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. Author manuscript; available in PMC: 2023 Oct 17.
Published in final edited form as: Environ Sci Technol Lett. 2021;8:66–72. doi: 10.1021/acs.estlett.0c00819

From Waste Collection Vehicles to Landfills: Indication of Per- and Polyfluoroalkyl Substance (PFAS) Transformation

Yalan Liu 1, Nicole M Robey 2, John A Bowden 3, Thabet M Tolaymat 4, Bianca F da Silva 5, Helena M Solo-Gabriele 6, Timothy G Townsend 7
PMCID: PMC10581401  NIHMSID: NIHMS1919593  PMID: 37850075

Abstract

Municipal solid waste contain diverse and significant amounts of per- and polyfluoroalkyl substances (PFAS), and these compounds may transform throughout the “landfilling” process from transport through landfill degradation. Fresh vehicle leachates, from commercial and residential waste collection vehicles at a transfer station, were measured for 51 PFAS. Results were compared to PFAS levels obtained from aged landfill leachate at the disposal facility. The landfill leachate was dominated by perfluoroalkyl acids (PFAAs, including perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkyl sulfonic acids (PFSAs); 86% of the total PFAS, by median mass concentration), while the majority of PFAS present in commercial and residential waste vehicle leachate were PFAA-precursors (70% and 56% of the total PFAS, by median mass concentration, respectively), suggesting precursor transformation to PFAAs during the course of landfill disposal. In addition, several PFAS, which are not routinely monitored—perfluoropropane sulfonic acid (PFPrS), 8-chloro-perfluoro-1-octane sulfonic acid (8Cl-PFOS), chlorinated polyfluoroether sulfonic acids (6:2, 8:2 Cl-PFESAs), sodium dodecafluoro-3H-4,8-dioxanonanoate (NaDONA), and perfluoro-4-ethylcyclohexanesulfonate (PFECHS)—were detected. Potential degradation pathways were proposed based on published studies: transformation of polyfluoroalkyl phosphate diester (diPAPs) and fluorotelomer sulfonic acids (FTS) to form PFCAs via formation of intermediate products such as fluorotelomer carboxylic acids (FTCAs).

Graphical Abstract

graphic file with name nihms-1919593-f0004.jpg

1. INTRODUCTION

For several decades, manufacturers have added per- and polyfluoroalkyl substances (PFAS) to commonplace consumer products and packaging.1–8 The introduction of these chemicals to the environment and subsequent animal and human exposure is pervasive.9–13 While some PFAS originating from products are released into wastewater from bathing, cleaning, and laundering,14–16 the solid waste stream may contain a larger PFAS mass. Waste collection vehicles transport municipal solid waste (MSW—the garbage from businesses, schools, government buildings, and similar institutions (commercial waste), combined with the discarded materials from homes, apartments, and other living units (residential waste)—and deliver it to an appropriate management facility. In the U.S., the majority of commercial and residential waste is disposed of in a MSW landfill.17,18

PFAS are routinely detected in MSW landfill leachate,3,19–28 leading to growing concerns regarding the contribution of landfills to environmental PFAS loading and interest in leachate treatment prior to discharge.29–31 Perfluoroalkyl acids (PFAAs), consisting of perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkyl sulfonic acids (PFSAs), are the most studied PFAS in landfill leachate, especially perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS), which have been reported at concentrations up to 82,000 ng L−1 in landfills,26,32 orders of magnitude higher than the 70 ng L−1 U.S. Environmental Protection Agency drinking water health advisory. Commercial and residential waste represent one input of PFAS to MSW landfills; other potential sources include industrial waste, media from soil and groundwater remediation, wastewater treatment plant sewage sludge, and chemicals used to suppress landfill fires.14,21,33–36 The relative contribution of different PFAS sources to landfills is not well understood.

Here, the role of commercial and residential waste as a possible PFAS source is described by characterizing PFAS in liquids from waste collection vehicles (designated as vehicle leachate) and comparing it to liquids collected from the bottom leachate collection systems at the landfill that receives the waste (designated as landfill leachate). Well-maintained waste collection vehicles contain liquids drained from the waste after collection and during transport. This vehicle leachate, defined and regulated similar to landfill leachate, is released where the waste is unloaded (usually a landfill). In this study, vehicle leachate from commercial and residential sources was collected during the waste unloading process at a MSW transfer station (a facility that combines waste from multiple collection vehicles into larger transfer trailers for transport to the landfill).

