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. 2024 May 16;58(21):9303–9313. doi: 10.1021/acs.est.3c10098

The Ins and Outs of Per- and Polyfluoroalkyl Substances in the Great Lakes: The Role of Atmospheric Deposition

Chunjie Xia †, Staci L Capozzi †, Kevin A Romanak †, Daniel C Lehman †, Alice Dove ‡, Violeta Richardson ‡, Tracie Greenberg ‡, Daryl McGoldrick ‡, Marta Venier †,*
PMCID: PMC11137863  PMID: 38752648

Abstract

graphic file with name es3c10098_0005.jpg

As part of the Integrated Atmospheric Deposition Network, precipitation (n = 207) and air (n = 60) from five sites and water samples (n = 87) from all five Great Lakes were collected in 2021–2023 and analyzed for 41 per- and polyfluoroalkyl substances (PFAS). These measurements were combined with other available data to estimate the mass budget for four representative compounds, PFBA, PFBS, PFOS, and PFOA for the basin. The median Σ41PFAS concentrations in precipitation across the five sites ranged between 2.4 and 4.5 ng/L. The median Σ41PFAS concentration in lake water was highest in Lake Ontario (11 ng/L) and lowest in Lake Superior (1.3 ng/L). The median Σ41PFAS concentration in air samples was highest in Cleveland at 410 pg/m3 and lowest at Sleeping Bear Dunes at 146 pg/m3. The net mass transfer flows were generally negative for Lakes Superior, Michigan, and Huron and positive for Lakes Erie and Ontario, indicating that the three most northern lakes are accumulating PFAS and the other two are eliminating PFAS. Atmospheric deposition is an important source of PFAS, particularly for Lake Superior.

Keywords: PFAS, Great Lakes, mass budget, atmospheric deposition, tributaries

Short abstract

Atmospheric deposition contribution to PFAS loadings varies across the Lakes. PFAS mass budget in the Great Lakes will be helpful in understanding the fate and transport of PFAS for public health and regulatory purposes.

Introduction

Per- and polyfluoroalkyl substances (PFAS) are environmental pollutants that have been broadly used as surfactants in consumer products (e.g., textiles, furniture, and paper) and as components of aqueous film-fire-fighting foams.1−4 PFAS can be released into the environment via multiple routes, including atmospheric emissions, sewage water, and landfill leachate discharges.5,6

PFAS have become ubiquitous in the environment because of long-range aquatic and atmospheric transport.7,8 The ionic PFAS tend to enter the aquatic environment and can be transported considerable distances through groundwater.9 The neutral PFAS, such as the fluorotelomer alcohols (FTOHs), fluorotelomer acrylates, fluoroalkyl sulfonamides (FASAs), and fluoroalkyl sulfonamidoethanols (FASEs), are volatile, and they can undergo long-distance atmospheric transport.10,11 During the atmospheric transport process, these volatile compounds are subject to photochemical reactions to form persistent ionic PFAS [for example, perfluorooctanesulfonic acid (PFOS), and perfluorooctanoic acid (PFOA)], a process that can account for the presence of these compounds in remote regions including the Arctic, the Antarctic, and the Tibetan Plateau.11−15

As part of the Integrated Atmospheric Deposition Network (IADN), persistent organic pollutants (POPs) have been monitored at representative sites in the Great Lakes basin since 199016 and recently PFAS have been added to the list of target chemicals. Like other legacy compounds, such as polychlorinated biphenyls (PCBs) and organochlorine pesticides, atmospheric deposition could be a significant environmental pathway, particularly for the Great Lakes. PFAS are washed out from the atmosphere by wet and dry deposition as well as by the gas absorption process, of which wet deposition (precipitation) is generally considered the dominant mechanism.17−19 PFAS concentrations in precipitation can be affected by several factors, including proximity to sources, physico-chemical properties of individual PFAS compounds (for example, the lengths of the fluorocarbon chain, octanol–air partitioning coefficients, and acid dissociation constants), frequency and amount of wet deposition, and other weather-related conditions.20−22

To date, the mass transfer flow of PFAS via atmospheric deposition to the Great Lakes was not well addressed despite the fact that 10% of the United States (U.S.) population and 30% of the Canadian population live in the Great Lakes basin.23 In the last two decades, only a few studies have investigated PFAS levels in precipitation in the Great Lakes.24−29 Four studies had focused on the annual mass transfer flow of PFAS to the Great Lakes, but they were limited in scope because they did not include either wet deposition (precipitation),24 gas absorption,27,30 dry deposition,27 or tributaries, which are important conduits of PFAS into the Great Lakes.31,32

Thus, the goals of this study were as follows: (a) to provide updated PFAS concentrations in precipitation and air in the Great Lake basin as well as in lake water; (b) to evaluate the importance of atmospheric deposition in the overall mass transfer flow of PFAS to the Great Lakes; (c) to calculate the mass budget of PFAS in the Great Lakes; and (d) to compare PFAS atmospheric deposition to those of legacy pollutants. Ultimately, this study contributes to national and international efforts to control PFAS (e.g., the U.S. EPA Strategic Road Map,33 the Stockholm Convention and the Canadian risk management plan)34,35 by providing insights into the fate and transport of PFAS through atmospheric deposition, which can inform future actions and policies aimed at mitigating PFAS contamination in the Great Lakes and abroad.

