Achieving sub-part per trillion trace level PFAS quantification in drinking water using an optimized fast flow solid-phase extraction and UPLC-MS/MS method

Deepak Timalsina; Bhargavi Srija Ramisetty; Michael Zhuo Wang

doi:10.1371/journal.pwat.0000501

Abstract

Per- and poly-fluoroalkyl substances (PFAS) are not efficiently degraded and hence cycle through the environment, persist for a very long time, accumulate in living organisms, and cause potential health and ecological risks. PFAS monitoring in the drinking water relies on solid phase extraction (SPE) for sample concentration and ultra-performance liquid chromatography-tandem mass spectrometry (UPLC-MS/MS) for sensitive and specific detection, but conventional methods suffer from long processing time, inadequate sensitivity for sub-part per trillion (ppt) trace level detection, and high cost. In this study, we aimed to develop and validate an optimized fast flow SPE method to achieve sub-ppt trace level quantification of 40 PFAS compounds that are of environmental concern to the US Environmental Protection Agency. The impact of several key sample preparation parameters, including N₂ drying, syringe filtration, SPE elution volume, and SPE flow rate, on the PFAS recovery was determined. These results helped inform the development of the final optimized fast flow SPE method, which was demonstrated using blank water samples spiked with trace levels of PFAS and tap water samples. The new fast flow SPE method substantially reduced the sample loading time (6 min for a 500-mL sample vs. 100 min required for normal flow SPE; 60–70 min for a 4-liter sample vs. 800 min required for normal flow SPE) without compromising the PFAS recovery for 38 out of 40 PFAS compounds and achieved sub-ppt (as low as 0.01 ppt for method detection limit) trace level quantification of PFAS in the drinking water. As a result, the optimized fast flow SPE is a viable strategy to enhance method sensitivity, increase throughput, and reduce cost for PFAS analysis and will positively impact future PFAS monitoring in the drinking water.

Citation: Timalsina D, Ramisetty BS, Wang MZ (2026) Achieving sub-part per trillion trace level PFAS quantification in drinking water using an optimized fast flow solid-phase extraction and UPLC-MS/MS method. PLOS Water 5(3): e0000501. https://doi.org/10.1371/journal.pwat.0000501

Editor: Jiafu Li, University of South Carolina, UNITED STATES OF AMERICA

Received: December 16, 2025; Accepted: February 19, 2026; Published: March 30, 2026

Copyright: © 2026 Timalsina et al. This is an open access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.

Data Availability: All relevant data are within the paper and its Supporting Information files.

Funding: This study was supported in part by a pilot project award from the University of Kansas General Research Fund (2504120 to MZW), the Center for Targeted PFAS Analysis (PFAS Lab) at the University of Kansas, and a research grant (G25AP00168 to MZW) from the United States Geological Survey. The funders had no role in study design, data collection and analysis, decision to publish, or preparation of the manuscript. DT and MZW received salary from the University of Kansas and the grant from the United States Geological Survey.

Competing interests: The authors have declared that no competing interests exist.

Introduction

Per- and poly-fluoroalkylated substances (PFAS) consist of at least one perfluoro (-CF₃) or polyfluoro (-CF₂) alkyl group in their chemical structures [1, 2] and common PFAS contain a perfluoroalkyl chain (C_nF_2n+1) with a polar terminal functional group, such as -COOH, -SO₃H, -SO₂NH₂. PFAS are called “forever chemicals” due to their resistance to degradation in the environment. This higher thermodynamic, metabolic, and environmental stability is partly due to the high C-F bond strength (488 kJ/mol vs 413 kJ/mol for the C-H bond) in the alkyl chain [3]. The C-F chain lengths also influence the surfactant properties, making it a good repellent to oils and greases. This makes PFAS suitable for use in protective coatings [4–7], textiles [8], paints [9], adhesives [10], food wrappers, cookware, and cosmetics [11–14]. PFAS are widely used in firefighting foam [15,16], semiconductors, insecticides, and space materials [17]. PubChem, one of the largest chemical database collections, recorded over 7 million PFAS as of September 2022 [18]. Due to their extensive usage and persistence, these compounds are distributed widely in the environment. Once released, these compounds can travel through the water system and atmospheric pathways, leading to contamination in places far away from the sources, such as Arctic and Antarctic glaciers [19]. Recent studies report the prevalence of PFAS in soil [20], water [21,22], food, and consumer products from parts per trillion (ppt) to higher parts per billion (ppb) levels [23] and the United States Geological Survey (USGS) reported that individual PFAS concentrations ranged from 0.025 to 319 ppt (ng/L) in >600 point-of-use tapwater samples and estimated that at least one PFAS could be detected in about 45% of US drinking-water samples [24].

PFAS are of increasing concern due to their toxicity and potential for bioaccumulation. Studies have shown that these PFAS, such as PFOA and PFOS, are linked to several health issues, such as fetal development, suppression of vaccine response, thyroid disease [25,26], kidney cancer, and testicular cancer [27]. A study done in mice showed that PFOA affects metabolic pathways, accumulates in the brain, and alters the neurotransmitter level and synaptic formation [28]. Even low concentrations of PFAS are of concern due to their accumulation inside the human body. For example, PFOA has been found to accumulate in the body due to renal tubular reabsorption and enterohepatic recirculation [29–32]. The half-life of PFOA in humans has been estimated to range from 1.2 to 14.9 years [33–35].

Due to these potential health risks, the United States Environmental Protection Agency (US EPA) announced the National Primary Drinking Water Regulation (NPDWR) on April 10, 2024 for six PFAS and established legally enforceable levels, known as Maximum Contaminant Levels (MCLs) in drinking water [36]. The MCLs are 4 ppt (or ng/L) for PFOA and PFOS, and 10 ppt for PFHxS, PFNA, and HFPO-DA (commonly known as GenX chemical), and a Hazard Index of 1 for the combined and co-occurring levels of PFHxS, PFNA, HFPO-DA, and PFBS. Importantly, EPA also finalized health-based, non-enforceable Maximum Contaminant Level Goals (MCLGs) to be zero for PFOA and PFOS, 10 ppt for PFHxS, PFNA, and HFPO-DA, and a Hazard Index MCLG of 1 for the mixtures. As of January 12, 2026, the European Union (EU) has implemented a new EU-wide rule on the systematic monitoring of PFAS in drinking water to ensure compliance with the new EU limit values under the recast Drinking Water Directive [37]. The new rule established clear limit values to ensure drinking water contains no more than 500 ppt of total PFAS and 100 ppt of the sum of 20 individual PFAS [38]. Stricter limits are expected to follow in the coming years as individual EU member states like Germany and Denmark have opted for lower limit values to better protect against PFAS [39]. Similarly, the World Health Organization (WHO) issued provisional guidance values of 100 ppt individually for PFOA and PFOS in drinking water and a combined 500 ppt for total PFAS based on the 29 PFAS compounds [40]. Although exact details of these regulations are subject to change, it is generally expected that public water systems must monitor and mitigate PFAS contamination and provide the public with information on the PFAS levels in their drinking water. To meet these demands (especially the MCLG of zero for PFOA and PFOS), a sensitive, reliable, and cost-effective analytical method to detect and quantify sub-ppt trace-level PFAS in the drinking water is urgently needed.

EPA has released two analytical methods based on solid-phase extraction (SPE) and liquid chromatography-tandem mass spectrometry (LC-MS/MS) that are suitable for PFAS monitoring in the drinking water. The EPA Method 537.1 (Version 2.0; released in March 2020) was designed specifically for drinking water, targeting 18 PFAS. It utilizes SPE at the flow rate of 10–15 mL/min to preconcentrate the PFAS from a 250 mL starting sample volume, followed by N₂ drying of the eluate (<65°C) before LC-MS/MS analysis. In contrast, the EPA Method 1633A (released in December 2024) is a more comprehensive method covering diverse matrices (aqueous, solid, biosolids, and tissues), targeting 40 PFAS [41]. This method uses 500 mL starting sample volume with a reduced SPE flow rate of 5 mL/min without the use of N₂ drying before LC-MS/MS analysis. It utilizes the isotope-dilution or extracted internal standard quantification technique, which is recognized as the gold standard technique for achieving accurate quantifications.

Previous studies have reported the detection limit as low as 3.2 ppt for some of the PFAS using nano-electrospray ionization and high-resolution mass spectrometry (Nano-ESI-HRMS) following Method 537.1 [42]. The best detection limits recorded in other studies using either Method 1633A or 537.1 were about 1.04 - 20.98 ppt for some of the PFAS [43,44]. Although these reported detection limits are generally sufficient for monitoring MCLs for individual PFAS in drinking water, they are inadequate for MCLGs or monitoring mixtures which require sub-ppt detection limits. Furthermore, both EPA methods are quite time-consuming and laborious, requiring long SPE sample loading time (e.g., 17–25 min for a 250-mL sample using Method 537.1 and 100 min for a 500-mL sample using Method 1633A), substantially reducing the analytical throughput. To advance and expedite the PFAS analysis, several alternatives have been investigated, such as the total oxidizable precursor assay (TOP), analysis of adsorbable organofluorine and extractable organofluorine (AOF/EOF) [45]. These assays are non-specific, cannot speciate or quantify individual PFAS, and are associated with interference from other fluorinated compounds such as pesticides and pharmaceuticals. Paper spray mass spectrometry (PS-MS) also offered rapid analysis with higher sensitivity without the need for extensive sample preparation and chromatographic separation [46]. However, a lack of separation before the MS analysis can lead to analytical challenges when compounds with similar molecular weights are present in the sample. For instance, PFOS and taurodexycholic acid (TDCA) exhibit overlapping MRM transitions, resulting in a potential interference if not resolved properly before analysis [47].