By comparing PFAS in vehicle leachate to PFAS in landfill leachate from the disposal site, we provide unique data regarding the commercial and residential waste contribution to the overall PFAS levels in landfills. Furthermore, we highlight the importance of precursor transformation to PFAAs (also referred to as terminal PFAS) in the chemically and biologically complex landfill environment.17,24,37,38 While PFAAs are considered nondegradable under most environmental conditions,39–41 PFAA-precursors (e.g., fluorotelomers), to varying degrees, have a potential to transform into PFAAs in the landfill.19,24,27,42 Though previous studies have shown biodegradation pathways of PFAA-precursors in environmental media,43–51 only a few studies examined the conditions likely in landfills, either using lab-scale simulated landfill reactors37,52 or lab-scale leachate–sediment microcosms,38 and little is known about the varieties of PFAS entering landfills in as-disposed waste. By studying waste collection vehicle leachate, we can characterize the PFAS contents of this prelandfill waste and compare it to leachate after it has undergone landfill transformations. Leachate obtained from an individual waste collection vehicle may not be representative of the entire landfill waste mass; however, we propose that individual waste vehicle analysis provides crucial preliminary insight into PFAS loading from MSW and the fate of PFAS during the waste decomposition process.

2. MATERIALS AND METHODS

2.1. Sample Collection.

Vehicle leachate samples were collected as leachate was drained during the normal unloading process from nine commercial and nine residential waste collection vehicles at a transfer station. The transfer station serves a county in north central Florida (population of 270,000) where industry is not a major component of the commercial sources and where there is no known history of PFAS contamination. Three vehicles of each type were sampled per event, over three sampling events during the winter season. The waste stream is similar in makeup to typical Florida MSW based on waste composition studies from this region and others in Florida; these studies also find that season does not cause major waste compositional differences.53,54 Vehicle leachate originates from moisture inherently present in the waste, as well as rainfall contact with waste prior to and during waste collection; in this regard, it is similar to landfill leachate but with lower contact time with the waste and not subjected to the elevated temperatures and anerobic environment typically present in a landfill. During a separate sampling event, nine leachate samples were collected from nine lined landfill cells at one site (the oldest unit was opened in 1992, and all nine cells continue to operate). Approximately 70% of the disposed waste at this landfill originates from the studied transfer station, with the rest resulting from commercial and residential sources from more rural counties; the landfill does not accept hazardous waste and is typical of most MSW landfills in Florida. A field blank was collected for each sampling event, and all samples were stored at −20 °C until analysis; see Section 1 in the Supporting Information (SI) for detailed sample description.

2.2. Extraction and Analysis.

The sample extraction method was adapted from Robey et al.55 Briefly, leachate was centrifuged and a 50 mL subsample was adjusted to pH 4–5 using glacial acetic acid and spiked with mass-labeled internal standards (Table S2 in the SI). Phenomenex cartridges (Strata-X-AW 100 μm polymeric weak anion exchange, 500 mg/6 mL) were conditioned with 0.3% ammonium hydroxide in methanol (4 mL), methanol (3 mL), and an ammonium acetate/acetic acid buffer solution (4 mL; pH = 4). After sample loading (at a rate of 1 drop per second), cartridges were washed with the buffer solution (4 mL) and eluted with methanol (4 mL) followed by 0.3% ammonium hydroxide in methanol (4 mL). Extracts were reduced using a Biotage TurboVap II to 4 mL.

Extracts were analyzed for 51 PFAS using a Thermo Scientific Vanquish ultrahigh-pressure liquid chromatograph coupled to a TSQ Quantis triple quadrupole mass spectrometer (UHPLC-MS/MS):13 PFCAs (C4−C14, C16, C18), nine PFSAs (C3−C10, C12), 24 PFAA-precursors including fluorotelomer sulfonic acids (4:2, 6:2, 8:2, and 10:2 FTS), saturated fluorotelomer carboxylic acids (6:2, 8:2, and 10:2 FTCA), unsaturated fluorotelomer carboxylic acids (8:2 and 10:2 FTUCA), perfluoroalkane sulfonamido substances (FBSA, FHxSA, N-AP-FHxSA, FOSA; N-MeFOSA, N-EtFOSA, FOSAA, N-MeFOSAA, N-EtFOSAA), perfluoroalkyl phosphinic acids (6:6 and 6:8 PFPi), and polyfluoroalkyl phosphate diester (6:2, 8:2, and 6:2/8:2 diPAP, diSAmPAP). In addition, five PFAS (8-chloro-perfluoro-1-octane sulfonic acid (8Cl-PFOS), chlorinated polyfluoroether sulfonic acids (6:2, 8:2 Cl-PFESAs), sodium dodecafluoro-3H-4,8-dioxanonanoate (NaDONA), and perfluoro-4-ethylcyclohexanesulfonate (PFECHS)) not fitting into traditional PFAS classes (“Other PFAS” hereafter) were analyzed. This list includes many PFAS not routinely analyzed in landfill leachate (e.g., short-chain perfluoroalkane sulfonamido substances).