Materials and Methods

Sample Collection

Precipitation samples were collected every 2 weeks (Text S1) at five IADN sampling stations36 from 2021 and 2022. The IADN sites included two urban sites in Chicago, IL, and Cleveland, OH; one rural site at Sturgeon Point, NY; and two remote sites, one at Eagle Harbor, MI, and one at Sleeping Bear Dunes, MI (Figure S1). Samples were collected using an atmospheric precipitation sampler (ADS-120, NCON, GA, USA), equipped with a PFAS-free high-density polyethylene bucket, and a sensor that automatically opens the cover during precipitation events. The sampling bucket was thoroughly rinsed three times with PFAS-free water and methanol before and after sampling (Text S2). The total precipitation volume was recorded by using a graduated cylinder. A precipitation sample of 950 mL was transferred to 1 L polypropylene (PP) bottles, and the excess volume, if any, was discarded. A total of 50 mL of LC/MS grade methanol was then added for preservation,36 after which the samples were shipped to the laboratory at Indiana University for PFAS analysis. Field duplicates were obtained annually (n = 3 per site on average) when the precipitation volume exceeded 2 L. Travel blanks were collected annually (n = 3 per site on average) by using 50 mL of LC/MS grade methanol in the same bottles used for the actual samples to assess the extent of potential background contamination from sample handling and transport.

Lake water samples were collected by the US EPA’s Great Lakes National Program Office in May 2022 aboard the R/V Blue Heron from Lakes Superior (n = 15) and Michigan (n = 11) and by Environment and Climate Change Canada (ECCC) between April and June 2022 aboard the CCGS Limnos from Lakes Huron (n = 21), Erie (n = 3), and Ontario (n = 13). Due to the low sample size in 2022, ECCC collected additional samples from Ontario (n = 12) and Erie (n = 12) again in April and May 2023, respectively. Spring sampling ensures a high degree of vertical mixing in the water column. The samples were collected at a 1 m depth below the water surface.37 In each campaign, field duplicates (n = 12) and field blanks (n = 6) were also collected. A total of 950 mL of lake water was collected into precleaned PP bottles using a sampling pole that extended about 2 m away from the vessel. After collection, 50 mL of LC/MS grade methanol was added for preservation. The samples were shipped in coolers on ice. An inventory of lake water samples and total PFAS concentrations is provided in Table S1.

Air samples were collected using a passive air sampler fitted with a sorbent-impregnated polyurethane foam disk (SIP disk; Text S3). The SIP disks were uniformly impregnated with ∼0.5 g of XAD-4 resin as outlined in Ahrens et al.38 and Shoeib et al.39 Briefly, precleaned polyurethane foam disks (14 cm diameter × 1.35 cm thick, Tisch Environmental) were dipped into a XAD-4 resin (<0.75 μm)/hexane slurry (6.4 g L–1) three times to obtain a uniform coating and then dried in a fume hood overnight. The SIP disks were mounted in stainless steel outdoor passive samplers and shipped to the sites where they were set up on fencing or antenna towers, depending on the infrastructure available at each location. Air samples were collected for 30 days at each of the five U.S. IADN sites during 2021 for a total of about 12 samples per site.

Sampling volume was estimated using the Harner template40 using the average temperature for the sampling period and a sampling rate of 4 m3/day, which has been shown to be an acceptable estimate for PFAS as well as other contaminants.38,41

Sample Preparation

Precipitation and lake water samples were processed following the same protocol. The unfiltered sample was first adjusted to a pH ≈ 4.0 using 2% formic acid (v/v) or 1% ammonia hydroxide (v/v) solution. Next, 500 mL of the sample was spiked with surrogate standards (SS) and passed through an Oasis weak anion exchange cartridge (6 cc, 150 mg, Waters, MA, USA). Cartridges were preconditioned as follows: 10 mL of Optima LC/MS grade water, followed by 4 mL of 1% ammonia in methanol, then 4 mL of methanol, and finally 2 × 4 mL of acetate buffer (20 mM). The samples were loaded onto the cartridge using a solid phase extraction manifold at a rate of 2–3 mL/min by adjusting the vacuum. After the sample was loaded, the cartridge was eluted with 4 mL of water and dried under a gentle vacuum. Neutral PFAS were eluted first using 3 mL of 20% isopropyl alcohol in methanol twice (called fraction 1 or F1), followed by elution of the ionic PFAS using 3 mL of 1% ammonia in methanol twice (F2). The latter fraction was dried under a stream of N2, reconstituted in methanol, and combined with F1. The combined eluate was then cleaned using approximately 100 mg of Envi-Carb, filtered using a 0.2 μm centrifugal filter, transferred into a 1 mL PP vial, and spiked with internal standards (IS) prior to analysis.

