Open AccessReview

Review of Emerging and Nonconventional Analytical Techniques for Per- and Polyfluoroalkyl Substances (PFAS): Application for Risk Assessment

Andrew McQueen

^1,*

Ashley Kimble

¹,

Paige Krupa

¹,

Anna Longwell

²,

Alyssa Calomeni-Eck

and

David Moore

US Army Corps of Engineers, Engineer Research and Development Center, 3909 Halls Ferry Road, Vicksburg, MS 39056, USA

US Army Corps of Engineers, Buffalo District, 478 Main Street, Buffalo, NY 14202, USA

Author to whom correspondence should be addressed.

Water 2025, 17(3), 303; https://doi.org/10.3390/w17030303

Submission received: 7 December 2024 / Revised: 10 January 2025 / Accepted: 15 January 2025 / Published: 22 January 2025

(This article belongs to the Section Water Quality and Contamination)

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Abstract

Per- and polyfluoroalkyl substances (PFAS) are persistent environmental contaminants that pose significant risks to ecosystems and human health. Increasing regulatory demands for PFAS management have increased the need for rapid and deployable analytical technologies for both abiotic and biotic matrices. Traditional detection methods, such as standardized chromatography, often require weeks to months for analysis due to a limited number of appropriately accredited laboratories, delaying critical decision-making. This literature review is intended to identify promising emerging PFAS analytical techniques or technologies to facilitate more rapid (near real-time) analysis and explore their relevancy in supporting human and ecological risk assessments. Recently developed optical and electrochemical sensing approaches are enabling the detection of PFASs within minutes to hours, with detection limits typically aligning within reported ambient concentrations in water, soil, and sediment. These emerging technologies could (1) support planning and prioritization of sampling efforts during the problem formulation phase of risk assessment, (2) complement traditional chromatography methods to lower time and resource demands to improve sampling frequency over space and time, and (3) aid in risk-informed characterization of PFAS exposures based on identified chemical classes or groups. This review highlights those approaches and technologies that could potentially enhance the comprehensiveness and efficiency of PFAS risk assessment across diverse environmental settings in the future.

Keywords:

PFOA; PFOS; optical sensors; electrochemical sensors; ecological risk assessment

Graphical Abstract

1. Introduction

Per- and polyfluoroalkyl substances (PFAS) are a complex class of persistent environmental contaminants that pose potential threats to ecosystems and human health. This class of synthetic chemicals has a stable carbon–fluoride (C-F) bond and functional groups that have been used in a variety of applications since their development in the mid-20th century due to their chemical stability, oil- and water-repellant properties, and fire suppression of flammable liquids as an aqueous film-forming foam (AFFF) [1]. Consequently, these substances have become globally distributed and pervasive in the environment [2], found in surface water and groundwater [3], rainwater [4], sediment and soil [5], animals [6], and humans [7]. As awareness of their environmental persistence and potential human health and ecological concerns has grown [8,9,10], the United States Environmental Protection Agency (USEPA) and international environmental agencies are introducing regulations to manage PFASs. Consequently, there is growing demand for tools and approaches to detect and quantify this complex suite of contaminants in the environment [11].

In the United States, the USEPA and individual states are starting to develop regulations and advisories for PFASs, increasing the need for rapid, deployable technologies to provide near-real-time analysis of PFASs in the environment. In April 2024, the USEPA released the final National Primary Drinking Water Regulation for six PFASs: perfluorooctanoic acid (PFOA), perfluorooctane sulfonate (PFOS), perfluorohexane sulfonate (PFHxS), perfluorononanoic acid (PFNA), hexafluoropropylene oxide dimer acid (HFPO-DA; or “GenX” chemicals), and perfluorobutanesulfonic acid (PFBS) (USEPA 2024). This regulation includes legally enforceable maximum contaminant levels (MCLs) for PFOA (4.0 ng/L or parts per trillion (ppt)), PFOS (4.0 ppt), PFHxS (10 ppt), PFNA (10 ppt), and HFPO-DA (10 ppt) individually, as well as mixtures with at least two of PFHxS, PFNA, HFPO-DA, and PFBS based off a combined hazard index value (hazard index MCL = 1; Equation (1)):

H a z a r d I n d e x = (\frac{H F P O - D A}{10}) + (\frac{P F B S}{2000}) + (\frac{P F N A}{10}) + (\frac{P F H x S}{10})

(1)

The USEPA has also developed soil regional screening levels (RSLs) for various PFAS analytes. Though these are not legally enforceable limits or clean-up standards, they are used by risk assessors to screen sites, including to help determine if a remedial investigation is needed [12]. Some of these RSL values are very low, particularly for PFOS and PFOA (e.g., residential soil RSLs of 0.63 µg/kg for PFOS and 0.019 µg/kg for PFOA [12]), highlighting the importance of analytical detection limits. Prior to the release of the USEPA MCLs, several states developed their own regulatory levels for PFASs in drinking water, though these values tend to be less conservative than the current 2024 USEPA MCLs [13]. However, some states have proposed more stringent water standards below ppt ranges (i.e., North Carolina proposed groundwater standards of 0.7 ng PFOS/L and 0.001 ng PFOA/L [14]). Various states have also released soil PFAS screening and advisory levels, with a few, including Alaska, Connecticut, Massachusetts, New Jersey, Pennsylvania, and Texas, having soil clean-up thresholds that typically require action when exceeded [15].

In March 2024, the Department of Defense completed preliminary assessments/site inspections at over 700 installations and is now prioritizing actions under the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) process [16]. Risk assessment is an iterative process in which initial screening data are collected, informing subsequent more detailed and resource intensive assessments of a site [17,18]. As an example, hundreds of samples of different abiotic (e.g., water, soil, sediment) and biotic (e.g., fish tissue) exposure routes are often collected during the problem formulation phase to characterize potential sources and concentration gradients from the source to identify areas potentially requiring remediation. For CERCLA, sampling and constituent analysis is needed during problem formulation, remediation, and post-remedy monitoring. Because this is a data-rich process that requires information sharing among stakeholders, federal agencies are interested in rapid, field-ready analytical techniques that could streamline decision making.

Initially, the combination of low concentrations, lack of standards, and unusual physical and chemical properties (e.g., non-volatility and thermal stability) made it difficult to confirm PFAS identity using traditional chromatography techniques. Additionally, the ubiquitous nature of PFASs and presence in many laboratory supplies added to the complication of quantitation. It was not until 2009 that the EPA published its first standardized analytical method for drinking water, EPA Method 537. Since then, EPA Method 537 has been updated to 537.1 and is the required method for analysis of PFASs in drinking water using a solid-phase extraction clean-up and concentration step and liquid chromatography-tandem mass spectrometry (LC-MS/MS) [19]. This method quantitively measures 24 PFASs with detection ranges from 0.53 to 2.8 ppt using 250 mL of water and concentrating to 5 mL. USEPA Method 533 also measures PFASs in drinking water but uses an isotope dilution approach with LC-MS/MS [20]. It targets short-chain PFAS compounds that are not covered by Method 537.1, whereby PFASs with fewer than six fully fluorinated carbon atoms in their perfluoroalkyl chain for perfluoroalkyl sulfonic acids (e.g., PFBS) or fewer than eight for perfluoroalkyl carboxylic acids (e.g., PFBA) are considered short-chain [8].

The most recent EPA method was published in 2021, EPA Method 1633 [21], and is now on the 4th version. This method is specifically designed for matrices other than drinking water (e.g., non-potable, soils, sediment, and tissues). In recent years, this has become the mandated methodology for analysis of PFASs in media other than drinking water. A fundamental need in risk assessments is exposure route and toxicological data that are supported by high-quality analytical verification. However, PFAS exposure verification has been particularly challenging due to the lack of qualified and certified laboratories and lack of widely accepted methods for PFAS species and matrices [22]. Additionally, cost and turnaround time are significant factors to stakeholders, decision makers, and researchers as information is being gathered regarding PFASs at contaminated sites. Therefore, there is a demand for “lower tier” (i.e., screening) PFAS analytical methods that can augment the existing standardized chromatography methods (e.g., LC-MS/MS).