For quantification, a 14-level calibration (from 10 to 100,000 ng L−1) was developed for 51 PFAS through serial gravimetrically-derived dilutions of primary stock solutions. A mixture of 24 mass-labeled PFAS internal standards (Table S2) at concentrations of ~800 ng L−1 was also added to each calibration level. When a labeled standard was not available for a compound or had a poor response, a labeled standard similar by structure or retention time (or that provided the most stable response within the batch) was selected; similar practices have been reported in other studies.20,27,55–57 Those compounds using other isotopically labeled standards are denoted in Table S2 in the SI.

QA/QC samples include field, method, and solvent blanks; laboratory control samples; and sample replicates (see SI Section 2 for detailed QA/QC information). The instrumental method and mass spectrometric parameters can be found in SI Section 3.

2.3. Data Analysis.

If a compound was observed in blanks, the average peak area of blanks was subtracted from all associated samples. As data are not normally distributed, the results reported here and in the SI are all based on median concentrations.

Frequent nondetects in environmental data (including those from this study) can lead to inaccurate statistical analysis; consequently, there has been discussion in the scientific community regarding how to treat these data. One conclusion has been that traditional statistical analyses (e.g., replacing nondetects with a fraction of detection limits or simply ignoring them) are not appropriate. For this study, a statistical package called NADA (Nondetects and Data Analysis) in R software was used to analyze left-censored data,58–60 which is the data type present in this study. In NADA, a nonparametric Kaplan–Meier method and Peto–Peto–Prentice rank testing were used for computing descriptive statistics and to determine whether differences among different sample types were significant (p-value < 0.05), respectively. These methods allow for better and more accurate scientific interpretations for censored data.58–60

3. RESULTS AND DISCUSSION

Concentrations of the 51 PFAS, as well as the QC results, are presented in greater detail in the SI Section 4. Overall, the method performs well for PFAS quantification, with negligible contamination in blanks, high extraction efficiency (average of about 80%), and desirable reproducibility (average of <15% relative standard deviation).

3.1. PFAS Concentrations in Landfill and Vehicle Leachates.

For landfill leachates, 38 of the 51 PFAS were detected in at least one sample above their respective LOD. The detection frequencies (presence above LOD) varied among PFAS, from 11% (e.g., PFDS) to 100% of the samples (e.g., PFHxA) (Table S6). In vehicle leachates, 36 and 48 PFAS were detected in at least one commercial and residential vehicle leachate sample, respectively (Table S6). The detection frequencies ranged from 11% of samples (e.g., PFHxDA) to 100% of samples (e.g., PFHxA).

Figure 1 displays relative PFAS contributions of each class (% of ∑51PFAS) in landfill and vehicle leachates with the total concentrations of PFAS classes listed. The ∑51PFAS in landfill, commercial, and residential vehicle leachate were 9700, 3300, and 3400 ng L−1, respectively. Landfill leachate was dominated by PFAAs (57% ∑13PFCAs and 29% ∑9PFSAs by mass concentration), while ∑24PFAA-precursors only contributed to 14% of the ∑51PFAS. Comparatively, for commercial and residential waste vehicle leachates, ∑24PFAA-precursors were the dominant contributors to the ∑51PFAS, representing 70% and 56%, respectively. The detailed concentrations for each compound are in Table S6.

Figure 1.

Figure 1.

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Relative% contribution of each PFAS class by median mass concentration. PFAS classes include short-chain PFCAs (C4–C7), long-chain PFCAs (C8–C14, C16, C18), short-chain PFSAs (C3–C5), long-chain PFSAs (C6–C10, C12), PFAA-precursors, and “Other PFAS”. The total median concentration of each PFAS class (as displayed) was divided by the total median concentration of 51 PFAS.

3.1.1. PFCAs and PFSAs.

Of the 13 PFCAs, the C5–C12 PFCAs were detected in 100% of the landfill leachate samples with PFHxA detected at the highest concentration, 2400 ng L−1; the C15 and C16 PFCAs were not detected in any landfill leachate samples. All PFCAs were detected in both commercial and residential leachate with C5–C8, C10, and C12 and C4–C9, C11, and C12 being detected in 100% of the commercial and residential waste vehicle leachates, respectively.