In this study, we aimed to evaluate the impact of several key sample preparation parameters relevant to the EPA methods, including N₂ drying, syringe filtration, SPE elution volume, and SPE flow rate, on the PFAS recovery and develop an optimized method to improve method sensitivity and analytical throughput for simultaneous detection and quantification of sub-ppt trace levels of 40 PFAS (Table A in S1 Text) in the drinking water. By employing fast flow SPE and increasing starting sample volume, the optimized method was compared to the reference EPA methods using spiked water and tap water samples for validation and demonstration.

Experimental Section

Reagents, chemicals, and standard solutions

The mixtures of 40 target PFAS standards, 24 extracted internal standards (EIS), and 7 non-extracted internal standards (NIS) (Tables B-D in S1 Text) were purchased from Wellington Laboratories Inc. (ON, Canada). Optima LC/MS grade methanol, acetonitrile, water, ammonium acetate (99%), glacial acetic acid (99%), formic acid (99%), and ammonium hydroxide solution (30%) were purchased from Fisher Scientific (ON, Canada). Calibration stocks were prepared by diluting the obtained mixture of target PFAS standard solutions first two-fold and further diluted down to the instrument detection limit (IDL) to obtain ten-point calibrations. The corresponding calibration concentrations (Table E in S1 Text) were prepared in the matrix composed of 4% water, 1% ammonium hydroxide, and 0.625% acetic acid in methanol. EIS and NIS were diluted in methanol from the purchased stock before being added to the calibrations (Table F in S1 Text).

UPLC-MS/MS method for PFAS detection

PFAS were separated using a Waters Acquity UPLC BEH C18 column (1.7 µm, 2.1x50 mm) coupled with an Acquity column in-line filter (0.2 µm) operated at 40 °C. A 2 µL injection volume was chosen to minimize the matrix effect. This analysis employed an LC mobile phase comprising 2 mM ammonium acetate in a 95:5 (v/v) water/acetonitrile (A) and acetonitrile (B) in a gradient flow, with a run time of 12 minutes. The LC gradient program is shown in Table G in S1 Text. The PFAS were analyzed in a Waters Xevo TQ-S triple quadrupole mass spectrometer in a negative electrospray ionization (ESI) mode. Stock solutions of target PFAS (6.25-25 ppb (or ng/mL) were prepared by diluting 40-fold from the purchased stock and injected multiple times to optimize the MS parameters for best signal intensity. Optimized MS/MS parameters for all compounds were as follows: source temperature 140°C, desolvation temperature 500 °C, capillary voltage 0.7 kV, cone gas 70 L/h, and desolvation gas 800 L/h. IntelliStart-optimized collision energy and cone voltage were used to build multiple reaction monitoring (MRM) mass transitions for each compound of interest. Five out of 40 compounds had only one transition present, similar to previously reported methods [48]. For the PFAS with multiple transitions, the higher intensity one was selected as quantification ions, and the other as confirmation ions. The details of MRM and their corresponding EIS and NIS are tabulated in Tables H-I in S1 Text.

Sample preparation and extraction

A 500 mL aliquot of water (Optima LC/MS grade) was used as blank water for PFAS analysis. Blanks were spiked with the spiking concentrations 2–5 times their respective estimated method detection limit (eMDL). An EIS was added to each sample to monitor and correct for SPE extraction recovery. SPE cartridge (Oasis PFAS WAX 200 mg 60 μm/GCB 50 mg) was obtained from Waters Corporation and preconditioned with 15 mL of methanolic ammonium hydroxide (1% v/v) followed by 5 mL of 0.3 M formic acid. The entire sample was passed through the cartridge under vacuum at approximately 5 mL/min, unless noted otherwise. The cartridges were washed with 5 mL of water twice, followed by 5 mL of a 50% v/v 0.1 M formic acid and water. The PFAS were then eluted with 5 mL of methanolic ammonium hydroxide (1% v/v) solutions. The NIS was added to the final extract, and 100 μL of sample was transferred to a polypropylene vial for UPLC-MS/MS analysis. The overall experimental flow chart (Protocol 1-existing EPA Method 1633A) is shown in Fig 1. The recovery of both EIS and target PFAS were calculated by using Equation 1 and was expressed in average recovery ± SD.

Download:

Fig 1. Overall experimental flowchart of SPE-based PFAS analysis in aqueous samples.

Optimized protocols (Protocols 2, 3, and 4) are highlighted in bold, with the optimized parameters representing modifications from Protocol 1 that represents the EPA Method 1633A. (Created in BioRender. Wang, Z. (2026) https://BioRender.com/v4lomj1).

https://doi.org/10.1371/journal.pwat.0000501.g001

(1)

Instrument detection limits, method detection limits, and method limits

The instrument detection limit was defined as the concentration where the signal-to-noise ratio is at least 3:1 across multiple injections. The peak height, peak area, and relative response ratio (RR) or relative response factor (RF) of the signals were calculated using the MassLynx software v4.2 The method detection limit was calculated according to the method outlined in 40 CFR Part 136, Appendix B, Revision 2 [49]. In PFAS-free HDPE bottles, four replicates of each blank water and blank water spiked with target PFAS at levels near 5 times of estimated method detection limit (eMDL) were prepared and subjected to SPE. The eluted extract was subsequently analyzed by UPLC-MS/MS analysis.

The MDL was calculated using the formula below:

(2)

where t_n−1, _1−α=0.99 is Student’s t value for the single-tailed 99th percentile confidence level for n-1 degrees of freedom; and S is the sample standard deviation of replicate spiked samples (n replicates). For a study using quadruplets, n = 4 and MDLs = 4.541 × S.

Method limits (ML) or method quantification limits (MQL) were calculated by multiplying MDL by 3.18. [50]

(3)

Method optimizations to improve MDL, ML, and throughput

Concentration by N₂ drying.

The eluted extracts after SPE and syringe filtrate from Protocol 1 were evaporated at room temperature and 60°C under a gentle nitrogen stream, and the sample was reconstituted with 500 μL of methanolic ammonium hydroxide (1%) and this method was designated as Protocol 2. 50 μL of NIS was added, and 100 μL of the sample was taken for UPLC-MS/MS analysis. Recovery was calculated for both EIS and target PFAS using Equation 1 as in Protocol 1. A quadruplet of samples was analyzed, and new MDL and ML were determined.

Syringe filtration of SPE extract.

The syringe filters (25 mm, 0.2 μm, Nylon) were purchased from Fisher Chemical, and 5 mL syringes were obtained from Becton, Dickinson and Company (NJ, USA). Non-potable water samples (n = 5) were spiked with EIS, and SPE was carried out. To assess the impact of syringe filtration, the NIS was added before syringe filtration for the first set of samples, whereas in the second set of samples, NIS was added after syringe filtration. The recovery was calculated on both sets of samples and compared for all 24 EIS standards of PFAS (Table C in S1 Text).

Elution volume optimizations.

The blank water samples were spiked with EIS, processed using SPE, and eluted using three different volumes of methanolic ammonium hydroxide (1% v/v), i.e., 1 mL, 2.5 mL, and 5 mL. Each eluate was analyzed separately, and the recoveries of the 24 EIS standards and 40 target PFAS were determined and compared. The minimum volume with acceptable recovery for both EIS and target PFAS was selected for further experiments.

Rapid sample preparation with fast flow SPE.

The 500 mL blank water samples (n = 3) were first spiked with a mixture of four PFAS compounds (PFOA, PFOS, PFBA, and PFBS; each at 10 ppb), and processed via SPE at a faster flow rate than recommended in the EPA methods. The flow rates of 35–40 mL/min and 85–90 mL/min were used to screen across triplicate samples, and their recoveries were calculated using Equation 1 to assess the effect of fast flow SPE on PFAS recovery.

Improving MDL and ML with fast flow SPE and larger sample volume.

A 4 L blank water (n = 4) was spiked with EIS and target PFAS, aiming for 5 times of eMDL. The sample was loaded into the SPE after preconditioning of the cartridge, at a flow rate of 80–85 mL/min with the help of high vacuum (Protocol 3). The entire 4-liter sample was loaded within 50–60 min, significantly less than the time taken by a 500 mL sample using conventional SPE flow rate (5 ml/min) as recommended by the current EPA Method 1633A. To further improve MDL and ML, N₂ drying of SPE extract at 60°C was included and designated as Protocol 4. For both Protocols 3 and 4, the EIS and target PFAS recoveries were calculated and the new MDL and ML were determined using quadruplet samples.

Data and statistical analysis

The data comparison regarding recovery, IDL, MDL and ML were done by the descriptive statistics (mean and standard deviation) and Student’s t-tests were computed using MS Excel Version 16.78.3 and GraphPad Prism 7.05.

Results and discussion

UPLC-MS/MS detection of forty target PFAS compounds

The developed and optimized UPLC-MS/MS method specifically detected all 40 target PFAS compounds in a mixture, including the potential interferant taurodeoxycholic acid (TDCA), in a single run of 12 min (Fig 2), which is an improvement over the longer run times of 27, 15, and 23 min achieved in previous reports [51–54]. Short chain carboxylic acids, fluoroether-carboxylic acids, and fluorotelomer carboxylic acids such as PFBA, PFPeA, PFMPA, PFMBA, and 3:3 FTCA eluted earliest (RT < 4 min), whereas long chain neutral sulfonamides and sulfonamidoethanols such as NMeFOSA, NEtFOSA, NMeFOSE, and NEtFOSE eluted last (RT > 9 min) on the reversed-phase C18 analytical column. The observed retention times were consistent across multiple injections (CV < 5%). The smaller peak before the larger linear isomer peak of several PFAS, such as PFOSA, NMeFOSA, NEtFOSA, NMeFOSE, NEtFOSE, NMeFOSAA, and NEtFOSAA, suggested the detection of their corresponding branched chain isomers, which would allow us to monitor the total isomers in environmental samples. This method was also able to achieve about 2 min of retention time separation between TDCA and PFOS (Fig 2A), which share the exact same MRM and hence require chromatographic separation for specific detection.