While there are several review papers focused on recent progress of emerging PFAS analytical sensing technologies (e.g., [11,23,24]), to date, there has been a limited focus on the context or practical implications of these future tools for environmental monitoring supporting regulatory compliance or risk assessment and risk management processes. Therefore, a primary aim of this study is to conduct a literature review of PFAS analytical techniques focused on promising emerging technologies that may facilitate more rapid (near-real-time) analysis to better inform human and/or ecological risk assessments. This review is not intended to cover all emerging techniques but to provide illustrative examples of selected promising technologies currently being explored that may support improvements to environmental monitoring applications. Approaches will be compared based on reported reliability, practicality, sensitivity, selectivity, and other considerations to provide context as to which scenarios different technologies may be effective in supporting human and ecological risk assessments.

2. Methods

This review expanded upon prior recent literature reviews (i.e., [11,23,24]) using select technologies identified therein and provided practical context for use in environmental or human health risk assessment settings. Therefore, the technology categories identified in these prior reviews (e.g., optical-, paper-, porphyrin-, fluorescent-based methods) were retained in the current review and a few (e.g., 1 to 3) illustrative examples of each emerging technique were identified for each category. A focused literature review was conducted of Scopus and Google Scholar for articles, using permutations of the following search terms: “PFAS”, “PFOA”, “PFOS”, “perfluorinated” and “polyfluoroalkyl substances” with the terms “standardized methodologies”, “analytical”, “method validation”, and the terms “rapid”, “portable”, “qualitative”, “screening”, “novel”, “new”. An example search string was “PFAS AND standardized methodologies AND novel”. Titles and abstracts were reviewed to determine if the articles identified in this review captured emerging analytical techniques. For example, manuscripts often included methods that were modifications of the existing LC-MS/MS standard methods and these were excluded. Further, this review focused on experimental studies that provide quantitative or qualitative data on the evaluation of PFAS detection using synthetic media or real-world samples (e.g., river water). The accepted citations were limited to non-standard approaches; however, data relevant to field portable versions of standard analytical tools were also reviewed to provide additional insight for potential field deployable applications. While the search included articles and reports published from 1990 to 2024, the majority of the identified articles that fit the selection criteria were published between 2020 and 2024.

3. Results

Emerging technologies developed for PFAS detection generally use optical, electrochemical, and other nonconventional methods (i.e., hydrogen swelling, thermal detection, flow rate analysis) (Table 1). Optical methods, such as fluorescence spectroscopy and surface-enhanced Raman spectroscopy (SERS), rely on light-based interactions with PFAS molecules, enabling rapid and sensitive detection at trace levels. Electrochemical-based approaches, including differential pulse voltammetry (DPV) and electrochemical impedance spectroscopy (EIS), utilize modified electrodes and selective recognition elements like molecularly imprinted polymers (MIPs) to detect PFASs through measurable changes in electrical signals. Nonconventional methods, such as those employing hydrogen swelling techniques, flow rates, or thermal detection offer alternative approaches that may provide low-cost and qualitative information. Many of these sensing technologies are aimed at improving real-time detection/quantification of PFASs using unique approaches, materials, or processes with the goal of achieving field deployable or in situ capabilities. Foundational processes and approaches are discussed below with some relevant examples for each sensing category.

3.1. Optical-Based Methods

Optical-based methods for detecting PFASs involve using optical sensors or techniques to correlate changes in optical properties with analyte concentrations. Optical sensing techniques have been successfully commercially scaled and applied for sensing a variety of persistent organic pollutants (POPs), often achieving requisite detection limits using low-cost and timely analytical methods [56]. Advancements in optical methods for PFAS measurement rely on the interaction of target analyte molecules with tailored substances (e.g., nanomaterials, catalytic platforms). In some cases, adaptation of surface materials (e.g., nanomaterials) is used to enhance optical signals and quantify target analyte through either chemical reaction or physical mechanism. Some pertinent examples of optical sensing platforms are reviewed below (Table 2).

3.1.1. Colorimetric Porphyrin-Based Detector

Taylor et al. [26] developed a porphyrin-based sensor that undergoes a color change when exposed to PFOA. Porphyrins are molecules that are often used as sensor molecules due to their ability to be functionalized (modifying the surface of molecules by addition of functional groups) to act as selective receptors that can bind specific target molecules. In this study, the α,α,α,α porphyrin isomer, also called the “picket fence” isomer, was functionalized with pentadecafluorooctanoyl chloride so that the molecule could bind fluorocarbons including PFOA. When such binding occurs, the absorbance properties of the porphyrin isomer are changed, resulting in an instant color change visible to the human eye. Selectivity testing suggested this sensor was specific to molecules with both a carboxylate anion and a fluorinated alkyl chain. The color change occurred with the addition of PFOA salt (tetrabutylammonium perfluorooctanoate), but not with a neutral fluorinated compound, non-fluorinated carboxylates, and common analytes (including F⁻, Cl⁻, Br⁻, I⁻, HSO₄⁻, and OH⁻).

Benefits of this technology include ease of use and minimal sample preparation. Taylor et al. [26] described simply adding the PFAS water sample to the porphyrin-dichloromethane solution, resulting in an instantaneous color change. Although spectroscopy would be needed to quantify PFASs, a qualitative measurement (color change in the presence of PFAS > 3 ppm) using this sensor would not require specialized equipment and could be carried out in the field. This sensor also can be used for various matrices, including both water and soil.

Despite the positive aspects of this technology, this study was intended as a proof-of-concept communication and the sensor was not thoroughly tested with a wide range of analytes and environmental samples. PFOA was the only fluorinated-anion-tested and other PFASs were not evaluated. The water and soil samples used were spiked in the laboratory; further testing with field-collected samples with a range of parameters (pH, salinity, etc.) would be helpful to establish performance with more complex and environmentally relevant samples. Furthermore, with a minimum detection limit of 3 ppm, this sensor would only be suitable for detecting high levels of PFASs.

3.1.2. Paper-Based Colorimetric Sensors

Menger et al. [25] developed and tested a color-changing, paper-based analytical device for PFOS in water samples, emphasizing the need for rapid, cost-effective, and portable methods to detect PFASs. The paper-based device was prepared with a wax layer to form a barrier that controls sample flow; a 60 mm circle was left uncovered by the wax and was dipped in methylene green. When exposed to PFOS, this circle would turn from green to purple due to ion-pairing of the sulfonic headgroup of PFOS and methylene green. Measuring the diameter of the resulting purple circle and comparing to a calibration curve allowed for quantification of the PFOS concentration (10–500 mg/L).

This method of detecting PFOS has several benefits, including ease of use and low cost. No special equipment or training would be needed to use this sensor, and the results would be available in <15 min. At <USD 1 to make a test strip, this has the potential to be a very low-cost analysis option. There was no interference from PFOA, presumably due to differences in hydrophobicity and ionic pairing due to acid strength between the two analytes; however, interferences from other PFAS analytes were not tested. However, as this method of measuring PFASs needs further refinement, including issues with interference, it is not currently commercially available.

Several issues need to be addressed before this type of PFAS sensors can be commercialized and used for PFOS detection. As the color-changing reaction is caused by the sulfonic headgroup of PFOS interacting with the methylene green dye, other compounds with similar headgroups could also cause this color change, creating potential interferences. Sulfonate and surfactants interfered with the dye’s intensity and size; carbonate also caused an interference by another mechanism. Substances that interfere with the electrostatic interactions needed for ion pairing between PFOS and methylene green also cause interferences, including heavy metals and high levels of salt (>1.0 M NaCl). The device also has issues when the sample’s pH is ≤4 or ≥8, requiring samples outside the ideal pH range to be neutralized beforehand. Menger et al. [25] suggest several methods to deal with these interferences (barium addition, cation exchange resin, filter-based cation exchange membrane, neutralizing the pH), but these additional steps would detract from some of the main benefits of using this sensor, particularly, its ease of use and quick timeframe. Detection limit is another important consideration for this technology; at 10 mg/L, this sensor has the potential to detect and estimate high concentrations of PFOS at impacted sites but would not be relevant to lower PFAS concentrations (including those associated with regulatory or advisory values) without further refinement. Furthermore, this method is specific to PFOS detection and was not intended to measure or detect levels of other PFAS analytes.