The SIP disk samples and blanks were packed into 33 mL individual accelerated solvent extraction cells and spiked with SS for both ionic and neutral PFAS before extraction. Neutral PFAS were extracted using a mixture of petroleum ether and acetone (5:1, v/v; F1), while ionic PFAS were extracted using methanol (F2). Fraction 1 was concentrated using a gentle stream of N2, and the solvent was exchanged to ethyl acetate before being transferred into a 1 mL PP vial and spiked with IS (Table S2). Fraction 2 was loaded onto a precleaned Envi-Carb cleanup column and eluted with 3 mL of methanol. The resulting solution was concentrated to approximately 0.3 mL, filtered using a 0.2 μm centrifugal filter, transferred into a 1 mL PP vial, and spiked with IS before analysis.

Details on the instrumental methods used in the PFAS analyses were previously reported.4,42 A brief description is provided here, and a list of 41 target analytes is provided in Table S3. FASAs and ionic PFAS were analyzed by high-performance liquid chromatography coupled to a triple quadrupole mass spectrometer (LC-MS/MS; Agilent 1290 Infinity II UPLC—6470 QQQ-MS), operated in negative electrospray ionization mode. A gas chromatographic mass spectrometer, operated in the positive chemical ionization mode (Agilent 7890 GC—5977B PCI-MS), was used to measure the neutral PFAS, including FASE, FTOH, FTAc, and FTMAc. Detailed instrumental parameters are included in Tables S4–S7. The sulfonamides (i.e., MeFOSE, EtFOSE, MeFOSA, and EtFOSA) were not reported in the aqueous samples (precipitation and lake water) due to the low recoveries.

QA/QC

A procedural blank and a matrix spike were processed with each batch of 6–10 samples. Travel blanks and procedural blanks were treated using the same protocol as that for the corresponding field samples. Procedural (n = 9) and travel blanks (n = 11) consisted of 0.5 g of XAD with a PUF disk for the air samples. Procedural blanks for aqueous samples (precipitation and lake water) consisted of 500 mL of Optima grade LC/MS water, and travel blanks consisted of 50 mL of LC/MS methanol. For matrix spikes, a solution containing 5 ng of each ionic PFAS analyte and 20 ng of each neutral PFAS was added prior to extraction. Recoveries of SS and matrix spikes for ionic PFAS were within the range of 62–121 and 50–140%, respectively (see Tables S8 and S9). The neutral PFAS concentrations in air were corrected using appropriate SS due to lower mass recoveries (Table S9). Blank subtraction/corrections on a mass basis were applied for all samples using procedural blanks on a batch-specific basis. Blanks represented 2.0, 3.7, and 0.20% of the total PFAS for precipitation, lake water, and air samples, respectively. Laboratory and field duplicates were processed for precipitation and lake water (Table S10). The method detection limits (MDLs) were calculated based on the US EPA method;43 compounds with masses below the MDL were considered nondetected (Table S8). For mass budget flow calculations, concentrations below the MDL were replaced with the 1/2 MDL for the four selected compounds (PFBA, PFBS, PFOA, and PFOS). Additional information on the QA/QC procedures is presented in the Supporting Information.

Statistical Analysis

OriginPro 2017 and Microsoft Excel 2016 were used. Analyses of variance were performed on logarithmically transformed concentrations including Tukey’s post hoc and Student’s t tests. A significance level of 5% was used. Calculations used in the mass budget analysis are provided below and detailed in the Supporting Information.

Mass Budget Analysis

Mass budgets were calculated for four individual compounds (i.e., PFBA, PFBS, PFOA, and PFOS). The flow resulting from the deposition of PFAS in precipitation to the lake surface (kg yr–1) was calculated using the following equation

graphic file with name es3c10098_m001.jpg 1

where Cprecip is the volume-weighted mean of PFAS concentrations in precipitation for 2021 and 2022 (ng L–1), p is the precipitation rate (m/year), and A is the lake’s surface area (km2). The relative standard error was calculated using the formula described in Venier and Hites.44 Concentrations for precipitation collected in 2017–2018 from Canadian sites, Evansville and Point Petre, were used for Lake Huron and Ontario, respectively.25 The flow from the dry deposition of PFAS in particles to the lake surface (kg yr–1) was calculated as follows