Of the nine PFSA, C3, C4, C6, and C8 PFSA were detected in 100% of the landfill leachate samples with the highest detected concentration of 1900 ng L−1 for PFHxS. In contrast, C3, C4, and C6 PFSA and C3–C6 and C8 PFSA were detected in all commercial and residential vehicle leachates, respectively, with the highest detected concentration of 95 ng L−1 for PFPrS in commercial vehicle leachate and 150 ng L−1 for PFHxS in residential vehicle leachate. It is noteworthy that PFPrS has received little attention, only Robey et al.55 reported this compound in landfill leachate, at 11 ng L−1. The C7 and C9 PFSA were detected in one residential vehicle leachate sample and no landfill or commercial vehicle leachate samples; the C12 PFSA were not detected in any landfill or vehicle leachate. Respective concentrations of PFOA and PFOS were 1100 and 100 ng L−1 in landfill leachate, 46 and 57 ng L−1 in commercial vehicle leachate, and 29 and 28 ng L−1 in residential vehicle leachate.

Short-chain PFCAs (C4–C7) have historically accounted for a majority of PFAS in landfill leachate.3,19,42,61 Figure 2 presents the percent contribution of each PFAA as a function of fluorinated carbon chain length. Consistent with the literature,19,20,22 PFAAs distribution in all leachate types is right-skewed, with more short-chain species overall (especially C4–C7), with vehicle leachate slightly more so (predominantly C3–C6). This is likely due to higher water solubilities and lower organic carbon−water partition coefficients in short-chain PFAAs as well as the phase out of some long-chain PFAAs and their precursors by manufacturers.62,63

Figure 2.

Figure 2.

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PFAAs percentage of ∑51PFAS (median mass concentration) relative to the fluorinated carbon chain length in landfill and vehicle leachates.

3.1.2. PFAA-Precursors.

In landfill leachate, 20 of the 24 PFAA-precursors were detected; seven species (4:2, 6:2, 10:2 FTS, 6:2 FTCA, 8:2 FTUCA, N-MeFOSAA, and 6:2 diPAP) had 100% detection frequency, with 6:2 FTCA detected at the highest median concentration of 920 ng L−1. For vehicle leachate, 17 and 22 species were detected in at least one of the commercial and residential leachate samples, respectively, while four species (4:2 FTS, 6:2 FTS, 6:2 FTCA, 6:2 diPAP) were consistently detected in all commercial and residential vehicle leachate samples. The predominant precursors in vehicle leachate were 6:2 FTS and 6:2 diPAP (respective concentrations of 1200 and 880 ng L−1 in commercial, 810 and 920 ng L−1 in residential vehicle leachate).

3.1.3. “Other PFAS”.

The global phase out of long-chain PFAAs and their precursors has prompted a shift toward the manufacture of replacement PFAS,64 many of which do not fit into traditional categorization and are referred to as “Other PFAS”. Five “Other PFAS” (8Cl-PFOS, 6:2 Cl-PFESAs, 8:2 Cl-PFESAs, NaDONA, and PFECHS) were analyzed in this study. Data are sparse regarding these PFAS in environmental samples. MacInnis et al.65 measured the five “Other PFAS” in sediments from Arctic lakes in Canada, and only low concentrations of PFECHS were detected. PFECHS has also been detected in fish, sediment, surface water, and drinking water.66–68 To our knowledge, except Robey et al.,55 who measured low concentrations of PFECHS in landfill leachate, no “Other PFAS” have been previously reported in landfill leachate. In this study, PFECHS was detected in seven out of nine landfill leachate samples, and all five “Other PFAS” were detected in a single residential vehicle leachate sample.

3.2. PFAS Comparison among Landfill, Commercial Waste Vehicle, and Residential Waste Vehicle Leachates.

Total median concentrations of ∑51PFAS in commercial and residential vehicle leachate were on average three times lower than in landfill leachate, while conductivity (Table S1), a general measurement of ionic strength, was essentially the same in all three leachate types. This suggests that PFAS transformation and leaching from MSW may not be complete in waste prior to landfilling and that the landfill environment (i.e., decomposition, elevated temperatures, low oxygen, and prolonged contact time) is conducive to PFAS transformation and leaching.