3.1.3. Water-Soluble Fluorescence Probes

Researchers have evaluated fluorescence as an analytical technique for measuring aqueous PFASs [28,32]. PFOS is quantified by altering the fluorescence intensity of a fluorescent probe or compound that emits light at a specific excitation wavelength. The fluorescent intensity of the probe can either be increased or decreased in a PFOS-concentration-dependent manner. Fluorescent intensity is decreased when PFOS interacts directly with solutions comprising a cationic porphyrin [33] and a cationic perylene diimide derivative [32]. Fluorescent intensity was increased when PFOS was added to a solution comprising siloxane-based tertiary ammonium salts and solutions containing erythrosine B [28]. When PFOS was added, disaggregation occurred between the siloxane based tertiary ammonium salts and erythrosine B, restoring the fluorescent intensity of the probe.

Detection of PFASs using fluorescent “probes” (i.e., encapsulated solution) has multiple benefits. Analyses are relatively simple and rapid (minutes), requiring mixing of a couple solutions and measuring the sample in a cuvette on a spectrofluorometer. Initial upfront costs would include the purchase of a spectrofluorometer, for which there are field-ready versions and consumable solutions. Based on the reviewed literature, this is an emerging technology and relevant field interferences require evaluation. Wang and Zhu [39] demonstrated high PFOS recoveries (>90%) in stream water and water containing homogenized fish tissue. However, copper (Cu²⁺) created interference, but was alleviated by the addition of EDTA. Low interference was reported with additions of common cations and anions found in natural waters [28,32]. Reported detection limits using these methods range from 3.3 ppb PFOS (0.008 µM) to 1350 ppb PFOS (4.65 µM) in laboratory buffered water [28,32,33].

3.1.4. Fluorescence of Amplifying Fluorescent Polymers/Conjugated Polymers

Conjugated polymers (CPs) are polymers that can transform the binding and release of a constituent with the CPs into a visual and measurable optical response. The advantages of CPs are that they can fluoresce brightly and result in clearly visible fluorescent changes at low concentrations of constituents [57]. In Concellón et al. [34], CPs in the form of a film and two different nanoparticles were used to detect PFOA and PFOS in milliQ water, DI water and groundwater. The CP used had fluorinated regions (e.g., lower-energy traps for the excitons and emission), allowing for binding of the two PFASs onto the polymer surface and resulting in a shift in the fluorescent peak towards lower energy wavelengths (e.g., blue to blueish-green). Using a different CP, Chen et al. [35] added PFOA and PFOS to stock solution containing the CP, which was prepared in DMSO and then diluted in HEPES buffer. The presence of PFOA and PFOS changed the fluorescence of the CP from blue to magenta with a concentration-dependent gradient of color and intensity.

CPs for detection of PFASs are an emerging technology with multiple advantages. Data would be available nearly in real time, dependent only on the time required for PFASs to diffuse through water to the CP surface. Concellón et al. [34] allowed for 1 h before determining a result; however, less time may be sufficient. Costs are anticipated to be relatively low and would include the purchase of the light source to induce fluorescence and CP material that may be reusable following washing [34]. Chen et al. [35] developed a portable device for measuring fluorescence of the CPs that is compatible with a cellular phone and associated applications. The device consisted of an on/off switch, UV light and cuvette holder attached to the camera of a cellular phone. An existing app was then used to determine color. Detection limits depend on the CP and its form (i.e., film or nanoparticle) with lowest reported limits of 0.08 ppb (0.2 nM) PFOA and 0.35 ppb (0.7 nM) PFOS in the work of Concellón et al. [34] and 2.53 ppb (6.12 nM) PFOA and 7.2 ppb (14.3 nM) PFOS in the work of Chen et al. [35].

3.1.5. Fluorescent Imprint-and-Report Sensor

The approach developed by Harrison and Waters [36] is a modification to a fluorescent sensor array using dynamic combinatorial chemistry (see Corbett et al. [58]) and dynamic combinatorial libraries (e.g., approach of using collections of molecules that can interconvert through reversible chemical reactions under thermodynamic control) containing macrocyclic building blocks imprinted with a fluorophore termed “imprint-and-report” sensing. The mechanism for this analytical process involves using a molecular building block (in this case, dithiol) that can reversibly interact with each other (other dithiol molecules) to create the dynamic combinatorial libraries. When a foreign molecule (i.e., PFASs) is introduced, the equilibrium of the dithiol interactions changes and can lead to an amplification of the number of interactions with the foreign molecule that can be sensed based on the fluorophore and analyzed using statistical tools like principle component analyses.

In this case, the authors tested multiple PFASs individually and in mixtures. The PFASs evaluated included PFOS, PFOA, perfluoro2-propoxypropanoic acid (GenX), perfluoroheptanoic acid (PFHpA), perfluorohexanoic acid (PFHxA), and perfluoropentanoic acid (PFPeA). Mixture concentration ratios were based on environmental contamination observed from an impacted location in North Carolina, USA. The results from this study indicated that 5.3–10 ppb (5–20 nM ranges) could be achieved in spiked tap water. The authors provide a proof-of-principle approach that could potentially be refined to environmentally relevant water samples. A benefit of this process if scaled would be the application for high-throughput analysis for multiple samples analyzed simultaneously with a multi-well microplate spectrophotometer.

3.1.6. Oil Interfacial/Amphiphilic Approaches

Trinh et al. [27] investigated the detection of PFASs and fluorinated surfactants using changes in optical emission in oil–water droplets and simple spectroscopy. This approach is based on the observed surfactant properties of PFASs, with higher concentrations resulting in lower surface and interfacial (boundary between the two immiscible phases) tensions [59]. Trinh and coauthors [27] used a Janus emulsion, that is, a droplet containing two distinct phases, comprising hydrocarbon-oil and fluorocarbon-oil with fluorescent dyes. As the droplet’s internal interface curves due to the introduction of PFASs, the emissive light of the fluorescent dye is focused, producing higher intensity when measured spectroscopically. In this study, there were positive correlations between increasing PFAS concentrations and decreasing surface tensions. This approach was demonstrated to detect a range of concentrations of PFOS, ranging from 2.5 ug/L to 600 mg/L and PFOA ranging from 4.5 ug/L to 600 mg/L. While this approach has relatively high detection limits as compared to most reported analytical tools or technologies, there may be applications for screening higher concentration sources (e.g., reject water, concentrated samples, leachates, etc.) to better inform more intensive sampling and analysis. Benefits would potentially include benefits to rapid and on-site monitoring of PFASs.

3.1.7. Bacterial Biosensor

Bacterial biosensors include genetically engineered bacteria, typically designed to fluoresce in a concentration-dependent manner when a promoter gene is induced or activated [60]. Sunantha and Vasudevan [30] developed a sensor for water samples using genetically engineered bacteria that fluoresce in the presence of PFOS and PFOA. After first testing several bacteria species and strains for optimal growth when cultured with PFOA and PFOS, Pseudomonas aeruginosa was chosen for the biosensor. This bacterium was transformed by the addition of a regulatory gene (defluorinase) and a reporter gene (green fluorescent gene). The defluorinase enzyme specifically acts on and breaks down halogenated compounds including PFOS, PFOA, and other PFASs. Upon induction of the defluorinase regulatory gene, the green fluorescent gene is also expressed. This fluorescence can be photographed with a microscope and measured with a spectrofluorometer, allowing PFAS levels to be quantified.

Using this biosensor for PFAS analysis has several benefits, particularly a detection limit as low as 10 ppt and high specificity for PFASs in general, but not a high degree of selectivity for specific PFASs. The selectivity of this biosensor was tested using various organic pollutants: PFOA, PFOS, sodium fluoroacetate, tri-chlorophenol, penta-chlorophenol, fluorene, naphthalene, phenanthrene, anthracene, α-endosulfan, β-endosulfan, chlorpyrifos and cypermethrin; interferences were only found with tri- and penta-chlorinated compounds, as defluorinase acts by breaking down chemical bonds with halogens. The sensor was also shown to work with complex field-collected water samples from polluted sites with a wide range of salinity and total dissolved solid concentrations.