Download:

Fig 2. Representative UPLC-MS/MS chromatograms of 40 target PFAS compounds in a mixture.

Panel A includes 22 PFAS with taurodeoxycholic acids (TDCA) to demonstrate the separation from PFOS (peak 16) and panel B includes the rest of 18 PFAS. Difficult-to-see peaks and peaks for linear PFAS compounds are indicated by the arrows and traces with different colors.

https://doi.org/10.1371/journal.pwat.0000501.g002

Calibration linearity and instrument detection limit

The calibration required two types of internal standards (IS), i.e., extracted internal standard (EIS) and non-extracted internal standard (NIS), to achieve isotope-dilution quantification and monitor SPE extraction recovery simultaneously. EIS and NIS are the stable isotope-labelled version of the same PFAS compounds and are sometimes used as surrogate IS as stable isotope-labelled PFAS are not always available for every target PFAS. The lists of target ions of PFAS compounds with their corresponding EIS and NIS are included in Tables H-I in S1 Text. EIS were spiked into samples and co-extracted with target PFAS to monitor extraction recovery, whereas NIS were spiked into final extracts prior to UPLC-MS/MS analysis to enable quantification using the isotope-dilution technique.

At least ten calibration concentrations for each target PFAS compound were injected, of which, at least seven calibrations were within the calibration range and the lowest standard was at or below the limit of detection (Table E in S1 Text). The linearity of the calibration was evaluated by calculating the relative standard deviation (RSD) of the response ratio (RR) or response factor (RF) obtained from the calibration samples. The RSD was well <15%, demonstrating the linearity (Table J in S1 Text). The compounds 3:3 FTCA and 7:3 FTCA displayed higher RSD than other compounds, likely due to their lower sensitivity than other compounds, consistent with the previous study [53].

Instrument detection limit (IDL) ranged from 0.05 to 5 ppb (Table 1). The IDL was less than 0.2 ppb for the classes of carboxylic acids, sulfonic acids, sulfonamidoacetic acids, and sulfonamides. The highest detection limit observed was for the HFPO-DA (5 ppb). The instrument detection limit for the higher molecular weight per- and polyfluoroether carboxylic acids, such as HFPO-DA, NFDHA, and ADONA, had a 2–10-fold higher IDL compared to the rest of the other PFAS.

Download:

Table 1. Instrument detection limit (IDL) and method limits (ML) in different method conditions.

https://doi.org/10.1371/journal.pwat.0000501.t001

Method detection limit, method limit, and recoveries before optimization

The initial method detection limit was calculated by spiking Optima LC/MS-grade blank water with the concentration of 2.5 to 62.5 ppt, depending upon the compounds, aiming for 5-fold of the eMDL, and extracted using the SPE Protocol 1 described in the experimental section and in Fig 1. The MDLs ranged from 0.18 to 18.12 ppt (Table L in S1 Text) and MLs ranged from 0.58 to 57.62 ppt (Table 1). The developed analytical method was able to detect and quantify 39 out of 40 PFAS listed in the EPA Method 1633A within the acceptable recovery limit (Fig 3). The MDLs calculated were sufficient for analyzing the drinking water and other aqueous matrices within the current regulatory MCLs. The MDLs and MLs were better, or comparable to those of Gremmel et al [52]. The extraction method had a reproducible recovery and was within the limit outlined by the EPA Method 1633A. Among the 24 EIS, 22 had a recovery of 81–116% while sulfonamides NMeFOSA and NEtFOSA had a recovery of 63 and 68%, respectively, but still within the acceptable recovery range (Fig 3A). Among the 40 target PFAS, 38 had an acceptable recovery, ranging from 81-124% with an RSD of 3–19% (Fig 3B). While most of them had a recovery range from 70-135%, PFHxA had a slightly higher recovery range of 139%, while NFDHA failed to detect. The absence of signal for NFDHA could be due to the combined effect of sample loss as well as the higher instrument detection limit observed in our study (Table 1).

Download:

Fig 3. Extracted internal standard (EIS) recovery (A) and Target PFAS recovery (B) before optimization (Protocol 1) along with their corresponding recovery limits as provided by the EPA Method 1633A.

Data are shown in bar diagram as mean ± SD (n = 4).

https://doi.org/10.1371/journal.pwat.0000501.g003

Effect of N₂ drying on MDL, ML and recovery

Adding drying steps and reconstituting with the elution solvent after elution (Protocol 2) resulted in the improvement of the MDL and ML values. The 5 mL extracted sample was dried and reconstituted with 500 μL, concentrating the extract by 10-fold. To equalize MRM signals, the water blank was spiked with a 10-fold lower concentration of target PFAS and EIS than the original method (Protocol 1) in order to calculate MDL and ML. Results showed that further concentration of extracted samples by nitrogen drying lowered the MDL (Table L in S1 Text) and ML (Table 1) by 2–10-fold compared to without nitrogen drying and it also increased the recovery of NFDHA to be within the acceptable limits (Fig 4B). However, drying negatively impacted the recovery of some of the neutral PFAS such as sulfonamides (PFOSA, NMeFOSA, and NEtFOSA) and sulfonamidoethanols (NMeFOSE and NEtFOSE) in the EIS recovery (Fig 4A) and fluorotelomer carboxylic acids (3:3 FTCA, 5:3 FTCA, and 7:3 FTCA) in the target PFAS recovery (Fig 4B). After N₂ drying at 60 °C, these PFAS had only 2–4% recovery, well below the EPA recovery limit (Fig 4), which also contributed to their large variability between replicates. Drying at room temperature still impacted sulfonamides and sulfonamide ethanol while improving the recovery of fluorotelomer carboxylic acids (Table M in S1 Text). In contrast, without N₂ drying, their recoveries were 50–80%, within the EPA recovery limit (Fig 3). To investigate the cause of low recovery for some PFAS after N₂ drying, PFOSA and NMeFOSE were studied individually and subject to N₂ drying under the same conditions. The recovery of these PFAS dropped by 70–96% (Table N in S1 Text) without forming the known terminal degradation product such as PFOS (Fig A in S1 Text), suggesting that the loss was likely due to evaporation rather than degradation. The low recovery of these compounds upon drying could have been caused by the semi-volatile nature of these compounds, accelerating evaporative loss during N₂ drying [55,56].

Download:

Fig 4. Comparison of(A) EIS recovery and (B) target PFAS recovery between the existing method (Protocol 1) and the method Protocol 2 (with N₂ drying) along with their corresponding recovery limits as provided by the EPA Method 1633A. Data are shown as mean ± SD (n = 4).

https://doi.org/10.1371/journal.pwat.0000501.g004

Effect of syringe filtration on PFAS recovery

Syringe filters help to remove particulates from the extract before UPLC-MS/MS analysis, which may cause damage to the column or interfere with the signal detection. The non-specific adsorption of the PFAS in the filters may cause lower recovery, impacting quantification results and underestimating method limits. Hence, we evaluated the impact of syringe filters on the recovery of PFAS compounds. There was no significant difference in the PFAS recoveries before and after syringe filtration of the SPE eluate samples that contained approximately 5 mL of 1% methanolic ammonium hydroxide (Fig 5).

Download:

Fig 5. EIS recovery of the extract eluted with 5 mL of eluent, compared before and after syringe filtration in non-potable water samples following Protocol 2.

Data are shown as mean ± SD (n = 5).

https://doi.org/10.1371/journal.pwat.0000501.g005

Previous study [57] have demonstrated that the longer-chain PFAS are more likely retained in the filters than the shorter-chain PFAS. However, the impact of syringe filtration is also affected by the volume and the types of solvent used. For example, the recovery of long-chain PFAS such as PFOS and PFNA was 0% in aqueous solvent for all volumes (1–6 mL), while the recovery of short-chain PFAS gradually increased with increasing volume. However, for the methanol, the recoveries were impacted only in the 1 mL filtered solution. This could be due to the fact that the smaller volume could leave pores unwetted due to the high surface tension of the solvent [58]. Any volume from 3-6 mL had 100% recovery, consistent with our results (Fig 5), which also indicate that the lower recoveries of sulfonamides and sulfonamidoethanols were not caused by the loss during syringe filtration.

Effect of SPE elution volume on PFAS recovery

Since N₂ drying led to substantial loss of some PFAS (Fig 4), reducing SPE elution volume was evaluated as an alternative way to concentrate the PFAS and improve the detection limit. The effect of varying elution volume (1, 2.5, 5 mL) was investigated on the recovery efficiency of target PFAS (Fig 6). At lower elution volume (1 mL), the PFAS were incompletely desorbed, and average recovery was below the acceptable recovery limit (<30%). This is likely due to the insufficient contact of the eluent solvent with the sorbent and failing to deprotonate. Increasing the volume to 2.5 mL substantially increased the recovery above the EPA minimum limit of EIS recovery. While most target PFAS and all EIS passed the minimum recovery limit, the recovery for some of the target compounds such as PFDS, PFHpS, NMeFOSE and NFDHA were below the minimum recovery limit (<60%). Further increasing the elution volume to 5 mL provided the best recovery for all of the compounds tested. This trade-off between concentration and recovery underscored the importance of optimizing the elution volume based on the analyte characteristics and the sensitivity of the method.