graphic file with name es3c10098_m002.jpg 2

where Cp is the mean of PFAS concentrations on airborne particles (pg m–3), vd is the deposition velocity of particles (cm s–1), and 3.16 × 10–4 accounts for units conversions. Although in the literature vd ranged from 0.01 to 0.8 cm s–1, here, we used 0.2 cm s–1 for consistency with previous studies.45,46 Because SIP-PAS captured both gas-phase and particles, we used the particulate associate fraction (Φ) from Ahrens et al.47 to estimate the portion of mass on SIP disks attributable to gas-phase and to particles

graphic file with name es3c10098_m003.jpg 3
graphic file with name es3c10098_m004.jpg 4

where CSIP is the concentration of the target chemical in the SIP disk, Cg is the concentration in the gas phase (pg m–3), and Cp is the concentration in the particle phase. The value of Φ was listed in Table S11.47 Concentration data in air from Sleeping Bear Dunes and Cleveland were used for Lake Huron and Ontario, respectively.

The air–water exchange flow of PFAS across the air–water interface (kg yr–1) was calculated using the following equation

graphic file with name es3c10098_m005.jpg 5

where Fvol is the flow due to volatilization out of the lake (kg yr–1), and Fabs is the flow due to absorption from the air into the lake (kg yr–1). Cg is the PFAS concentration in the gas phase (pg m–3) as estimated from the passive samplers, Cw is the dissolved-phase PFAS concentrations in Great Lakes water (ng L–1), kol is the overall mass transfer coefficient (m d–1), H′ is the dimensionless, temperature-specific Henry’s law constant, which varies for each compound (see Table S11), and 3.65 × 10–7 and 0.365 account for units conversions. The atmospheric deposition flow of PFAS to the lake surface (kg yr–1) was calculated as the sum of the three above flows as follows

graphic file with name es3c10098_m006.jpg 6

The flow from tributaries Ftrib was estimated as follows

graphic file with name es3c10098_m007.jpg 7

where Ctrib is the PFAS concentration (ng L–1) in the tributaries, for which data were obtained from USGS data release,32 from literature,27,31,48 and from state agencies (see Table S12 for a complete list). Qtrib is the flow (m3 s–1) of each tributary obtained from USGS Current Water Data for the Nation,49 and from Bentley et al.50 that also included ungauged tributaries, and 3.16 × 10–2 accounts for units conversions.

The flow for wastewater treatment plants (WWTPs) effluents discharging directly into Great Lakes from both the Canadian and US sides was calculated as follows

graphic file with name es3c10098_m008.jpg 8

where Cwwtp is the PFAS concentration (ng L–1) in the WWTPs effluent. For the Canadian side, median values were used for each compound obtained from Gewurtz et al.,51 while for the US side, median values from Thompson et al.52 were used. Qwwtp is the flow rate (m3 day–1) of WWTP effluent discharge and 3.65 × 10–7 accounts for units conversions. Detailed flow rate values are presented in Table S13.

The loss of PFAS in the water leaving the lakes (kg yr–1) was calculated as follows

graphic file with name es3c10098_m009.jpg 9

where Cout is the dissolved phase concentration (ng L–1) in water leaving the lake at specific locations, Qout is the outflow (m3 s–1) of water, f is the fraction of PFAS in the dissolved water phase, and 3.16 × 10–2 accounts for units conversions (see Text S4 and Table S12). The estimated f values for PFBS, PFBA, PFOS, PFOA, and Σ41PFAS are 0.85 ± 0.04, 0.96 ± 0.01, 0.94 ± 0.13, 0.63 ± 0.11, and 0.87 ± 0.10 based on results from a previous study.53

Only outflow and sedimentation were considered as elimination mechanisms. Since degradation in the atmosphere and open water systems are minimal for PFCAs and perfluoroalkyl sulfonic acids (PFSAs),27 photodegradation and biodegradation were considered negligible in our model. Because PFAS concentrations in sediments were not measured in this study, the sediment burial flow, Fsedi, was obtained using values from Christensen et al.54

The inflow of PFAS in the water that entered the lake (kg yr–1) was calculated as follows

graphic file with name es3c10098_m010.jpg 10

where Cin is the dissolved phase PFAS concentration (ng L–1) in the water that entered the lakes, Qin is the inflow (m3 s–1), and 3.16 × 10–2 accounts for units conversions. Water flows contributing to the PFAS include those from Lake Superior via the St. Mary’s River, from Lake Michigan through the Straits of Mackinac into Lake Huron, from Lake Huron down the St. Clair-Detroit River system to Lake Erie, and from Lake Erie through the Niagara River into Lake Ontario (see Text S4 and Table S12).

The net mass transfer flow of PFAS (kg yr–1) was calculated as the difference between the output and input flows as follows

graphic file with name es3c10098_m011.jpg 11

Each parameter was assigned an uncertainty, and errors were propagated, as previously described.55 Complete equations and values for parameters are included in Text S4 and Tables S11–S14.