To further explore PFAS concentrations and diversity differences in landfill and vehicle leachates, statistical analysis was conducted among every compound in each leachate type (Table S7, with the p-value presented). Almost all PFAS showed in higher concentrations in landfill leachate than in vehicle leachate, with a few exceptions: three PFAA-precursors, 6:2 FTS, 6:2 and 8:2 diPAP. A review of biodegradation pathways of these precursors37,38,43–52,69–72 identifies potential intermediate and terminal breakdown products (Figure 3), and the concentrations of these byproducts in the landfill leachate samples suggests that these precursors are indeed undergoing biodegradation in the landfill. For example, while 6:2 diPAP concentrations are 31 and 32 times higher in the commercial and residential vehicle leachates as compared to landfill leachate, respectively, concentrations of its intermediate degradation product, 6:2 FTCA, are 10 and 11 times lower (commercial and residential vehicle leachates, respectively), and the terminal PFCAs breakdown products (C4–C7 PFCA) are on average 10 and 9 times lower in commercial and residential vehicle leachates, respectively. Fluorotelomer 6:2 FTS, with concentrations 5 and 4 times higher in the commercial and residential vehicle leachate, respectively, follows similar degradation pathways as presented in Figure 3.46,70

Figure 3.

Figure 3.

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Suspected biodegradation pathways of diPAPs and FTS to form PFCAs, adapted from the literature.38,43–47,69–72

3.3. Study Implications.

In this study, a diverse suite of PFAS were detected in the leachate of commercial and residential waste collection vehicles prior to waste disposal in a landfill. The landfill where the landfill leachate samples were collected has not historically accepted special wastes which are likely to have high levels of PFAS (e.g., remediation waste,33,35 industrial waste14,21). PFAS in this leachate have thus been predominantly derived from MSW (i.e., household, consumer waste). These results indicate that PFAS historically measured in landfill leachates are not necessarily derived from special waste, and the basic waste stream components from homes and businesses are significant contributors to PFAS loading at municipal landfill sites. While a direct comparison of concentrations between landfill and vehicle leachates has limitations because of differences in exposure conditions (i.e., liquid to solid ratio, contact time) and waste stream variability among collection vehicles, the relative diversity of the compounds measured in these two leachate types is noteworthy and warrants further attention.

To summarize, the majority of PFAS present in vehicle leachate were precursors, whereas landfill leachate was predominantly terminal PFAS. PFECHS was detected in landfill leachate, while five “Other PFAS” were detected in vehicle leachate; this is the first time these PFAS (except PFECHS) have been detected in leachate samples. While most PFAS showed higher concentrations in landfill leachate, a subset (6:2 FTS, 6:2 and 8:2 diPAP) were measured at higher concentrations in vehicle leachate, suggesting a potential biodegradation pathway in landfills: transformation of diPAPs and FTS to form PFCAs through the formation of intermediate degradation products (e.g., FTCAs). However, further studies are needed to add to our understanding of transformation pathways in landfills. The landfill transformation of precursor PFAS (including unknown compounds) to terminal species, as well as the presence of uncommon PFAS, especially regulated compounds identified with greater risk, warrants further research and scrutiny given society’s heavy reliance on landfills as a disposal sink for household and commercial waste.

Supplementary Material

Supplementary Material

ACKNOWLEDGMENTS

The researchers have received support from the Hinkley Center for Solid and Hazardous Waste Management and from the U.S. Environmental Protection Agency under the Science To Achieve Results (STAR) grant program (EPA-G2018-STAR-B1; Grant 83962001-0) to investigate the occurrence, source, and fate of PFAS in landfills.

Footnotes

Supporting Information

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.estlett.0c00819.

Sample description, acronyms and structures of analyzed PFAS, additional information about QA/QC procedure and instrumental analysis, detailed PFAS and QA/QC results (PDF)

Complete contact information is available at: https://pubs.acs.org/10.1021/acs.estlett.0c00819

The authors declare no competing financial interest.

Contributor Information

Yalan Liu, Department of Environmental Engineering Sciences, College of Engineering, University of Florida, Gainesville, Florida 32611, United States.

Nicole M. Robey, Department of Environmental Engineering Sciences, College of Engineering, University of Florida, Gainesville, Florida 32611, United States

John A. Bowden, Department of Environmental Engineering Sciences, College of Engineering and Center for Environmental and Human Toxicology & Department of Physiological Sciences, College of Veterinary Medicine, University of Florida, Gainesville, Florida 32611, United States

Thabet M. Tolaymat, National Risk Management Laboratory, Office of Research and Development, U.S. Environmental Protection Agency, Cincinnati, Ohio 45268, United States

Bianca F. da Silva, Center for Environmental and Human Toxicology & Department of Physiological Sciences, College of Veterinary Medicine, University of Florida, Gainesville, Florida 32611, United States

Helena M. Solo-Gabriele, Department of Civil, Architectural, and Environmental Engineering, College of Engineering, University of Miami, Coral Gables, Florida 33146, United States

Timothy G. Townsend, Department of Environmental, Engineering Sciences, College of Engineering, University of Florida, Gainesville, Florida 32611, United States

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