Download:

Fig 6. EIS recovery comparison of 1, 2.5, and 5 mL SPE elution volume of methanolic ammonium hydroxide in Oasis WAX sorbent.

Data are shown as mean ± SD (n = 3).

https://doi.org/10.1371/journal.pwat.0000501.g006

Different PFAS bind with different affinities in a mixed-mode Oasis WAX sorbent, due to the difference in their physicochemical properties. Previous studies demonstrated the role of pH and composition of the solvent in the desorption of the analyte from the sorbents. The higher pH of about 8 was favored over the lower pH of 2 for most of the PFAS. This is due to the neutralization of weak anion exchange tertiary ammonium sites (pKa of 8) by deprotonation, disrupting the ion-ion interaction and releasing the analyte [59]. Given the size of the cartridge and amount of sorbent in this study, the optimal volume of eluent was determined to be 5 mL, which appeared to have effectively wetted the sorbent and modified the pH of the environment, and hence released the PFAS from binding to achieve the best recovery.

Effect of SPE Loading Flow Rate on PFAS Recovery

To improve sample throughput, the effect of SPE loading flow rate on PFAS recovery was first investigated using a mixture of four PFAS (PFOA, PFOS, PFBA, PFBS). To our surprise, good recoveries (87–123%) were achieved at two substantially higher sample loading rates of 40–45 and 80–85 mL/min than the original 5 mL/min (Table 2 and Table K in S1 Text). Generally, in a mixed-mode SPE, the conventional SPE protocol often recommends slower flow rates of about 1–2 mL/min to ensure sufficient time for interaction between analytes and the sorbents. The latest method for PFAS analysis by the EPA Method 1633A uses a 5 mL/min flow rate for sample loading, significantly impacting the throughput, especially when processing a larger number of samples. This is based on the understanding that the anion exchange interaction exhibits slower mass transfer kinetics compared to other interactions, requiring a longer time of contact for optimal retention of the analyte [60]. However, our empirical data challenged this assumption and suggested the feasibility of fast-flow SPE in a high-throughput environment.

Download:

Table 2. Comparison of average recovery of four legacy PFAS at different SPE flow rates.

https://doi.org/10.1371/journal.pwat.0000501.t002

Although, previous study by Rossi et. al. successfully quantified the trace level chromium (HCrO₄^-) in the anion exchange resins at higher sample loading rate such as 50 mL/min indicating sufficient ion-ion interaction even at elevated flow rate [61]. This study, to the best of our knowledge, is the first one to optimize the flow rate and enhance the throughput by ten to twenty-fold than existing method for PFAS analysis. The sample loading time for 500 mL could be drastically reduced to 6 minutes from 100 minutes recommended by the EPA Method 1633A, without compromising the recoveries.

Impact of an optimized fast flow SPE protocol on the MDL and ML

An elevated flow rate was employed to process a larger volume of sample, such as 4 L, in a considerably shorter amount of time (50 min vs. 13.3 hours that would otherwise take using the 5 mL/min flow rate). The 4 L blank water was spiked with a concentration aiming two-fold concentration of the eMDL in quadruplet samples to calculate MDL and ML. The 1000-fold preconcentration of PFAS by the fast flow SPE (Protocol 3) decreases the MDL and ML by 10 – 50-fold as compared to the existing method (Table 1 and Table L in S1 Text). The MDL ranged from 0.01 – 1.12 ppt while ML ranged from 0.04-3.6 ppt. Nearly all target PFAS (38 out of 40; except for NMeFOSE and NFDHA) and all of their corresponding EIS were within the EPA recovery limit (Fig 7A and 7B). To further improve MDL and ML, SPE extracts were concentrated by N₂ drying (Protocol 4) before UPLC-MS/MS analysis. As expected, sulfonamides (PFOSA, NMeFOSA, and NEtFOSA) and sulfonamidoethanols (NMeFOSE and NEtFOSE) in the EIS were not recovered to the minimum recovery limit after N₂ drying (Fig 7A), consistent with our previous observations (Fig 4A). Majority of target PFAS (34 out of 40; except for NMeFOSA, NEtFOSA, 3:3 FTCA, 5:3 FTCA, 7:3 FTCA, and NFDHA) were within the EPA recovery limit (Fig 7B). However, results showed minimal difference in the MDL and ML between with and without concentration by N₂ drying (Table 1 and Table L in S1 Text). Interestingly, when examining background PFAS levels in the blank water (Optima LC/MS grade), several PFAS were present at or near the MDL (Table O in S1 Text) established by our fast-flow SPE methods (Protocols 3 and 4). This suggests that background PFAS levels in the blank water could eventually establish a detection threshold, precluding further improvement in PFAS analysis below this threshold.

Download:

Fig 7. EIS recovery (A) and target PFAS recovery (B) comparison of fast flow SPE (Protocol 3) and fast flow SPE combined with N₂ drying methods (Protocol 4) with existing EPA Method 1633A (Protocol 1).

Data are shown as mean ± SD (n = 4).

https://doi.org/10.1371/journal.pwat.0000501.g007

Application of the optimized fast flow SPE protocol in drinking water samples

The newly developed method was applied to determine the forty PFAS in tap water samples using 500 mL and 4 L volume and the results were compared to the existing method. Our new method was able to detect at least 10 PFAS out of the 40 target PFAS (Table 3), which were undetectable in the existing EPA Method 1633A. The concentrations were consistent among the optimized methods for each of the PFAS and were present in the range of 0.19 to 4.10 ppt. All regulated PFAS such as PFOS, PFOA, PFHxS, and PFNA, concentrations were below the current MCL level.

Download:

Table 3. Analysis of PFAS in tap water using existing and optimized methods.

https://doi.org/10.1371/journal.pwat.0000501.t003

Conclusion

An optimized fast flow SPE-UPLC-MS/MS method was developed to improve analytical sensitivity and throughput for simultaneous detection and quantification of sub-ppt trace levels of 40 PFAS compounds in the drinking water. The impact of several key sample preparation parameters, including N₂ drying, syringe filtration, SPE elution volume, and SPE flow rate, on the PFAS recovery was determined, which was used to inform the optimization process. It was demonstrated that fast flow SPE substantially reduces the sample loading time without compromising the PFAS recovery and, hence, it constitutes a viable strategy to increase the throughput and reduce the cost for PFAS analysis. The combination of the fast flow SPE with a larger sample volume greatly shortened the sample preparation time that would otherwise take multiple hours and improved the method sensitivity to allow sub-ppt trace level quantification of PFAS in the drinking water. As a result, our investigation is expected to help reduce the cost of PFAS analysis and move PFAS monitoring in the drinking water beyond MCLs and closer to MCLGs.

Supporting information

S1 Table. Complete dataset used to generate summary statistics reported in the manuscript.

https://doi.org/10.1371/journal.pwat.0000501.s001

(XLSX)

S1 Text. List of supplementary tables and figures in S1 Text.

Table A Forty PFAS compounds from EPA Method 1633A and their CAS numbers. Table B PFAS stock concentrations of the 40 PFAS mixture. Table C: Extracted Internal Standard (EIS) and their stock concentrations in the mixture. Table D Non-Extracted Internal Standard (NIS) and their stock concentrations in the mixture. Table E Calibration standards (ppb) used for method development and validation. Table F: Concentrations of EIS and NIS spiked in the calibration standards. Table G LC gradient program for UPLC-MS/MS Method Development. Table H MRM for 40 target PFAS during LC-MS/MS analysis and corresponding EIS used for each target PFAS. Table I MRM for 24 EIS and 7 NIS during LC-MS/MS analysis and corresponding NIS used for each EIS. Table J Response ratio (RR), response factor (RF), and their RSD (<15%) demonstrating linearity of calibration curve. Table K Recoveries of PFAS during detection limit screening using fast flow SPE in 4 L aqueous samples. Table L Instrument Detection Limit (IDL) and Method Detection Limits (MDL) achieved in different method conditions for 40 target PFAS. Table M Recovery of sulphonamides, sulphonamide ethanol, and flurotelomer carboxylic acids as a result of N₂ drying at room temperature. Table N Recovery comparison of PFOSA, NMeFOSE at 200 ppb as a result of N₂ drying at 60°C. Table O Estimated background trace PFAS contamination present in the blank water used in this study. Fig A. Chromatogram traces of PFOS (a widely reported degradation byproduct of precursor PFAS) obtained after N2 drying of PFOSA (A) and NMeFOSE (B).

https://doi.org/10.1371/journal.pwat.0000501.s002

(DOCX)