Results and Discussion

PFAS in Precipitation

Descriptive statistics of individual PFAS compounds measured in precipitation samples are listed in Table S15. The median Σ41PFAS concentration in precipitation across five sites was 3.1 ng/L, with a range of 2.4–4.5 ng/L, which was comparable to those in previous studies reported in the Great Lake region (see Table S16).25,27−29 PFAS concentrations in precipitation were not significantly different across the sites (Table S15). The relative homogeneity of concentrations across sites indicates that PFAS, and in particular ionic PFAS, are ubiquitous.

Notably, the median of the ∑41PFAS concentration in precipitation from Sleeping Bear Dunes and Eagle Harbor was comparable to that from Chicago, Cleveland, and Sturgeon Point. Since Sleeping Bear Dunes and Eagle Harbor are remote sites and Sturgeon Point a rural one, this suggests that PFAS levels, especially perfluoroalkyl acids (PFAAs), in precipitation, are not strongly associated with the human population density around each site.56,57 A comparable finding was reported for precipitation collected in Wisconsin in 2020 by Pfotenhauer et al.,28 and across the Great Lakes in 2006–2018 by Gewurtz et al.25 This phenomenon was not observed for precipitation for other legacy compounds [i.e., polycyclic aromatic hydrocarbons (PAHs), pesticides, and polybrominated diphenyl ethers (PBDEs)], all of which were associated with population density.23,57,58 PFAAs in precipitation may not reflect localized sources as observed for other POPs.28

PFAS concentrations were dominated by PFCAs (C4–C16), accounting for 75–95% of total PFAS. Short-chain PFCAs (C4–C6), i.e., PFBA, PFPeA, and PFHxA, were particularly abundant, accounting for 57–81% of total PFAS. PFSAs (C4–C8) represented up to 6% of total PFAS (Figures 1 and S2). Site specific differences in PFAS composition were observed. For instance, 6:2 FTS was present in samples from the Cleveland site with a high detection frequency of 72%, where it contributed more than 12% to the total PFAS. Our speculation is that the higher concentrations of 6:2 FTS detected in Cleveland may be attributed to the city’s industrial activity, as 6:2 FTS is commonly used in electronics, vehicles, and other industrial sectors, all of which are abundant in the area. PFOS was detected less often at the remote sites at Eagle Harbor and Sleeping Bear Dunes, while the more urbanized sites, Chicago, Cleveland, and Sturgeon Point, showed higher detection frequencies; however, when detected, the concentrations were comparable across all sites.

Figure 1.

Figure 1

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Composition of PFAS above the detection limit in Great Lakes precipitation and lake water. Note that there are no IADN sampling sites on the shores of Lake Huron and Ontario; hence, the bar for precipitation is missing.

PFAS in Great Lakes Water

Descriptive statistics of individual PFAS compounds measured in water samples from the five lakes are listed in Table S17. Not surprisingly, Σ41PFAS median concentrations in lake water from Lake Ontario (11 ng/L) were the highest, followed by those in Lake Michigan (8.5 ng/L), Erie (7.3 ng/L), and Huron (5.3 ng/L), and lowest for Lake Superior (1.3 ng/L) (Table S17 and Figure 2). PFBA, PFOA, and PFOS also followed the same pattern. This spatial trend is likely due to Lake Ontario’s position downstream from the other lakes compared with Lake Superior’s position furthest upstream. In addition, Lake Ontario basin is the most urbanized and industrialized of the five Great Lakes, while the Lake Superior basin is the least.25,26,59 PFAS concentrations in Lake Erie water were comparable to those in Lake Michigan, unlike other studies that found that PFOS and PFOA concentrations in Lake Erie were higher than those in Lake Michigan.26,60

Figure 2.

Figure 2

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Concentrations of total PFAS (Σ41PFAS) in Great Lakes water. The maximum bar height is 48 ng/L, and all other bars have been scaled accordingly.

Similar to precipitation samples, PFCAs also dominated the lake water profiles, accounting for 66–100% of Σ41PFAS. Short-chain PFCAs (C4–C6) accounted for 47–87% of Σ41PFAS (Figure 1), while PFSAs contributed up to 22% to Σ41PFAS. It is noteworthy that the proportion of PFSAs increases in the eastern lakes, which might be due to regional source variations.

6:2 FTS, a marker of industrial activities,61 accounted for 8% of Σ41PFAS in Lake Ontario water, which might be due to specific sources to this lake. More data are needed to clarify this pattern. Perfluoroethylcyclohexanesulfonate (PFECHS), a cyclic PFAS, was detected in nearly all the lake water samples, with a detection frequency of 94%, except in Lake Superior. The median of PFECHS concentration in lake water followed the order: Lake Michigan > Ontario > Huron > Erie (p < 0.05), although the levels measured here are an order of magnitude lower than those reported by De Silva et al.60 for samples collected in 2005–2010. This decline may be due to the 3 M phase out of the production of PFECHS in 2002.60,62 Incidentally, this is a chemical mainly used in aircraft hydraulic fluids since the 1940s.