One major issue that hinders the use of this technology in a field-setting is the sample incubation time. A 24 h incubation period was necessary to reach the maximum fluorescent intensity required for PFOA and PFOS detection, as sufficient time is required for the bacteria to breakdown the PFASs and for green fluorescent gene protein expression. This method also would not be able to distinguish between different PFAS analytes.

In another study, Young et al. [29] evaluated propane 2-monooxygenase alpha (prmA) from the bacterium Rhodococcus jostii strain as a potential promoter gene. When exposed to PFOA, gene expression increased by a factor of 3. A genetically engineered plasmid was then designed consisting of the promoter, prmA and the reporter gene, monomeric red fluorescent protein and was inserted into R. jostii. However, upon exposure of the genetically engineered R. jostii to PFOA, a fluorescent signal was not visually apparent.

Biosensors from bacteria can be made available in a variety of forms, including both in vivo and in vitro assays [60]. These biosensors can be sensitive to changes in environmental conditions; therefore, a benefit of an in vivo assay is that the cellular membrane protects the internal reactions within a cell necessary for the workings of the biosensor. Researchers have evaluated methods to immobilize bacterial biosensors using biocompatible techniques. Therefore, biosensors evaluated using cultured cells may be able to be transitioned to more field-ready sensor types.

3.1.8. Ligand Binding

Human liver fatty-acid-binding protein (hLFABP) has been engineered to create fluorescence-based sensors for PFAS detection [37,38]. PFASs are known to bind with this protein, as they share structural similarities to hLFABP’s natural ligands. In the work of Mann et al. [37], the protein’s ligand binding pocket was genetically modified by adding fluorophores, of which acrylodan had the lowest detection capabilities and was subsequently used for further testing. When PFASs bind to this pocket, the polarity around the pocket decreases, leading to lowered energy emission and a blue-shifted fluorescence emission spectrum that can be quantified with a spectrometer. In the work of Mann et al. [38], a green fluorescent gene was used instead of acrylodan and the resulting sensor was introduced into E. coli cells. The E. coli would then fluoresce in the presence of PFOA, showing the feasibility of using a whole-cell biosensor.

Sensors with hLAFBP have the potential to be a useful tool for measuring PFASs in the environment. Though Mann et al. [37,38] only tested the sensors in aqueous samples, they can be adapted to detect PFASs in soil and biosolids, particularly by modification into a whole-cell biosensor, as in the work of Mann et al. [38]. With detection limits of 112 ppb for PFOA, 345 ppb for PFOS, and 1090 ppb for PFHxS using the acrylodan-based sensor [37] or 236 ppb using the GFP-based biosensor [38], these are capable of measuring environmentally relevant levels of PFASs associated with contaminated sites. Another benefit of this method is a sample turnaround time, with a 5 min incubation time and a 5 min equilibrium time reported in the work of Mann et al. [37].

Although this sensor technology appears promising, further testing is needed to investigate performance with field-collected samples. Only a single interferent was tested, the anionic surfactant sodium dodecyl sulfate. Additional testing with other possible interferants and across a range of water quality parameters would further demonstrate the applicability of the sensor to environmental samples. The sensor was tested using field-collected water from a creek, but this water was not characterized. Mann et al. [37] discuss that whole-cell biosensors can be used for soil and solid sample analysis, but their hLFABP-E. coli biosensor was only tested in aqueous matrices and was not demonstrated on soil or solid matrix samples. As a spectrometer or microplate reader was used to analyze samples, this sensor was more suitable to a laboratory rather than field setting.

3.1.9. Aptamer-Based

Park et al. [31] developed a sensor to quantify PFOA using DNA aptamers: short, single-stranded DNA molecules that undergo three-dimensional shape changes in the presence of target molecules, allowing them to bind these target molecules with high affinity and selectivity. Aptamers shown to exhibit a high binding affinity for PFOA were modified by adding fluorescein to the 5′-end. These aptamers were then bound with capture DNA (cDNA) modified with dabcyl to act as a quencher for fluorescence. Upon exposure to PFOA, this aptamer would bind with the PFOA instead, detaching the cDNA and resulting in fluorescence. This fluorescent signal could then be quantified with a fluorescence reader.

This sensor has the potential to be used for relatively fast detection of PFOA in water samples by simply mixing the aptamer and sample and waiting 40 min for the reaction. This sensor has been successfully tested with wastewater samples, indicating it can be used for complex environmental water samples. Overall, the sensor is not susceptible to interferences; interferences were not observed with inorganic salts, dissolved organic matter, or most 4–8 carbon PFCAs and PFSAs, except PFHxS and PFHpA. Wastewater results from this method were compared to LC-MS results; overall, the two methods were similar (percent error ranging from 0.6 to 13.3%), demonstrating the accuracy of this method.

Park et al. [31] mention several limitations associated with this sensor technology. Though this method was intended to only measure PFOA, it was found to also bind PFHxS and PFHpA, but not other 4–8 carbon PFCAs and PFSAs. A spectrometer or fluorescence reader is needed to quantify the samples; so, this analysis is more suitable for a laboratory rather than field setting. Though the detection limits of this method are suitable for measuring elevated environmental PFAS concentrations (i.e., 70.4 ppb limit of detection (LOD)), such as those associated with industrial sites, contaminated groundwater, and spills, this method would not be suitable for measuring low levels of PFASs, including those relevant to drinking water regulatory levels.

3.1.10. Dual DNA Oligonucleotide and Lysozyme Fiber Biosensors

Two biosensors were developed by Zhu et al. [39] consisting of a DNA oligonucleotide (short section of DNA) and thioflavin T (Tht) and the other consisting of a synthesized lysozyme fiber and Tht. These were evaluated for their ability to detect hydrophobic long-chain PFASs including perfluorododecanoic acid (PFDoA), perfluorodecanoic acid (PFDA) and perfluorotetradecanoic acid (PFTeDA) in tap water and serum. Tht is the fluorescent component of these biosensors and its fluorescent intensity increases when the rotation of the chemical structure is restricted by binding with the other components in the biosensor. For example, when Tht binds with the DNA oligonucleaotide, the fluorescent signal of Tht is enhanced. The addition of PFDoA, because of its hydrophobic nature, results in binding of PFDoA with the DNA bases of the oligonucleotide, which releases Tht, and the fluorescent signal is decreased. Because of differences in the hydrophobicity of the different long-chain PFASs, the ratio of the change in fluorescent intensity during binding and unbinding for the two separate biosensors could be used to identify different PFAS analogues. These biosensors were also evaluated against shorter-chain PFASs (e.g., PFOA and PFHxS) and long-chain surfactants (e.g., sodium dodecyl sulfate and lauric acid), resulting in no detectable impact to the fluorescent signal.

Presumably, based on the requirements of sample preparation, this method would be challenging to perform in a field setting as tap water and serum samples require a preconcentration and purification step with solid-phase extraction. This involves a series of washing, extraction and drying steps with a vortex and centrifuge. However, with SPE, the limit of detection for PFDoA decreased 98.2 ppb (0.16 µM) to 0.46 ppb (0.75 nM).

3.1.11. Portable Surface-Enhanced Raman Spectroscopy Sensors with Graphene + Silver Nanoparticles on a Kapton Film

There has been progress in developing low-cost surface-enhanced Raman spectroscopy (SERS) substrates for PFAS detection [41]. The mechanism of SERS is generally adsorption of a compound or molecule to a metal substrate using nanoparticles, which changes the electromagnetic signal due to the excitation of surface plasmons. A benefit of this approach is that electromagnetic enhancement typically has higher signal-to-noise ratios compared to some chemical enhancement methods [61]. McDonnel and co-authors [41] developed an aerosol jet-printed SERS sensor sensitive to PFOA and PFOS using adsorption properties of graphene + silver nanoparticles. The substrate is printed with an aerosol Jet printer applying silver nanoparticles and graphene inks to a Kapton film allowing for a low-cost SERS application. The addition of graphene is necessary to enhance signal increasing sensitivity.