References

1. Ullah S, Alsberg T, Berger U. Simultaneous determination of perfluoroalkyl phosphonates, carboxylates, and sulfonates in drinking water. J Chromatogr A. 2011;1218(37):6388–95. pmid:21791340
- View Article
- PubMed/NCBI
- Google Scholar
2. Al Amin Md, Sobhani Z, Liu Y, Dharmaraja R, Chadalavada S, Naidu R, et al. Recent advances in the analysis of per- and polyfluoroalkyl substances (PFAS)—A review. Environ Technol Innov. 2020;19: 100879.
- View Article
- Google Scholar
3. Boiteux V, Dauchy X, Rosin C, Munoz JF. National screening study on 10 perfluorinated compounds in raw and treated tap water in France. Arch Environ Contaminat Toxicol. 2012;63:1–12.
- View Article
- Google Scholar
4. Glüge J, London R, Cousins IT, DeWitt J, Goldenman G, Herzke D, et al. Information requirements under the essential-use concept: PFAS case studies. Environ Sci Technol. 2022;56(10):6232–42. pmid:34608797
- View Article
- PubMed/NCBI
- Google Scholar
5. Moggi G, Lenti D, Ingoglia D. Process for protecting stony materials, marble, tiles, and cement from atmospheric agents and pollutants. 1991. https://patents.google.com/patent/US5077097A/en
6. Ebnesajjad S. Introduction to fluoropolymers: materials, technology, and applications. William Andrew; 2020.
7. Kissa E. Fluorinated surfactants and repellents, second edition. CRC Press; 2001.
8. Drage DS, Sharkey M, Berresheim H, Coggins M, Harrad S. Rapid determination of selected PFAS in textiles entering the waste stream. Toxics. 2023;11(1):55. pmid:36668781
- View Article
- PubMed/NCBI
- Google Scholar
9. Cahuas L, Muensterman DJ, Kim-Fu ML, Reardon PN, Titaley IA, Field JA. Paints: a source of volatile PFAS in air─potential implications for inhalation exposure. Environ Sci Technol. 2022;56(23):17070–9. pmid:36367233
- View Article
- PubMed/NCBI
- Google Scholar
10. Domalanta MRB, Caldona EB. Toward enhancing the surface adhesion of fluoropolymer-based coating materials. Polym Rev. 2024;64:980–1030.
- View Article
- Google Scholar
11. Seltenrich N. PFAS in food packaging: a hot, greasy exposure. Environ Health Perspect. 2020;128(5):54002. pmid:32463326
- View Article
- PubMed/NCBI
- Google Scholar
12. Encarnacion EKP, Alcantara AC, Armario HE, Alejandro WP, Zhan Z, Sun Z. Preliminary screening of per- and polyfluoroalkyl substances (PFAS) in Philippine fast food packaging using liquid chromatography–mass spectrometry (LC-MS). Curr Res Nutr Food Sci J. 2024;12:423–36.
- View Article
- Google Scholar
13. Wolf N, Müller L, Enge S, Ungethüm T, Simat TJ. Analysis of PFAS and further VOC from fluoropolymer-coated cookware by thermal desorption-gas chromatography-mass spectrometry (TD-GC-MS). Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 2024;41(12):1663–78. pmid:39311776
- View Article
- PubMed/NCBI
- Google Scholar
14. Whitehead HD, Peaslee GF. Directly fluorinated containers as a source of perfluoroalkyl carboxylic acids. Environ Sci Technol Letters. 2023;10:350–5.
- View Article
- Google Scholar
15. Dubocq F, Wang T, Yeung LWY, Sjöberg V, Kärrman A. Characterization of the chemical contents of fluorinated and fluorine-free firefighting foams using a novel workflow combining nontarget screening and total fluorine analysis. Environ Sci Technol. 2020;54(1):245–54. pmid:31789512
- View Article
- PubMed/NCBI
- Google Scholar
16. Mazumder N-U-S, Hossain MT, Jahura FT, Girase A, Hall AS, Lu J, et al. Firefighters’ exposure to per-and polyfluoroalkyl substances (PFAS) as an occupational hazard: a review. Front Mater. 2023;10:10.3389/fmats.2023.1143411. pmid:38074949
- View Article
- PubMed/NCBI
- Google Scholar
17. Gaines LGT. Historical and current usage of per- and polyfluoroalkyl substances (PFAS): a literature review. Am J Ind Med. 2023;66(5):353–78. pmid:35614869
- View Article
- PubMed/NCBI
- Google Scholar
18. Schymanski EL, Zhang J, Thiessen PA, Chirsir P, Kondic T, Bolton EE. Per- and Polyfluoroalkyl Substances (PFAS) in PubChem: 7 million and growing. Environ Sci Technol. 2023;57:16918–28.
- View Article
- Google Scholar
19. Wong F, Hung H, Dryfhout-Clark H, Aas W, Bohlin-Nizzetto P, Breivik K, et al. Time trends of persistent organic pollutants (POPs) and Chemicals of Emerging Arctic Concern (CEAC) in Arctic air from 25 years of monitoring. Sci Total Environ. 2021;775: 145109.
- View Article
- Google Scholar
20. Brusseau ML, Anderson RH, Guo B. PFAS concentrations in soils: background levels versus contaminated sites. Sci Total Environ. 2020;740:140017. pmid:32927568
- View Article
- PubMed/NCBI
- Google Scholar
21. Maruzzo AJ, Hernandez AB, Swartz CH, Liddie JM, Schaider LA. Socioeconomic disparities in exposures to PFAS and other unregulated industrial drinking water contaminants in US public water systems. Environ Health Perspect. 2025;133(1):17002. pmid:39812474
- View Article
- PubMed/NCBI
- Google Scholar
22. Lemay AC, Bourg IC. Interactions between per- and polyfluoroalkyl substances (PFAS) at the water–air interface. Environ Sci Technol. 2025;59:2201–10.
- View Article
- Google Scholar
23. Koenig P, Brand B, Hamscher G, Stahl T. Development, optimization and validation of a highly sensitive and selective method for the determination of PFAS in animal-based food. Chemosphere. 2025;372:144123. pmid:39862654
- View Article
- PubMed/NCBI
- Google Scholar
24. Smalling KL, Romanok KM, Bradley PM, Morriss MC, Gray JL, Kanagy LK, et al. Per- and polyfluoroalkyl substances (PFAS) in United States tapwater: Comparison of underserved private-well and public-supply exposures and associated health implications. Environ Int. 2023;178:108033. pmid:37356308
- View Article
- PubMed/NCBI
- Google Scholar
25. Du X, Xu X, Yu H, Du Z, Wu Y, Qian K, et al. Thyrotoxic effects of mixed exposure to perfluorinated compounds: integrating population-based, toxicogenomic, animal, and cellular evidence to elucidate molecular mechanisms and identify potential effector targets. Environ Sci Technol. 2024;58(41):18177–89. pmid:39359169
- View Article
- PubMed/NCBI
- Google Scholar
26. Bali SK, Martin R, Almeida NMS, Saunders C, Wilson AK. Per- and Polyfluoroalkyl (PFAS) Disruption of Thyroid Hormone Synthesis. ACS Omega. 2024;9:39554–63.
- View Article
- Google Scholar
27. Vieira VM, Hoffman K, Shin H-M, Weinberg JM, Webster TF, Fletcher T. Perfluorooctanoic acid exposure and cancer outcomes in a contaminated community: a geographic analysis. Environ Health Perspect. 2013;121(3):318–23. pmid:23308854
- View Article
- PubMed/NCBI
- Google Scholar
28. Yu N, Wei S, Li M, Yang J, Li K, Jin L, et al. Effects of Perfluorooctanoic Acid on Metabolic Profiles in Brain and Liver of Mouse Revealed by a High-throughput Targeted Metabolomics Approach. Sci Rep. 2016;6:23963. pmid:27032815
- View Article
- PubMed/NCBI
- Google Scholar
29. Cao H, Zhou Z, Hu Z, Wei C, Li J, Wang L. Effect of enterohepatic circulation on the accumulation of per- and polyfluoroalkyl substances: evidence from experimental and computational studies. Environ Sci Technol. 2022;56:3214–24.
- View Article
- Google Scholar
30. Liu D, Yan S, Wang P, Chen Q, Liu Y, Cui J, et al. Perfluorooctanoic acid (PFOA) exposure in relation to the kidneys: a review of current available literature. Front Physiol. 2023;14:1103141. pmid:36776978
- View Article
- PubMed/NCBI
- Google Scholar
31. Han X, Nabb DL, Russell MH, Kennedy GL, Rickard RW. Renal elimination of perfluorocarboxylates (PFCAs). Chem Res Toxicol. 2012;25(1):35–46. pmid:21985250
- View Article
- PubMed/NCBI
- Google Scholar
32. Louisse J, Dellafiora L, van den Heuvel JJMW, Rijkers D, Leenders L, Dorne J-LCM, et al. Perfluoroalkyl substances (PFASs) are substrates of the renal human organic anion transporter 4 (OAT4). Arch Toxicol. 2023;97(3):685–96. pmid:36436016
- View Article
- PubMed/NCBI
- Google Scholar
33. Nilsson S, Thompson J, Mueller JF, Bräunig J. Apparent half-lives of chlorinated-perfluorooctane sulfonate and perfluorooctane sulfonate isomers in aviation firefighters. Environ Sci Technol. 2022;56(23):17052–60. pmid:36367310
- View Article
- PubMed/NCBI
- Google Scholar
34. Li Y, Andersson A, Xu Y, Pineda D, Nilsson CA, Lindh CH, et al. Determinants of serum half-lives for linear and branched perfluoroalkyl substances after long-term high exposure-A study in Ronneby, Sweden. Environ Int. 2022;163:107198. pmid:35447437
- View Article
- PubMed/NCBI
- Google Scholar
35. Wallis DJ, Kotlarz N, Knappe DRU, Collier DN, Lea CS, Reif D, et al. Estimation of the half-lives of recently detected per- and polyfluorinated alkyl ethers in an exposed community. Environ Sci Technol. 2023;57:15348–55.
- View Article
- Google Scholar
36. US EPA O. Per- and polyfluoroalkyl substances (PFAS). 2021. Accessed 2025 July 20. https://www.epa.gov/sdwa/and-polyfluoroalkyl-substances-pfas
37. New EU rules limit PFAS in drinking water. Environment. Accessed 2026 February 15. https://environment.ec.europa.eu/news/new-eu-rules-limit-pfas-drinking-water-2026-01-12_en
38. Treatment of drinking water to remove PFAS. Accessed 2026 February 15. https://www.eea.europa.eu/en/european-zero-pollution-dashboards/indicators/treatment-of-drinking-water-to-remove-pfas-signal
39. White & Case L. New standards for drinking water in Germany – new limits for PFAS. 2026. Accessed 2026 February 15. https://www.whitecase.com/insight-alert/new-standards-drinking-water-germany-new-limits-pfas
40. Southerland E, Birnbaum LS. What limits will the world health organization recommend for PFOA and PFOS in drinking water?. Environ Sci Technol. 2023;57:7103–5.
- View Article
- Google Scholar
41. Method 1633A analysis of per- and polyfluoroalkyl substances (PFAS) in aqueous, solid, biosolids, and tissue samples by LC-MS/MS. 2024.
42. Wu C, Wang Q, Chen H, Li M. Rapid quantitative analysis and suspect screening of per-and polyfluorinated alkyl substances (PFASs) in aqueous film-forming foams (AFFFs) and municipal wastewater samples by Nano-ESI-HRMS. Water Res. 2022;219:118542. pmid:35550967
- View Article
- PubMed/NCBI
- Google Scholar
43. Pelch KE, McKnight T, Reade A. 70 analyte PFAS test method highlights need for expanded testing of PFAS in drinking water. Sci Total Environ. 2023;876:162978. pmid:37059129
- View Article
- PubMed/NCBI
- Google Scholar
44. Bai X, Son Y. Perfluoroalkyl substances (PFAS) in surface water and sediments from two urban watersheds in Nevada, USA. Sci Total Environ. 2021;751:141622. pmid:32871315
- View Article
- PubMed/NCBI
- Google Scholar
45. Zweigle J, Simon F, Meermann B, Zwiener C. Can qualitative nontarget data be indicative of PFAS contamination? First evidence by correlation with EOF in environmental samples. Environ Sci Technol Lett. 2024;11:996–1001.
- View Article
- Google Scholar
46. Hassan MT-A, Chen X, Fnu PIJ, Osonga FJ, Sadik OA, Li M, et al. Rapid detection of per- and polyfluoroalkyl substances (PFAS) using paper spray-based mass spectrometry. J Hazard Mater. 2024;465:133366. pmid:38185081
- View Article
- PubMed/NCBI
- Google Scholar
47. Bangma J, Barry KM, Fisher CM, Genualdi S, Guillette TC, Huset CA, et al. PFAS ghosts: how to identify, evaluate, and exorcise new and existing analytical interference. Anal Bioanal Chem. 2024;416(8):1777–85. pmid:38280017
- View Article
- PubMed/NCBI
- Google Scholar
48. Willey J, Hanley A, Anderson R, Leeson A, Thomson T. Report on the multi-laboratory validation study of PFAS by isotope dilution LC-MS/MS wastewater, surface water, and groundwater.
49. EPAMDL procedure. https://www.epa.gov/sites/default/files/2016-12/documents/mdl-procedure_rev2_12-13-2016.pdf
50. Appendix B to Part 136, Title 40 -- definition and procedure for the determination of the method detection limit—revision 2. Accessed 2025 September 6. https://www.ecfr.gov/current/title-40/appendix-Appendix%20B%20to%20Part%20136
51. Lin EZ, Nason SL, Zhong A, Fortner J, Godri Pollitt KJ. Trace analysis of per- and polyfluorinated alkyl substances (PFAS) in dried blood spots - Demonstration of reproducibility and comparability to venous blood samples. Sci Total Environ. 2023;883:163530. pmid:37094673
- View Article
- PubMed/NCBI
- Google Scholar
52. Gremmel C, Frömel T, Knepper TP. HPLC-MS/MS methods for the determination of 52 perfluoroalkyl and polyfluoroalkyl substances in aqueous samples. Anal Bioanal Chem. 2017;409(6):1643–55. pmid:27928608
- View Article
- PubMed/NCBI
- Google Scholar
53. Coggan TL, Anumol T, Pyke J, Shimeta J, Clarke BO. A single analytical method for the determination of 53 legacy and emerging per- and polyfluoroalkyl substances (PFAS) in aqueous matrices. Anal Bioanal Chem. 2019;411(16):3507–20. pmid:31073731
- View Article
- PubMed/NCBI
- Google Scholar
54. Frigerio G, Cafagna S, Polledri E, Mercadante R, Fustinoni S. Development and validation of an LC-MS/MS method for the quantitation of 30 legacy and emerging per- and polyfluoroalkyl substances (PFASs) in human plasma, including HFPO-DA, DONA, and cC6O4. Anal Bioanal Chem. 2022;414(3):1259–78. pmid:34907451
- View Article
- PubMed/NCBI
- Google Scholar
55. Weiwei Z, Songsong C, Yongzhi W, Ru Z, Chengcheng B, Jinpeng Y, et al. The soil-air interfacial migration process of volatile PFAS at the contaminated sites: Evidence from stable carbon isotopes with CSIA. Environ Pollut. 2024;363(Pt 1):125111. pmid:39419467
- View Article
- PubMed/NCBI
- Google Scholar
56. Martin JW, Muir DCG, Moody CA, Ellis DA, Kwan WC, Solomon KR, et al. Collection of airborne fluorinated organics and analysis by gas chromatography/chemical ionization mass spectrometry. Anal Chem. 2002;74(3):584–90. pmid:11842814
- View Article
- PubMed/NCBI
- Google Scholar
57. Noda MK, Lee H, Han C, Niu XZ. Syringe filters and autosampler vials impact on analysis of long-, short-, and ultrashort-chain per- and polyfluoroalkyl substances. ACS EST Water. 2025;5:1344–52.
- View Article
- Google Scholar
58. Kochan J, Wintgens T, Hochstrat R, Melin T. Impact of wetting agents on the filtration performance of polymeric ultrafiltration membranes. Desalination. 2009;241(1–3):34–42.
- View Article
- Google Scholar
59. Brumovský M, Bečanová J, Karásková P, Nizzetto L. Retention performance of three widely used SPE sorbents for the extraction of perfluoroalkyl substances from seawater. Chemosphere. 2018;193:259–69. pmid:29145087
- View Article
- PubMed/NCBI
- Google Scholar
60. Watson DW. How it works: Ion-exchange SPE. 2017;36:66–66.
61. Rossi E, Errea MI, de Cortalezzi MMF, Stripeikis J. Selective determination of Cr (VI) by on-line solid phase extraction FI-SPE-FAAS using an ion exchanger resin as sorbent: An improvement treatment of the analytical signal. Microchem J. 2017;130:88–92.
- View Article
- Google Scholar