PFOS and PFOA median concentrations in Lake Huron (0.41 and 0.75 ng/L), Michigan (0.69 and 1.2 ng/L), Erie (0.36 and 0.86 ng/L), and Ontario (1.3 and 1.4 ng/L) are comparable with those in Niagara river (1.5 and 1.6 ng/L) for samples collected in 2023,63 but they are lower than those from previous studies that used samples collected in 2005–2017.25,60,64 For comparison, Gewurtz et al.25 found a median PFOS level of 5.0 ng/L in Lake Ontario and 2.1 ng/L in Lake Erie for samples collected between 2006 and 2017. PFOA median levels were reported as 2.3, 1.6, 0.012 ng/L (LOQ), and 0.61 ng/L in Lakes Ontario, Erie, Huron, and Superior, respectively.25 The low levels and detection frequencies for Lake Superior (<0.21 ng/L-MDL- and 0.16 ng/L) are consistent with previous results.25 Muir and Miaz reported significant declines in PFSAs and long-chain PFCAs for 2004–2017 in the Great Lakes region (Table S16).65 The decline of PFOS and PFOA in lake water is not surprising given the phase-out of these two compounds in the North American market starting in the early 2000s, together with other long-chain PFCAs as well as some of their precursors.66

The median levels of PFBA are 0.63, 1.2, 1.3, 1.7, and 2.5 ng/L in Lake Superior, Michigan, Huron, Erie, and Ontario, respectively, which are comparable to values (0.64, 1.3, 1.7, and 3.5 ng/L in Lake Superior, Huron, Erie, and Ontario) reported by Gewurtz et al.25 PFBA was the most abundant compound detected in lake water, as well as in precipitation. This indicates that as a substitute for long-chain PFAS, PFBA is still in the market and continuously discharges into the environment. The presence of PFBA in the environment could be the result of breakdown from precursors as well as other sources including chlorofluorocarbons replacement compounds with C4 chains.67 This hypothesis of its sources aligns with the increasing trend of PFBA in precipitation at three IADN sites on the Canadian side reported by Gewurtz et al.25 and the relative uniform distribution of PFBA in the atmosphere observed by Wong et al.68

Although the PFAS pattern in precipitation was similar to that observed in lake water samples, a few differences are worth noting (Figure 1). PFSAs in water samples from Lakes Michigan and Erie accounted for >20% of the total PFAS, but only ∼5% for precipitation. While some of these differences might be due to the presence of other sources of PFSAs to these two lakes (e.g., WWTPs discharges, tributary loadings),31,69 it is also important to note that a few compounds (i.e., PFOS, 6:2 FTS, and FBSA) other than PFCAs were detected in precipitation with low detection frequencies of 11, 4.5, and 39%. If one were to consider all of the concentrations below the MDL, these differences are likely to be less pronounced. The absence of PFOS from lake water might be explained by the higher sorption capacity of PFOS and other PFAS to the sediment, as proposed by Higgins and Luthy.70

PFAS in Air

Table S18 shows the descriptive statistics for individual PFAS compounds measured with SIP disk passive air samplers, which represent both gas and particle phases.71−75 The most abundant PFAS in the air samples were FTOHs, accounting for 55–85% of the total PFAS, while ionic PFAS represented 15–38% of the total PFAS (Figure 3). The higher percentage of FTOHs is consistent with data from previous studies for air,47,68,76,77 but it is different than that observed in lake water and precipitation. This difference is likely due to the fact that neutral PFAS, such as FTOHs, are more volatile and less water-soluble than ionic PFAS.78 Additionally, neutral precursors can be transported in the atmosphere where they are oxidized into stable ionic PFAS. Previous studies have shown that volatilization and degradation of precursors can be a prominent pathway for PFCAs in the environment, especially in rural and remote areas.79−81

Figure 3.

Figure 3

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Comparison of contributions (%) of targeted PFAS in air.