SERS is field-portable and has a shelf life of 9 months (longer timelines have not been investigated). These substrates are intended to be used as swabs for contaminated surfaces. The SERS substrates detected concentrations of ~0.5 ng/L and 70 ng/L for PFOS and PFOA, respectively, in basic (pH = 9) conditions. The relative differences in detection limits were attributed to the higher hydrophobicity of PFOS, which would lead to more effective adsorption to the nanoparticles and the lower pKA of PFOS, which would lead to an increased number of hydrogen bonding sites as compared to PFOA.

This technique would need to be tested for other PFASs; however, the ability to achieve low limits of detection is appealing considering that is not possible with many portable detection methods for PFASs. The low-cost printing of SERS sensors has been developed and applied to other emerging technologies for a variety of environmental health and human safety monitoring purposes [62,63]. In the case of application to PFOA and PFOS, the promising aspect of this technology for application to field investigations is the low limit of detections potentially feasible using SERS substrates. This approach does not require a preconcentration step as compared to an HPLC approach and could likely offer point-of-sampling sensing capability. Additionally, the reported detection levels for PFOA and PFOS are comparable to LC/MS. Due to the stage of this approach in development, it has yet to be rigorously evaluated in terms of potential confounding factors in real-world matrices. This study was conducting using pure PFAS standards in RO water only; therefore, competition of molecular adsorption has yet to be determined. The authors state that the approach provides a path forward to simple, cheap, and portable, which would be enticing for on-site applications [41].

Table 2. Summary of optical-based emerging technologies for the detection of PFASs.

Method	Media	Analyte, LOD (ppb)	Selectivity Evaluated	Relative Cost	Reference
Optical/colorimetric porphyrin-based detector	Soil, water, organic solvent	PFOA, 3000	No	Low	[26]
Optical/colorimetric paper-based sensor	Deionized water	PFOS, 10,000	Yes	Very Low	[25]
Fluorescence with cationic siloxane and erythosine B	Not reported	PFOS, ~1350	Yes; ions (NO₂, NO₃, F, Cl, S₂O₃; SO₄)	Low	[28]
Fluorescence with perylene diimide	Buffered water	PFOS, 14	Yes—ions (e.g., Na, K Ca, Cl) and other PFASs (e.g., PFOA, PFBS)	Low	[32]
Fluorescence with porphyrin	Buffered MilliQ water	PFOS, 3.3	Yes—cations including Cu, Mn, Fe, Cr, Mg, Zn, Ba, Cd, Pb, Hg, Al, Ni and surfactants. stream water samples, water containing blended fish tissue	Low	[33]
Fluorescence with conjugated polymer	MilliQ, deionized, and well water	PFOA, 0.08 PFOS, 0.35 PFBA, NA	Yes	Low	[34]
Fluorescence with conjugated polymer	Buffered water	PFOA, 2.53 PFOS, 7.2	Yes—surfactants, PFAS analogues, anions (Cl, I, and others) and cations (Ca, Fe, and others)	Low	[35]
Fluorescence imprint-and-report with dithiol dynamic combinatorial libraries	Spiked tap water	PFOA, 8.3 PFOS, 10 PFHxA, 6.3 GenX, 6.6 PfPeA, 5.3 PFHpA, 7.3	Yes—fully distinguished among mixtures of six PFASs evaluated	Medium	[36]
Fluorescence with biphasic oil–water droplets	Water, formation of Janus droplets using surfactants, PFASs, and oil mixtures	PFOA, 4.5 PFOS, 2.5	Yes	Low	[27]
Fluorescence synthetic bacterial biosensor	Drinking, river, and wastewater	PFOA, 0.01 PFOS, 0.01	Yes—field water samples, cross reactivity across PCBs, PAHs, pesticides	Medium, >24 h response time	[30]
Fluorescence genetically modified bacterium	Culture medium	PFOS, NA	No	Low	[29]
Fluorescence ligand-binding, human liver fatty-acid-binding protein	Buffered water, creek water	PFOA, 112 PFOS, 345 PFHxS, 1090	Yes	Low	[37,38]
Fluorescence aptamer-based	Wastewater effluent	PFOA, 70.4	Yes	Medium	[31]
Fluorescence dual DNA and lysozyme fiber biosensors	Tap water and serum	PFDoA, 98.2 PFDA, 648 PFTeDA, 57.1	Yes—short-chain PFASs (e.g., PFOA and PFHxS) and long-chain surfactants (e.g., sodium dodecyl sulfate and lauric acid)	Low	[39]
Aerosol jet printed surface-enhanced Raman substrates (SERS)	Buffered water (Basic; pH = 9)	PFOA, 0.07 PFOS, 0.0005	No	Medium	[41]

Note(s): LOD = limit of detection reported by corresponding study; PCB = polychlorinated biphenyls; PAH = polycyclic aromatic hydrocarbons.; NA = not available.

3.2. Electrochemical-Based Methods

PFAS molecules are generally considered to be electrochemically inert, thereby limiting the application of electrochemical sensors. However, novel approaches have been developed to circumvent these challenges and achieve environmentally relevant detection concentrations of PFAS (Table 3). In many cases, the sensor approach is aimed at using the target molecule to block the signal, which results in a quantifiable reaction [47]. As an example, as the electrode surface area decreases with increasing binding of PFASs, the result is a proportional signal change. This electrochemical impedance approach has been explored for its potential in rapid detection using (1) functionalization with a molecularly imprinted polymer (MIP); (2) a modified ion exchange polymer; and (3) increasing adsorption to electrodes via sorbent materials. Many of these approaches offer unique characteristics in terms of effecting scalability to commercialization, sensor miniaturization, and lowering costs [11].

3.2.1. Gold Nanostar Molecularly Imprinted Polymer

Lu et al. [44] detected PFOS in tap water at a LOD of 7.5 ppt with an ultra-sensitive voltammetric sensor that used a glassy carbon electrode (GCE) with a thin coating of gold nanostar (AuNS) and MIP. The AuNS coating can boost the voltammetric response of the blank signal intensity with the oxidation of iron with a FcCOOH probe a commonly used electrochemical redox probe. An advantage of MIP sensors is how quickly the layer is formed via electropolymerization of monomers in the presence of the target analyte (PFOS). PFOS is detected by measuring the FcCOOH oxidation peak through differential pulse voltammetry (DPV). An advantage of the MIP/AuNS/GCE sensor is the low relative standard deviation, indicating high reproducibility. The relative standard deviation was lower than that of results achieved via standard LC-MS analysis. Additionally, optimization of the thickness of the sensitive AuNS coating and MIP layer enables a low LOD, comparable to the USEPA method 537.1 detecting PFOS in tap water.

A limitation of the method is the potential for interference due to the presence of other PFAS compounds occupying binding sites and affecting the FcCOOH binding isotherm. Once a threshold (5 nM) of binding occurs, the sensitivity of the analysis decreases by more than half. This has the potential to severely limit the practical application of the MIP based sensor. As a result, an underestimation of detection can occur based on the interference from PFBA and PFBS and, to a lesser extent, PFOA, PFHxA and PFHxS. Since PFBA and PFBS have a decreased molecular size, they can accumulate in the MIP layer, preventing PFOS from being able to bind. Humic acid and chloride ions can also decrease the DPV response of PFOS, thus increasing the detection limits due to matrix interferences.