[ref1] 1. Ullah S, Alsberg T, Berger U. Simultaneous determination of perfluoroalkyl phosphonates, carboxylates, and sulfonates in drinking water. J Chromatogr A. 2011;1218(37):6388–95. pmid:21791340
View Article
PubMed/NCBI
Google Scholar

[2] View Article

[3] PubMed/NCBI

[4] Google Scholar

[ref2] 2. Al Amin Md, Sobhani Z, Liu Y, Dharmaraja R, Chadalavada S, Naidu R, et al. Recent advances in the analysis of per- and polyfluoroalkyl substances (PFAS)—A review. Environ Technol Innov. 2020;19: 100879.
View Article
Google Scholar

[6] View Article

[7] Google Scholar

[ref3] 3. Boiteux V, Dauchy X, Rosin C, Munoz JF. National screening study on 10 perfluorinated compounds in raw and treated tap water in France. Arch Environ Contaminat Toxicol. 2012;63:1–12.
View Article
Google Scholar

[9] View Article

[10] Google Scholar

[ref4] 4. Glüge J, London R, Cousins IT, DeWitt J, Goldenman G, Herzke D, et al. Information requirements under the essential-use concept: PFAS case studies. Environ Sci Technol. 2022;56(10):6232–42. pmid:34608797
View Article
PubMed/NCBI
Google Scholar

[12] View Article

[13] PubMed/NCBI

[14] Google Scholar

[ref5] 5. Moggi G, Lenti D, Ingoglia D. Process for protecting stony materials, marble, tiles, and cement from atmospheric agents and pollutants. 1991. https://patents.google.com/patent/US5077097A/en

[ref6] 6. Ebnesajjad S. Introduction to fluoropolymers: materials, technology, and applications. William Andrew; 2020.

[ref7] 7. Kissa E. Fluorinated surfactants and repellents, second edition. CRC Press; 2001.