The median Σ41PFAS concentrations in air samples were significantly higher (p < 0.05) at the two urban sites (range 179–1146 pg/m3 at Chicago and Cleveland) compared to those at two remote sites (range 38–442 pg/m3; Table S18). PFAS levels, particularly those of FTOHs, were higher than those reported in Toronto by Ahrens et al.47 but comparable to the levels reported for other urban sites in Japan, India, China, and Europe (Table S16).76,82,83 The levels of ionic PFAS were similar to previous studies,76,84 but lower than a recent report by Saini et al. that found ΣPFCAs ranging from 0.128 to 781 pg/m3 and ΣPFSAs from 6.85 to 124 pg/m3 across 40 sites globally.83 In addition, the Σ41PFAS concentration in the particle phase, calculated as described above, ranged between 1.3 and 44 pg/m3, which is comparable to values reported in previous studies.47,82

Mass Budget in Lakes

The net mass transfer flows (kg yr–1) for four individual compounds (i.e., PFBA, PFOA, PFBS, and PFOS) for the Great Lakes are given in Tables S19–S22 and Figure 4. PFOA and PFOS were the most extensively researched and utilized long-chain PFCAs and PFSAs, whereas PFBA and PFBS were identified as short-chain counterparts, offering potential alternatives to PFOA and PFOS. A negative flow indicates that the lake is accumulating PFAS, while a positive flow indicates that the lake is eliminating PFAS.

Figure 4.

Figure 4

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Bar plot showing the contribution of the five main flows (outflow, sedimentation, tributaries, inflow, and atmospheric deposition) that contributed to the net mass flow (kg/year) for four selected compounds, PFBS, PFOS, PFBA, and PFOA. Lake inputs (atmospheric deposition, tributaries, and inflow) are positive, while removal processes (outflow and sedimentation) are negative. The arrows represent the direction of the net mass flow where significantly different from zero. Arrows are directed upward if negative (i.e., the lake is accumulating that chemical) and downward if positive (i.e., the lake is discharging that chemical). The length of the arrow is not representative of the magnitude of the flow.

One can analyze these results by looking at the contribution of each flow to the net flow or at the magnitude and sign of the net flow itself. We first address the contributions. Atmospheric deposition was the dominant accumulation process in Lake Superior, with its contribution declining in each downstream lake due to the increasing role of input flows, including tributaries. Inputs from tributaries were significant for Lake Michigan for all four compounds and for PFOS and PFBS in Lake Huron. WWTP contributions were always below 5%; more information about facilities discharging into the lakes and on the specific PFAS concentrations in their effluents is needed to confirm this finding.

Within atmospheric deposition, the relative contribution of each of these three components (i.e., air–water exchange, wet deposition, and dry deposition) varied according to the compound and lake. About 60% of the PFOA and 50% of the PFOS deposited from the atmosphere were in wet deposition, while the more volatile, shorter chain compounds PFBA (71%) and PFBS (78%) were deposited mostly by gas absorption. Approximately one-third of the atmospheric deposition of PFOS was deposited via dry deposition, as could be expected due to its long chain length and affinity to particles, while smaller portions were contributed by gas absorption (7–34%). Other compounds such as PBDEs and PAHs,44,85 are known primarily to be contributed by wet deposition. Few studies exist on the atmospheric deposition of PFAS and while these have been done elsewhere, they suggest precipitation is a main mechanism for removing PFAS from the atmosphere.18,19,86 Additional studies on the partitioning of PFAS between the gas and particulate phases are needed to confirm these findings.

Removal processes include sedimentation and outflows from the lakes, with outflows as the dominant process for all compounds, except for PFBS in Lake Erie. The sedimentation flow of PFBS was much higher than that of PFOS in Lake Erie, which was not consistent with the higher tendency of PFOS to strongly adsorb to particles/sediment and a lower water solubility than PFBS.53,87 More data are needed to confirm this finding. Sedimentation contributed on average 9.4% of PFOA removal across the lakes, similar to the 6.7% reported for Lake Ontario in 1998.27,30

With respect to net flows (Figure 4 and Tables S19–22), all four compounds (i.e., PFBA, PFBS, PFOA, and PFOS) are accumulating in Lake Superior. Net flows were significant for PFOA only in Lake Ontario (218 ± 106 kg/year) and in Lake Superior (−12.5 ± 10.3 kg/year). Lake Michigan is accumulating PFBA and PFSAs (the latter represented by PFOS and PFBS).

Considering the large uncertainties in these estimates, Lake Huron is close to steady state for PFBS and PFOA, and accumulates PFOS (−77.1 ± 25.0 kg/year) and PFBA (−231 ± 56.9 kg/year). The net flows for Lake Erie were significant only for PFBS (58.8 ± 35.7 kg/year), suggesting that Lake Erie is close to steady state for PFBA, PFOS, and PFOA, while it seems to be eliminating PFBS.

Lake Ontario is discharging all four PFAS compounds via the St. Lawrence River outflow and sedimentation, although the net flow for PFBA was not significant (122 ± 135 kg/year). The net flows for Lake Erie and Ontario are higher than the estimates for the upper Great Lakes, especially for short-chain PFAS, which suggests that additional sources may be contributing PFAS here or that these lakes are still not at steady state.