3.2.2. Co-Doped Carbon Nanoarchitectures Molecularly Imprinted Polymer

Another study investigating MIP structures, Pierpaoli et al. [48] tailored the MIP substrate using a combination of vapor-deposited boron and nitrogen co-doped with diamond-rich carbon nanoarchitectures. The detection method was a combination of DPV and EIS. The MIPs’ structure was manipulated to change the nanoscale morphology and chemical composition to improve the surface interactions with PFASs. This approach showed strong affinity for PFOS, roughly an order of magnitude higher affinity versus similar analytes (PFHxS, PFOA, and PFDA). The modified MIP was demonstrated to detect PFOS with an LOD of 1.2 µg/L (ppb) in tap water and wastewater. However, performance was diminished in a more complex landfill leachate water, likely due to the presence of relatively high concentration of sulphates and humic substances. This study successfully demonstrated the MIPs approach using more complex wastewater, highlighting the potential of novel enhanced MIP architectures to be selective for PFOS with competing ligands. Moreover, the study demonstrated detection capabilities in the low ppb region.

3.2.3. Molecularly Imprinted Polymers: Oxygen as a Redox Probe

Clark and Dick [47] applied a unique approach of using oxygen in river water as a mediator to calibrate PFOS concentrations on a MIP-modified carbon substrate. The MIP approach was paired with DPV and EIS to measure PFOS in river water [47]. The authors primary objective was to evaluate whether ambient oxygen concentrations in river water could be used as an electron mediator to quantify PFOS. Typically, the MIP approach requires the use of electron mediators (e.g., a ferrocene derivative) to achieve oxidation in the process for successful reaction to occur. Clark and Dick [47] recognized the limitation of this step for field-deployable sensors and evaluated whether oxygen in ambient river conditions could be used as an alternate approach. The results indicated that the sensor showed good selectivity toward PFOS in the presence of competing molecules (e.g., chloride, humic acid). However, selectivity for PFOS in the presence of other PFAS analytes was not evaluated. The sensor detection range was 0.0025 to 0.250 ppb, with a limit of detection of 0.0017 ppb for PFOS.

This approach for MIP-based technologies is very promising, as these can achieve low limits of detection in the ppt range. Additionally, the performance in the presence of natural inferences (e.g., humic acid and chloride) did not degrade the sensor performance.

3.2.4. Bare Platinum Electrode

Other studies have explored the application of a platinum electrode for electrochemical sensing of PFASs. Gogoi et al. [42] evaluated a platinum electrode using a potassium phosphate buffer with potassium ferricyanide as the electrolyte/redox probe solution and differential pulse voltammetry (DPV) and electrochemical impedance spectroscopy (EIS) to detect concentrations of PFOS. The study indicated that the adsorption of PFOS to the electrode occurred in three distinct adsorbate–adsorbate interactions occurring in concentrations ranging from 0 to 50 ppb. However, this study did not evaluate the selectivity of this method for other PFAS analytes. Based on electrochemical data, Gogoi and coauthors [42] posited adsorption mechanisms based on the electrochemical data obtained during the study. At low concentrations, the adsorption had a random orientation on the platinum surface, with molecules lying either flat or “head on” (i.e., sulfonate group attached to electrode surface). At medium concentrations (between 5 and 50 ppb), adsorption to the electrode was primarily head on as a consequence of electrostatic repulsion between neighbor PFOS molecules. At concentrations above 50 ppb, molecular interactions among the PFOS molecules increased due to micelle-like interactions and there was greater competition between platinum surface and a “bi-layer” of PFOS molecules. An additional finding of this study was that PFOS could be selectively adsorbed on the platinum electrode even in the presence of other inferences commonly present in natural waters (e.g., humic acids and chloride).

There are numerous benefits of developing bare platinum electrodes paired with electrochemical sensing for PFASs due to the relatively simple design. If successfully developed, platinum sensors could provide a fairly robust and versatile option for field sensing capabilities without the need for complex sample preparation. Additionally, in the context of site investigations, it provides a practitioner with the real-time capability of bounding environmental concentrations to inform more intensive sampling with approaches that are more selective and sensitive.

3.2.5. Platinum Sensor with Selective Perfluorinated Anion Exchange Ionomer Coating

Sahu et al. [43] developed a method for quantifying PFOA with an electrochemical sensor. The platinum sensor has a selective perfluorinated anion exchange ionomer (PFAEI) coating for direct sensing of PFOA with a commercially available screen-printed electrode. For direct sensing to occur electrostatic, van der Waal, or hydrophobic interactions work cooperatively between the PFAEI coating and PFOA. Different coatings for the electrochemical sensor were tested, including a quaternary benzyl pyridinium chloride poly (arylene ether sulfone), hydrocarbon cation exchange ionomer, and perfluorinated cation exchange ionomer. These other coatings were ineffective for sensing PFOA in comparison to the PFAEI coating. The interaction between the perfluoro backbones and side changes through van der Waal interactions and quaternary ammonium groups on the PFAEI coating allow for direct sensing of PFOA through anion exchange to occur. Testing the platinum electrode with different PFOA concentrations revealed that as the PFOA concentration increased, the current response increased until it peaked and shifted from −0.35 to −0.45 volts. The effects of interfering anions (nitrate, phosphate, and sulfate) on PFOA sensing were evaluated; the result was that the current response increased as the concentration of PFOA increased. This indicated that the PFAEI coated wire is a promising candidate to detect PFOA with interfering anions present.

The use of an electrochemical sensor with a PFAEI coating has the advantage of being able to quantify PFOA without the need for redox probes. The LOD was 6.51 ppb in buffered deionized water and drinking water with interfering anions present (nitrate, phosphate, sulfate). The LOD at 6.51 ppb is orders of magnitude above the current MCL in drinking-water standards (i.e., 0.004 ppb for PFOA and PFOS). This method is promising for the future since it is a direct method of measuring PFOA that does not require the addition of redox active probes. Karimian et al. [64] tested an electrochemical sensor to indirectly detect PFOS with a LOD of 17 ppt in deionized water using a ferrocenecarboxylic acid redox-active probe. In another study by Glasscott et al. [65], GenX was quantified in river water to a concentration as low as ~1 pM (~0.33 ppb). Improvements to lower the LOD have the potential to be completed by optimizing the PFAEI coating thickness by using the Langmuir–Blodgett technique.

Limitations of the electrochemical sensor with a PFAEI coating include the requirement of an optimal pH of ~1.5. The acidic pH keeps PFOA neutral for adsorption onto the PFAEI while ensuring a large enough proton concentration exists to minimize ohmic losses. Other future work includes improving the electrochemical sensor by increasing the surface area through an alternative circuit design. In turn, the PFAEI thickness needs to be optimized for the promotion of electrocatalyst. A limitation of the ferrocenecarboylic acid probe method of Karimian et al. [64] and Glasscott et al. [65] is the need for additional redox active probes and chemical reagents, which limits the convenience and field capability of this PFOA detection method.

3.2.6. Ionized PFAS Detection at Interfaces Between Immiscible Solutions

Islam et al. [45] utilized ion transfer voltammetry to detect PFASs at micropipette-based interfaces between two immiscible electrolyte solutions (µITIES). This technique is used to overcome the chemical stability of PFASs by applying electrochemistry at the interface between two immiscible electrolyte solutions producing measurable electrochemical signal. Simply, this process relies on the movement of PFAS ions across the interface of two immiscible liquids. Garada et al. [66] developed a procedure to detect PFOS by ion transfer stripping voltammetry at the ITIES formed between an aqueous electrolyte and a thin film on an electrode. They were able to achieve a LOD of 0.025 ppb for PFOS requiring a 30 min preconcentration step. Viada et al. [67] expanded on this work, studying matrix effects and achieved an LOD for PFOS of 0.015 ppb with a 5 min preconcentration step. Islam and coauthors expanded on these concepts focusing on selectivity by investigating its compatibility with four PFAS analytes individually and as mixtures.

Islam et al. [45] fabricated micropipettes for this application and they function by placing an organic phase electrolyte inside the pipette and an organic reference solution; then, the pipette is placed into a beaker with the aqueous phase forming ITIES at the tip of the pipette. Cyclic voltammetry (CV) and differential pulse voltammetry (DPV) were both used in forward/reverse scans. CV was viable for individual analytes; however, DPV needed to be used for mixtures. The study specifically focused on PFOA, PFBS, PFHxS, and PFOS individually and as mixtures; the limits of detection were 21 ppb (0.05 µM), 9 ppb (0.03 µM), 8 ppb (0.02 µM), and 20 ppb (0.04 µM), respectively, using DPV. It was shown that the combinations of PFOA/PFBS and PFHxS/PFOS were not detectable due to interferences. Using mixtures, it is not possible to achieve the same LODs as those of previous studies. To have utility and support commercial applications, sensors will need to be capable of screening for multiple PFASs.