[ref8] 8. Drage DS, Sharkey M, Berresheim H, Coggins M, Harrad S. Rapid determination of selected PFAS in textiles entering the waste stream. Toxics. 2023;11(1):55. pmid:36668781
View Article
PubMed/NCBI
Google Scholar

[19] View Article

[20] PubMed/NCBI

[21] Google Scholar

[ref9] 9. Cahuas L, Muensterman DJ, Kim-Fu ML, Reardon PN, Titaley IA, Field JA. Paints: a source of volatile PFAS in air─potential implications for inhalation exposure. Environ Sci Technol. 2022;56(23):17070–9. pmid:36367233
View Article
PubMed/NCBI
Google Scholar

[23] View Article

[24] PubMed/NCBI

[25] Google Scholar

[ref10] 10. Domalanta MRB, Caldona EB. Toward enhancing the surface adhesion of fluoropolymer-based coating materials. Polym Rev. 2024;64:980–1030.
View Article
Google Scholar

[27] View Article

[28] Google Scholar

[ref11] 11. Seltenrich N. PFAS in food packaging: a hot, greasy exposure. Environ Health Perspect. 2020;128(5):54002. pmid:32463326
View Article
PubMed/NCBI
Google Scholar

[30] View Article

[31] PubMed/NCBI

[32] Google Scholar

[ref12] 12. Encarnacion EKP, Alcantara AC, Armario HE, Alejandro WP, Zhan Z, Sun Z. Preliminary screening of per- and polyfluoroalkyl substances (PFAS) in Philippine fast food packaging using liquid chromatography–mass spectrometry (LC-MS). Curr Res Nutr Food Sci J. 2024;12:423–36.
View Article
Google Scholar

[34] View Article

[35] Google Scholar

[ref13] 13. Wolf N, Müller L, Enge S, Ungethüm T, Simat TJ. Analysis of PFAS and further VOC from fluoropolymer-coated cookware by thermal desorption-gas chromatography-mass spectrometry (TD-GC-MS). Food Addit Contam Part A Chem Anal Control Expo Risk Assess. 2024;41(12):1663–78. pmid:39311776
View Article
PubMed/NCBI
Google Scholar

[37] View Article

[38] PubMed/NCBI

[39] Google Scholar

[ref14] 14. Whitehead HD, Peaslee GF. Directly fluorinated containers as a source of perfluoroalkyl carboxylic acids. Environ Sci Technol Letters. 2023;10:350–5.
View Article
Google Scholar

[41] View Article

[42] Google Scholar

[ref15] 15. Dubocq F, Wang T, Yeung LWY, Sjöberg V, Kärrman A. Characterization of the chemical contents of fluorinated and fluorine-free firefighting foams using a novel workflow combining nontarget screening and total fluorine analysis. Environ Sci Technol. 2020;54(1):245–54. pmid:31789512
View Article
PubMed/NCBI
Google Scholar

[44] View Article

[45] PubMed/NCBI

[46] Google Scholar

[ref16] 16. Mazumder N-U-S, Hossain MT, Jahura FT, Girase A, Hall AS, Lu J, et al. Firefighters’ exposure to per-and polyfluoroalkyl substances (PFAS) as an occupational hazard: a review. Front Mater. 2023;10:10.3389/fmats.2023.1143411. pmid:38074949
View Article
PubMed/NCBI
Google Scholar

[48] View Article

[49] PubMed/NCBI

[50] Google Scholar

[ref17] 17. Gaines LGT. Historical and current usage of per- and polyfluoroalkyl substances (PFAS): a literature review. Am J Ind Med. 2023;66(5):353–78. pmid:35614869
View Article
PubMed/NCBI
Google Scholar

[52] View Article

[53] PubMed/NCBI

[54] Google Scholar

[ref18] 18. Schymanski EL, Zhang J, Thiessen PA, Chirsir P, Kondic T, Bolton EE. Per- and Polyfluoroalkyl Substances (PFAS) in PubChem: 7 million and growing. Environ Sci Technol. 2023;57:16918–28.
View Article
Google Scholar

[56] View Article

[57] Google Scholar

[ref19] 19. Wong F, Hung H, Dryfhout-Clark H, Aas W, Bohlin-Nizzetto P, Breivik K, et al. Time trends of persistent organic pollutants (POPs) and Chemicals of Emerging Arctic Concern (CEAC) in Arctic air from 25 years of monitoring. Sci Total Environ. 2021;775: 145109.
View Article
Google Scholar

[59] View Article

[60] Google Scholar

[ref20] 20. Brusseau ML, Anderson RH, Guo B. PFAS concentrations in soils: background levels versus contaminated sites. Sci Total Environ. 2020;740:140017. pmid:32927568
View Article
PubMed/NCBI
Google Scholar

[62] View Article

[63] PubMed/NCBI

[64] Google Scholar

[ref21] 21. Maruzzo AJ, Hernandez AB, Swartz CH, Liddie JM, Schaider LA. Socioeconomic disparities in exposures to PFAS and other unregulated industrial drinking water contaminants in US public water systems. Environ Health Perspect. 2025;133(1):17002. pmid:39812474
View Article
PubMed/NCBI
Google Scholar

[66] View Article

[67] PubMed/NCBI

[68] Google Scholar

[ref22] 22. Lemay AC, Bourg IC. Interactions between per- and polyfluoroalkyl substances (PFAS) at the water–air interface. Environ Sci Technol. 2025;59:2201–10.
View Article
Google Scholar

[70] View Article

[71] Google Scholar

[ref23] 23. Koenig P, Brand B, Hamscher G, Stahl T. Development, optimization and validation of a highly sensitive and selective method for the determination of PFAS in animal-based food. Chemosphere. 2025;372:144123. pmid:39862654
View Article
PubMed/NCBI
Google Scholar

[73] View Article

[74] PubMed/NCBI

[75] Google Scholar

[ref24] 24. Smalling KL, Romanok KM, Bradley PM, Morriss MC, Gray JL, Kanagy LK, et al. Per- and polyfluoroalkyl substances (PFAS) in United States tapwater: Comparison of underserved private-well and public-supply exposures and associated health implications. Environ Int. 2023;178:108033. pmid:37356308
View Article
PubMed/NCBI
Google Scholar

[77] View Article

[78] PubMed/NCBI

[79] Google Scholar

[ref25] 25. Du X, Xu X, Yu H, Du Z, Wu Y, Qian K, et al. Thyrotoxic effects of mixed exposure to perfluorinated compounds: integrating population-based, toxicogenomic, animal, and cellular evidence to elucidate molecular mechanisms and identify potential effector targets. Environ Sci Technol. 2024;58(41):18177–89. pmid:39359169
View Article
PubMed/NCBI
Google Scholar

[81] View Article

[82] PubMed/NCBI

[83] Google Scholar

[ref26] 26. Bali SK, Martin R, Almeida NMS, Saunders C, Wilson AK. Per- and Polyfluoroalkyl (PFAS) Disruption of Thyroid Hormone Synthesis. ACS Omega. 2024;9:39554–63.
View Article
Google Scholar

[85] View Article

[86] Google Scholar

[ref27] 27. Vieira VM, Hoffman K, Shin H-M, Weinberg JM, Webster TF, Fletcher T. Perfluorooctanoic acid exposure and cancer outcomes in a contaminated community: a geographic analysis. Environ Health Perspect. 2013;121(3):318–23. pmid:23308854
View Article
PubMed/NCBI
Google Scholar

[88] View Article

[89] PubMed/NCBI

[90] Google Scholar

[ref28] 28. Yu N, Wei S, Li M, Yang J, Li K, Jin L, et al. Effects of Perfluorooctanoic Acid on Metabolic Profiles in Brain and Liver of Mouse Revealed by a High-throughput Targeted Metabolomics Approach. Sci Rep. 2016;6:23963. pmid:27032815
View Article
PubMed/NCBI
Google Scholar

[92] View Article

[93] PubMed/NCBI

[94] Google Scholar

[ref29] 29. Cao H, Zhou Z, Hu Z, Wei C, Li J, Wang L. Effect of enterohepatic circulation on the accumulation of per- and polyfluoroalkyl substances: evidence from experimental and computational studies. Environ Sci Technol. 2022;56:3214–24.
View Article
Google Scholar

[96] View Article

[97] Google Scholar

[ref30] 30. Liu D, Yan S, Wang P, Chen Q, Liu Y, Cui J, et al. Perfluorooctanoic acid (PFOA) exposure in relation to the kidneys: a review of current available literature. Front Physiol. 2023;14:1103141. pmid:36776978
View Article
PubMed/NCBI
Google Scholar

[99] View Article

[100] PubMed/NCBI

[101] Google Scholar

[ref31] 31. Han X, Nabb DL, Russell MH, Kennedy GL, Rickard RW. Renal elimination of perfluorocarboxylates (PFCAs). Chem Res Toxicol. 2012;25(1):35–46. pmid:21985250
View Article
PubMed/NCBI
Google Scholar

[103] View Article

[104] PubMed/NCBI

[105] Google Scholar

[ref32] 32. Louisse J, Dellafiora L, van den Heuvel JJMW, Rijkers D, Leenders L, Dorne J-LCM, et al. Perfluoroalkyl substances (PFASs) are substrates of the renal human organic anion transporter 4 (OAT4). Arch Toxicol. 2023;97(3):685–96. pmid:36436016
View Article
PubMed/NCBI
Google Scholar

[107] View Article

[108] PubMed/NCBI

[109] Google Scholar

[ref33] 33. Nilsson S, Thompson J, Mueller JF, Bräunig J. Apparent half-lives of chlorinated-perfluorooctane sulfonate and perfluorooctane sulfonate isomers in aviation firefighters. Environ Sci Technol. 2022;56(23):17052–60. pmid:36367310
View Article
PubMed/NCBI
Google Scholar

[111] View Article

[112] PubMed/NCBI

[113] Google Scholar

[ref34] 34. Li Y, Andersson A, Xu Y, Pineda D, Nilsson CA, Lindh CH, et al. Determinants of serum half-lives for linear and branched perfluoroalkyl substances after long-term high exposure-A study in Ronneby, Sweden. Environ Int. 2022;163:107198. pmid:35447437
View Article
PubMed/NCBI
Google Scholar

[115] View Article

[116] PubMed/NCBI

[117] Google Scholar

[ref35] 35. Wallis DJ, Kotlarz N, Knappe DRU, Collier DN, Lea CS, Reif D, et al. Estimation of the half-lives of recently detected per- and polyfluorinated alkyl ethers in an exposed community. Environ Sci Technol. 2023;57:15348–55.
View Article
Google Scholar

[119] View Article

[120] Google Scholar

[ref36] 36. US EPA O. Per- and polyfluoroalkyl substances (PFAS). 2021. Accessed 2025 July 20. https://www.epa.gov/sdwa/and-polyfluoroalkyl-substances-pfas

[ref37] 37. New EU rules limit PFAS in drinking water. Environment. Accessed 2026 February 15. https://environment.ec.europa.eu/news/new-eu-rules-limit-pfas-drinking-water-2026-01-12_en

[ref38] 38. Treatment of drinking water to remove PFAS. Accessed 2026 February 15. https://www.eea.europa.eu/en/european-zero-pollution-dashboards/indicators/treatment-of-drinking-water-to-remove-pfas-signal

[ref39] 39. White & Case L. New standards for drinking water in Germany – new limits for PFAS. 2026. Accessed 2026 February 15. https://www.whitecase.com/insight-alert/new-standards-drinking-water-germany-new-limits-pfas

[ref40] 40. Southerland E, Birnbaum LS. What limits will the world health organization recommend for PFOA and PFOS in drinking water?. Environ Sci Technol. 2023;57:7103–5.
View Article
Google Scholar

[126] View Article

[127] Google Scholar

[ref41] 41. Method 1633A analysis of per- and polyfluoroalkyl substances (PFAS) in aqueous, solid, biosolids, and tissue samples by LC-MS/MS. 2024.