To put these budgets in a broader context and compare them with well-known legacy chemicals, the wet deposition of the four chemicals combined to Lake Michigan was 104 ± 12.0 kg yr–1 which is lower than the value reported for all PBDEs in 2006 (250 ± 66 kg yr–1),44 and higher than that for PCBs (24.6 ± 3.1 kg yr–1) in 2015. It should be noted that also PBDEs and PCBs are the sum of 40 or more individual compounds.55

Data Gaps and Limitations

Although we strived to be as comprehensive as possible in our calculations, model complexity, and data availability did not allow us to be exhaustive. Aerosol injection into the atmosphere from the surface of the lake should be considered in the calculation of PFAS exchange at the air–water interface.75 Several potential sources of PFAS to the lakes are not included. For example, PFAS have been found at high concentrations in biosolids,88 from which they can leach into surface water from amended fields and ultimately reach the lakes. Groundwater is another possible significant source of PFAS. Additionally, PFAA precursors present in air, water, and biosolids may need to be considered as these are known to produce PFAAs via photolytic or microbial degradation.27 Also, precursor transformation to these PFAAs will confound the mass-balance interpretation.

This study also highlights several data gaps that prevent a more robust understanding of the cycling of PFAS in the Great Lakes. We stress that discharge data are lacking for many Great Lakes tributaries, and tributary concentration data are generally scarce and lack temporal and spatial resolution. Our finding of high abundance of PFBA in precipitation suggests that ultrashort chain PFCAs, i.e., trifluoroacetic acid and perfluoropropionic acid, which have been detected in precipitation and surface water samples,89−91 should be monitored in future studies. Moreover, more lake water samples are needed from different seasons due to the effects of stratification and temperature. Sedimentation rates for short chain PFAS should be updated; until then, sedimentation flows for Lake Erie and Ontario, especially for PFBS, should be interpreted with caution. Finally, the partitioning of PFAS in air needs to be further studied, possibly including active samplers alongside passive samplers,76,92 to improve the strength of estimates of dry deposition and air–water exchange flows.

Despite these limitations, our findings suggest that atmospheric deposition plays an important role in delivering PFAS to the Great Lakes, in particular for short chain compounds and in Lake Superior, where atmospheric deposition contributes more than 90% of inputs for PFBA. The upper lakes seem to accumulate PFAS and possibly act as continued source to the lower lakes (i.e., from sediments).31 As part of ongoing efforts to understand the fate and behavior of PFAS in the environment, continued measurements of PFAS in the atmosphere, as well as in tributary water and lake surface water, are recommended.

Acknowledgments

The authors thank the U.S. Environmental Protection Agency’s Great Lakes National Program Office for Great Lakes Restoration Initiative funding via Cooperative Agreement Number GL00E02730 (Derek Ager, Program Manager, and Ben Alsip, Project Officer); Derek Ager, who collected the lake water samples from Lakes Superior and Michigan; the site operators and Team IADN at Indiana University Bloomington for their ongoing efforts; Environment and Climate Change Canada for collecting water from Lakes Huron, Erie, and Ontario; Steve Corsi at the USGS for the tributary water data; Sarah Gewurtz and Shirley Ann Smyth for valuable inputs on tributaries and WWTPs data; and Cooper Sykes for the maps. The precipitation data used in this study were obtained from the U.S. EPA and is the only data set formally reviewed by the EPA. This article has not been formally reviewed by EPA and the views expressed in this document are solely those of the authors and do not necessarily reflect those of the Agency. EPA does not endorse any products or commercial services mentioned in this publication. We also thank the anonymous reviewers for their recommendations that helped significantly improve this manuscript.

Supporting Information Available

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.3c10098.

  • Details about SIP disk protocol; mass budget analysis; determination of the duration for precipitation collection; sampling bucket rinse test for cleaning; details of five U.S. IADN sites for precipitation and air sample collection and lake water samples location; targeted PFAS analytes and SS and IS; instrument parameters for LC-MS/MS and GC/MS; blanks, matrix spike recoveries, and MDLs in water and air samples; surrogate recoveries; lab/field duplicate; parameter values used for the mass budget calculation; summary of previously published data on PFAS in precipitation, Great Lakes’ water and air samples; descriptive statistics of PFAS concentrations in precipitation, Great Lakes’ water and air samples; mass transfer flow (kg/yr) for all five individual Great Lakes for PFBA, PFOA, PFBS, and PFOS; and contribution of individual detected PFAS in precipitation in five IDAN sites (PDF)

  • Concentrations of PFAS (total PFAS, PFBA, PFBS, PFOA, and PFOS) and discharge rate in tributaries, tributary flow, wet deposition flow, dry deposition flow, air-water exchange flow, and inflow and outflow (XLSX)

The authors declare no competing financial interest.

Supplementary Material

es3c10098_si_001.pdf (1.6MB, pdf)
es3c10098_si_002.xlsx (63.8KB, xlsx)

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