3.2.7. Ion-Transfer Electroanalytical Detection

Lamichhane and Arrigan [46] developed a method of detecting PFOA through its electrochemical behavior via ion transfer voltammetry. The ion transfer voltammetry also uses a microinterface array between two immiscible electrolytic solutions (i.e., µITIES, as discussed in previous section). Testing the cyclic voltammetry of PFOA was accomplished with applied potential at the interphase between the aqueous and organic layer. At the interphase, ionized PFOA is transferred either forward through the aqueous phase or with a reverse scan through the organic phase to be detected. Both chloride and sulfate salt were evaluated to determine which would allow for the greater potential window for increased detection of PFOA. Sulfate had the larger window for a greater sensitivity and a lower detection limit (0.3 µM) compared to chloride salt, but chloride salt was used in the study since it is common in environmental and biological samples.

Testing differential pulse voltammetry (DPV) for the detection of PFOA revealed that the peak current linearly increased in correlation with PFOA concentration. The lowest detection of PFOA (0.1 µM; 0.497 ppb) occurred with a backward scan to produce the highest sensitivity. The differential pulse stripping voltammetry (DPSV) has the potential for an improved LOD compared to DPV because DPSV uses the advantages of the intrinsic preconcentration at a set potential, and a voltametric scan along with the detection methods from DPV. Using DPSV and a preconcentration time of 300 s, Viada et al. [67] were able to detect PFOS down to a LOD in the pM. A lower detection limit for PFOS compared to PFOA may be due to PFOS being more lipophilic, which is more favorable to the preconcentration step. The preconcentration step can be tailored in voltage (0.1 V) for an even lower LOD with more modifications in further studies. Other areas for optimization of the method for electrochemical detection of PFOA include factors that enhance or hinder the movement of PFOA from the aqueous phase to the organic phase, the geometry of the pores of the membrane that form the µITIES array, and the preconcentration time (300 s).

One limitation of the method is matrix interference and its impacts on PFOA detection. Both drinking water and laboratory tap water were tested for the aqueous phase with the addition of LiCl. There was a slight shift in the detection peak from the chloride content, as well as a decrease in the peak. There were no impacts in the detected peak when clean water was evaluated as the matrix. Matrix effects from bovine serum albumin and β-cyclodextrin were also evaluated since both constituents can form complexes with PFASs. Both bovine serum albumen and β-cyclodextrin reduced the detection of PFOA due to the impacted the electrochemical response of PFOA by the formation of complexes that bind PFOA to remove it from being freely ionized to move across the ITIES. Another limitation is that POFA needs to be completely dissociated at the ITIES. For this to occur, the pH of the solutions needs to be ≥4 to keep PFOA in an anionic form before transfer through the interphase can occur for detection. Overall, the electrochemical method involves expertise and is time-consuming to complete. An expert is needed to run the software and use the laser to make the grid of the membrane before the glass membrane can be glued to a cylinder. It takes the glue 24 h to cure before the experiment can occur. Further research also needs to be completed to add a systemic preconcentration step which has the potential to lower the LOD further.

3.2.8. Nano-Electrochemistry with Silver Nanoparticles

Niu et al. [68] demonstrated that that silver nanoparticles can be modified with thiol-terminated polystyrene to produce aggregation when exposed to ppm levels of PFOA. Alternatively, Khan et al. [50] used silver nanoparticles to detect PFOS by single-particle collision electrochemistry (SPCE). This technique works based on the change in Faradaic charge transfer of silver nanoparticles when exposed to PFOS. The selectivity and sensitivity of the SPCE method relies on the concentration of nanoparticles and the interaction time between nanoparticles and the target analyte. Khan et al. [50] used 18.5 pM silver nanoparticles and a 20 min contact time between silver nanoparticles and 1 ppb PFOS. The selectivity of this technique was tested in the presence of structurally similar PFASs and various functional groups. These tests showed that the presence of both the SO⁻³ group and relatively long-chain PFASs is responsible for the interaction.

This method has a theoretical LOD of 0.04 ppb for PFOS. The use of SPCE with silver nanoparticles for the detection of PFOS is still in the exploratory stages. While there are some benefits, such as its sensitivity and selectivity, significant refinement is required before it is a viable technique for in-field PFAS detection.

3.2.9. Altering Fluorinated Self-Assembled Monolayers

Moro et al. [49] used fluorinated self-assembled monolayers (SAMs) manipulating available architectures to detect short- and long-chain PFASs. Both long-chain (PFOA and PFOS) and short-chain PFASs (PFHxS, HFPO-DA, and PFPA) were evaluated. Moro et al. [49] investigated the structural changes in fluorinated SAMs, specifically 1H,1H,2H,2H-perfluorodecanethiol (FDT-SAM). The main features that were focused on were pinholes (a site on the substrate is not covered by the monolayer) and defects (formed when the thiol chain does not stand in an upright position). This process aims to detect a signal where interactions of PFASs either decreases a signal, “switch-off”, or increases a signal, “switch-on”. A “switch-off” is when a detectable signal (e.g., fluorescence, electrochemical current) decreases or is quenched when PFASs interact with the sensing surface.

Ordered SAMs, pinhole/defect-free, and unordered SAMs, randomized pinhole/defect architecture, were both employed to determine their compatibility with both long- and short-chain PFASs. Ordered SAM was not applicable to PFAS screening due to short-chain PFASs not producing any noticeable change in SAM properties and any changes seen with long-chain PFASs were not reproducible. Unordered FDT-SAM was compatible with switch-off sensing strategy for both long- and short-chain PFASs. This work suggests the potential for utilizing the different SAM architectures for PFAS screening. This technique potentially lends itself as a building block to for future portable and rapid screening tools for detection of PFASs in wastewaters.

Table 3. Summary of electrochemical-based emerging technologies for the detection of PFASs.

Method	Media	Analyte, LOD (ppb)	Selectivity Evaluated	Relative Cost	Reference
Gold nanostar molecularly imprinted polymer	Tap water	PFOS, 0.0075	Yes	Low	[44]
Co-doped carbon nanoarchitectures molecularly imprinted polymer	Tap water, wastewater, landfill leachate	PFOS, 1.2	Yes, competing analytes in wastewater (Cl, sulfate, humic acid)	Low to medium	[48]
Molecularly imprinted polymers: oxygen as a redox probe	River water	PFOS, 0.0017	Yes, Cl and humic acid	Low to medium	[47]
Bare platinum electrode	0.1 M phosphate-buffered solution	PFOS, NA ¹	No	Low	[42]
Platinum sensor with selective perfluorinated anion exchange ionomer coating	Buffered deionized water and drinking water	PFOA, 6.21	Yes	Low	[43]
Ionized PFAS detection at interfaces between immiscible solutions	Water	PFOA, 20.7 PFOS, 20 PFBS, 9.0 PFHxS, 8.0	Yes	Low	[45]
Ion-transfer electroanalytical detection	Drinking and laboratory tap water	PFOA, 0.497	Yes, ions Cl, SO₄	Low to medium	[46]
Nano-electrochemistry with silver nanoparticles	Water	PFOS, 0.04	Yes	Low to medium	[50]
Altering fluorinated self-assembled monolayers	Water	PFOA, NA ² PFOS, NA ² PFHxS, NA ² HFPO-DA, NA ² PFPA, NA ²	Yes, but poor selectivity among PFASs	Low to medium	[49]

Note(s): ¹ Data not available; based on sorption kinetics, there were three distinct behaviors at the following concentration ranges: 0–5; 5–50; and 50 ppb. ² Data not available; linear trends were observed in the concentration range from ~40 to 400 ppb.