[ref42] 42. Wu C, Wang Q, Chen H, Li M. Rapid quantitative analysis and suspect screening of per-and polyfluorinated alkyl substances (PFASs) in aqueous film-forming foams (AFFFs) and municipal wastewater samples by Nano-ESI-HRMS. Water Res. 2022;219:118542. pmid:35550967
View Article
PubMed/NCBI
Google Scholar

[130] View Article

[131] PubMed/NCBI

[132] Google Scholar

[ref43] 43. Pelch KE, McKnight T, Reade A. 70 analyte PFAS test method highlights need for expanded testing of PFAS in drinking water. Sci Total Environ. 2023;876:162978. pmid:37059129
View Article
PubMed/NCBI
Google Scholar

[134] View Article

[135] PubMed/NCBI

[136] Google Scholar

[ref44] 44. Bai X, Son Y. Perfluoroalkyl substances (PFAS) in surface water and sediments from two urban watersheds in Nevada, USA. Sci Total Environ. 2021;751:141622. pmid:32871315
View Article
PubMed/NCBI
Google Scholar

[138] View Article

[139] PubMed/NCBI

[140] Google Scholar

[ref45] 45. Zweigle J, Simon F, Meermann B, Zwiener C. Can qualitative nontarget data be indicative of PFAS contamination? First evidence by correlation with EOF in environmental samples. Environ Sci Technol Lett. 2024;11:996–1001.
View Article
Google Scholar

[142] View Article

[143] Google Scholar

[ref46] 46. Hassan MT-A, Chen X, Fnu PIJ, Osonga FJ, Sadik OA, Li M, et al. Rapid detection of per- and polyfluoroalkyl substances (PFAS) using paper spray-based mass spectrometry. J Hazard Mater. 2024;465:133366. pmid:38185081
View Article
PubMed/NCBI
Google Scholar

[145] View Article

[146] PubMed/NCBI

[147] Google Scholar

[ref47] 47. Bangma J, Barry KM, Fisher CM, Genualdi S, Guillette TC, Huset CA, et al. PFAS ghosts: how to identify, evaluate, and exorcise new and existing analytical interference. Anal Bioanal Chem. 2024;416(8):1777–85. pmid:38280017
View Article
PubMed/NCBI
Google Scholar

[149] View Article

[150] PubMed/NCBI

[151] Google Scholar

[ref48] 48. Willey J, Hanley A, Anderson R, Leeson A, Thomson T. Report on the multi-laboratory validation study of PFAS by isotope dilution LC-MS/MS wastewater, surface water, and groundwater.

[ref49] 49. EPAMDL procedure. https://www.epa.gov/sites/default/files/2016-12/documents/mdl-procedure_rev2_12-13-2016.pdf

[ref50] 50. Appendix B to Part 136, Title 40 -- definition and procedure for the determination of the method detection limit—revision 2. Accessed 2025 September 6. https://www.ecfr.gov/current/title-40/appendix-Appendix%20B%20to%20Part%20136

[ref51] 51. Lin EZ, Nason SL, Zhong A, Fortner J, Godri Pollitt KJ. Trace analysis of per- and polyfluorinated alkyl substances (PFAS) in dried blood spots - Demonstration of reproducibility and comparability to venous blood samples. Sci Total Environ. 2023;883:163530. pmid:37094673
View Article
PubMed/NCBI
Google Scholar

[156] View Article

[157] PubMed/NCBI

[158] Google Scholar

[ref52] 52. Gremmel C, Frömel T, Knepper TP. HPLC-MS/MS methods for the determination of 52 perfluoroalkyl and polyfluoroalkyl substances in aqueous samples. Anal Bioanal Chem. 2017;409(6):1643–55. pmid:27928608
View Article
PubMed/NCBI
Google Scholar

[160] View Article

[161] PubMed/NCBI

[162] Google Scholar

[ref53] 53. Coggan TL, Anumol T, Pyke J, Shimeta J, Clarke BO. A single analytical method for the determination of 53 legacy and emerging per- and polyfluoroalkyl substances (PFAS) in aqueous matrices. Anal Bioanal Chem. 2019;411(16):3507–20. pmid:31073731
View Article
PubMed/NCBI
Google Scholar

[164] View Article

[165] PubMed/NCBI

[166] Google Scholar

[ref54] 54. Frigerio G, Cafagna S, Polledri E, Mercadante R, Fustinoni S. Development and validation of an LC-MS/MS method for the quantitation of 30 legacy and emerging per- and polyfluoroalkyl substances (PFASs) in human plasma, including HFPO-DA, DONA, and cC6O4. Anal Bioanal Chem. 2022;414(3):1259–78. pmid:34907451
View Article
PubMed/NCBI
Google Scholar

[168] View Article

[169] PubMed/NCBI

[170] Google Scholar

[ref55] 55. Weiwei Z, Songsong C, Yongzhi W, Ru Z, Chengcheng B, Jinpeng Y, et al. The soil-air interfacial migration process of volatile PFAS at the contaminated sites: Evidence from stable carbon isotopes with CSIA. Environ Pollut. 2024;363(Pt 1):125111. pmid:39419467
View Article
PubMed/NCBI
Google Scholar

[172] View Article

[173] PubMed/NCBI

[174] Google Scholar

[ref56] 56. Martin JW, Muir DCG, Moody CA, Ellis DA, Kwan WC, Solomon KR, et al. Collection of airborne fluorinated organics and analysis by gas chromatography/chemical ionization mass spectrometry. Anal Chem. 2002;74(3):584–90. pmid:11842814
View Article
PubMed/NCBI
Google Scholar

[176] View Article

[177] PubMed/NCBI

[178] Google Scholar

[ref57] 57. Noda MK, Lee H, Han C, Niu XZ. Syringe filters and autosampler vials impact on analysis of long-, short-, and ultrashort-chain per- and polyfluoroalkyl substances. ACS EST Water. 2025;5:1344–52.
View Article
Google Scholar

[180] View Article

[181] Google Scholar

[ref58] 58. Kochan J, Wintgens T, Hochstrat R, Melin T. Impact of wetting agents on the filtration performance of polymeric ultrafiltration membranes. Desalination. 2009;241(1–3):34–42.
View Article
Google Scholar

[183] View Article

[184] Google Scholar

[ref59] 59. Brumovský M, Bečanová J, Karásková P, Nizzetto L. Retention performance of three widely used SPE sorbents for the extraction of perfluoroalkyl substances from seawater. Chemosphere. 2018;193:259–69. pmid:29145087
View Article
PubMed/NCBI
Google Scholar

[186] View Article

[187] PubMed/NCBI

[188] Google Scholar

[ref60] 60. Watson DW. How it works: Ion-exchange SPE. 2017;36:66–66.

[ref61] 61. Rossi E, Errea MI, de Cortalezzi MMF, Stripeikis J. Selective determination of Cr (VI) by on-line solid phase extraction FI-SPE-FAAS using an ion exchanger resin as sorbent: An improvement treatment of the analytical signal. Microchem J. 2017;130:88–92.
View Article
Google Scholar

[191] View Article

[192] Google Scholar

Figures

Abstract

Introduction

Experimental Section

Reagents, chemicals, and standard solutions

UPLC-MS/MS method for PFAS detection

Sample preparation and extraction

Instrument detection limits, method detection limits, and method limits

Method optimizations to improve MDL, ML, and throughput

Concentration by N2 drying.

Syringe filtration of SPE extract.

Elution volume optimizations.

Rapid sample preparation with fast flow SPE.

Improving MDL and ML with fast flow SPE and larger sample volume.

Data and statistical analysis

Results and discussion

UPLC-MS/MS detection of forty target PFAS compounds

Calibration linearity and instrument detection limit

Method detection limit, method limit, and recoveries before optimization

Effect of N2 drying on MDL, ML and recovery

Effect of syringe filtration on PFAS recovery

Effect of SPE elution volume on PFAS recovery

Effect of SPE Loading Flow Rate on PFAS Recovery

Impact of an optimized fast flow SPE protocol on the MDL and ML

Application of the optimized fast flow SPE protocol in drinking water samples

Conclusion

Supporting information

S1 Table. Complete dataset used to generate summary statistics reported in the manuscript.

S1 Text. List of supplementary tables and figures in S1 Text.

References

Concentration by N₂ drying.

Effect of N₂ drying on MDL, ML and recovery