadequately lined landfills was estimated by multiplying the PFAA concentrations in leachate (converted to kg m−3) by the average amount of leachate generated (m3 year−1), using the following equation: L ¼ Ci  k  Q

ð1Þ

where L is the Chinese national leakage rate of PFAA (kg3 year−1), Ci the mean concentration of PFAA in raw leachate in this study (converted to kg m− 3), Q the annual volume of leachate generated in China (m3 year−1), and k the percentage of landfills not being lined. The annual volume of leachate generated in China was estimated to be 4735 × 104 m3 in 2010 (Ministry of Housing and Urban–Rural Development of the People's Republic of China). Based on the mean

concentrations of PFAAs in raw leachate in this study, the national leakage of PFAAs was estimated, assuming that the percentage of landfills not being lined is 80% (before the new standard for pollution control on landfill sites issued in 2008, no technical requirements for liner systems were proposed in China and most landfills were not adequately lined), and that the leakage through lined landfills is negligible. Uncertainty ranges in the estimates of leakage rates of PFAAs were generated by using the uncertainty ranges of the PFAA concentrations in raw leachates. Approximately 3110 ± 4000 kg year−1 of ∑ PFAAs was estimated to be discharged into the subterranean environment surrounding landfills in China, which is 2 times higher than the national discharge load of ∑ PFAAs (1630 kg year−1) from wastewater treatment plants in Korea (Kim et al., 2012). For individual PFAAs, the national leakage rates (kg year−1) are 222 ± 131 for PFPrA, 133 ± 111 for PFBA, 95 ± 77 for PFPeA, 54 ± 61 for PFHxA, 49 ± 84 for PFHpA, 1860 ± 3085 for PFOA, 6 ± 6 for PFNA, 577 ± 555 for PFBS, 5 ± 7 for PFHxS and 103 ± 67 for PFOS, respectively. These leaching rates of PFAAs are considered high and would make an additional contribution to Chinese source emission inventories (Lim et al., 2011; Wang et al., 2014b). Therefore, the risk of groundwater pollution caused by PFAA leakage and subsequent threat to human health from exposure to PFAA contaminated matrices should not be ignored and requires further investigation. 4. Conclusions This study is important for understanding the PFAA contamination level in landfill leachates across China, along with the mass flow and fate of PFAAs during leachate treatment processes. It is the first investigation on the potential risks of PFAA contamination to subterranean environment through landfill leachate. The results showed that municipal landfills can be a significant source of PFAAs to the environment. PFOS continues to be detected in landfill leachates indicating that the 2000–2002 phase-out has not been fully effective in removing PFOS from waste sites in China. To assess the fate of PFAAs during each step of different leachate treatment processes, the mass flow of each PFAA was calculated. The results showed that the fate of individual PFAAs was substance and treatment system specific. The MBR-NF/RO system is generally effective for the removal of PFAAs from the aqueous phase. The national leakage of ∑PFAAs was estimated to be 3110 kg year−1 through landfill leachate, indicating that landfill leachate is a significant source of PFAAs to groundwater in China which could threaten the sustainable use of groundwater as a drinking water source. Significant amounts of the replacement substance PFBS are present in Chinese landfills and it is one of the most abundant PFAAs in leachate, indicating a new contamination threat to groundwater. These results suggest that waste disposal sites need to be further monitored to assess the risks of PFAA pollution on the global environment. Acknowledgment We acknowledge funding for this research from the National Science Foundation of China (Project Nos. 21177094, 41271465), and the Fundamental Research Funds for the Central Universities (Project No. 0400219271). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2015.03.111. References 3M, 2001. Multi-City Study: Water, Sludge, Sediment, POTW Effluent and Landfill Leachate Samples. 3M Laboratories. Ahmed, F.N., Lan, C.Q., 2012. Treatment of landfill leachate using membrane bioreactors: a review. Desalination 287, 41–54.

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Perfluoroalkyl acids in municipal landfill leachates from China: Occurrence, fate during leachate treatment and potential impact on groundwater.

Raw and treated landfill leachate samples were collected from 5 municipal landfill sites in China to measure the concentrations and contamination prof...
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