3.3. Nonconventional Methods

3.3.1. Hydrogel-Swelling-Enabled PFAS Detection

Savage et al. [51,69] researched the potential of using the swelling of hydrogels to detect PFASs. This hydrogel sensor consisted of the polymer poly (N-isopropylacrylamide) (PNIPAM) copolymerized with 2,2,2-trifluoroethylacrylate (TFEA). PFOS was shown to increase the swelling behavior and alter the thermodynamics of the PNIPAM gel [69]; PFOA, on the other hand, results in shrinking of the gel due to electrostatic or hydrogen bonding [69]. Quantification of the swelling behavior was measured with dynamic light scattering or fluorescence using Förster resonance energy transfer (FRET)-compatible dyes [51]. Benefits of this method include its straightforward application and minimal preparation: simply adding the sample to the hydrogel and quantifying the reaction in a particle size analyzer or spectrometer (for fluorescent samples).

Further refinements to this method are needed before it can be commercialized. Additional selectivity testing to common interferences would also be helpful; preliminary testing indicated that several compounds, including phenol and sodium 1-octanesulfonate, could cause a shrinking reaction similar to PFOA’s behavior. Refinements to the sensitivity of this sensor are needed to increase its relevance to environmental samples, as the method’s detection limit of 50 ppm (mg/L) for PFOS is well above regulatory levels [70], typical environmental levels [71] and levels of concern for aquatic life [72]. As Savage et al. [51,69] only tested samples prepared with deionized water, additional testing with environmental samples is needed. Further testing with multiple PFAS analytes would also be beneficial to demonstrating whether this method could be used to quantify analytes other than PFOS.

3.3.2. PFOA Sensing Through Flow Rate Analysis

Breshears et al. [52] developed a smartphone assay capable of detecting PFOA by analyzing capillary action on a cellulose-based chip. Different reagents (L-lysine, casein, and albumin) known to interact with PFOA via hydrophobic interactions, electrostatic interactions, and hydrogen bonding were loaded into the microfluidic channel of the chip. These competitive interactions, which were recorded on video and analyzed using smartphones, altered the surface tension and thus capillary flow rate.

This assay has the potential to rapidly assess samples for the presence of PFOA with minimal preparation, equipment, and training. Once the reagent has been preloaded and dried (<30 min process), sample analysis takes less than 5 min. As this assay uses a smartphone rather than specialized laboratory equipment, it is portable and can be performed in a field setting. Another benefit of this assay is its sensitivity, with detection limits as low as 0.01 ppt in laboratory water and 1 ppt in filtered wastewater. As such, this technology would be capable of detecting PFOA concentrations applicable to the USEPA’s drinking water regulation [70]. Specificity assays indicated low susceptibility of interference for the three surfactants tested. Initial tests with PFOS indicated similar results to PFOA, suggesting the technique could also be used to detect other PFASs.

3.3.3. Thermal Detection of PFASs Using Molecular Imprinting

Detection of PFOA using molecularly imprinted polymers (MIPs) involves measuring changes in thermal properties, such as heat transfer or degradation, using a thermal sensing platform. MIPs are created through polymerization in the presence of a template molecule that is extracted from the polymer, resulting in synthetic receptors complementary to the target analyte. This, in combination with the heat transfer method, allows for quick and reliable detection of PFOA. The heat transfer method monitors the thermal resistance across an aluminum chip through continuous temperature input by a heat provider and the output temperature in a liquid-flow cell above. The chip is coated with a receptor layer and as a target analyte binds to the receptor, its thermal resistance changes and can be registered.

Tabar et al. [53] tested this combination of techniques using a controlled buffer solution, river water, and soil. All test matrices were spiked with PFOA and soil samples were extracted using 20 mL of DI water. LODs for buffer solution, river water, and soil were 0.001 ppb (22 pM), 0.038 ppb (91 pM), and 0.064 ppb (154 pM), respectively. The LODs that Taber and co-authors were able to achieve are promising for the future applications of this technique and are environmentally relevant. However, this study is an initial proof of concept and will need to be further validated with environmental samples. The MIP used by Taber et al. [53] is specific to PFOA and additional MIPs would need to be developed for other PFAS analytes. Selectivity of the MIP was demonstrated by spiking samples with other PFASs and interference was only seen in high concentrations. This methodology is promising due to its compatibility with soil samples, in addition to aqueous samples. With a relatively simple soil extraction method, this can still be used in the field. Benefits of this methodology include ease of miniaturization, real-time quantitative results, selectivity, and cost effectiveness.

3.3.4. Fluorine Nuclear Magnetic Resonance (NMR)

Gauthier and Mabury [54] evaluated fluorine (¹⁹F) nuclear magnetic resonance (NMR) to determine the quantitative and structural characteristics of 47 PFAS compounds. The study evaluated PFASs in environmental and biological media, including rainwater, lake water, wastewater effluent (solids and liquids), serum, and urine. ¹⁹F NMR offers a large spectral window spanning more than 300 ppm, which enhances the resolution compared to traditional proton NMR. A LOD for total PFASs was ~1 ppb, which was achieved through an array of smaller experiments that reduced noise while also providing a high-resolution spectrum, free from significant background contamination. ¹⁹F NMR is a non-destructive technique that requires minimal sample preparation and does not rely on prior knowledge of the chemical structure of analytes. This provides a high degree of specificity for fluorinated compounds, making this an ideal tool for broad-spectrum, non-targeted analysis in environmental chemistry. ¹⁹F NMR does not rely on matrix-matched calibration or mass-labeled standards, which is very advantageous for complex biological media including serum and urine. ¹⁹F NMR can use a multivariate analysis, which allows for enhanced signal-to-noise ratios, which enables the detection of low-concentration PFASs that might otherwise go undetected.

One challenge of ¹⁹F NMR is its sensitivity, which is highly dependent on the number of transients (i.e., individual signal acquisitions) during the experiment. To achieve the necessary sensitivity for low-concentration samples, longer acquisition times are often required, making the method potentially time-consuming and expensive. Additionally, the relationship between the number of transients and signal strength is not linear; so, after a certain point, increasing the number of transients results in diminishing returns in terms of signal intensity. A prolonged experimental time can lead to issues such a baseline noise and peak broadening, particularly if there are fluctuations in sample temperature or homogeneity. Other limitations of NMR are that it requires specialized equipment and expertise, which may limit the rapid and field transferability, and have higher analysis costs as compared to other emerging approaches. Nonetheless, despite some potential challenges, ¹⁹F NMR is a potentially valuable nonconventional technology which can offer insight into PFAS contamination that may not be captured by more targeted approaches including LC-MS/MS.

3.3.5. Metal–Organic Framework (MOF) Paired with Deep Eutectic Solvents

Han et al. [55] developed a novel method for detecting perfluoroalkyl iodides using a metal–organic framework combined with deep eutectic solvents for detection in oil. Perfluoroalkyl iodides are neutral, nonionic PFAS compounds that can be converted into perfluoroalkyl carboxylates. This approach uses a deep eutectic-solvent-based magnetic solid-phase extraction system, composed of pyridine analogues and glycolic acid, which is immobilized on a metal–organic framework surface (UiO-66-NH2). This, in turn, enhances PFAS enrichment through hydrogen bonding with the iodide ion in the PFAS molecules. Magnetic solid-phase extraction utilizes a magnetic sorbent that allows for the direct dispersion of the adsorbent, which is then retrieved using an external magnetic field. This technique offers several advantages, including shorter run times and a reduced need for hazardous solvents. The method’s efficiency is improved by the high contact surface area between the targets and the adsorbents. Selecting the appropriate modification reagents is crucial for achieving higher selectivity and improving adsorption/extraction capacity.

The method demonstrated a high sensitivity and recovery rate for five different PFAS compounds, perfluorobutyl iodide (PFBI), perfluorohexyl iodide (PFHxI), perfluorooctyl iodide (PFOI), perfluorodecyl iodide (PFDeI), and perfluorododecyl iodide (PFDoI), with LODs ranging from 0.00281 to 0.0343 ppb. This technique was first tested under “optimized” conditions with several media types, including air, water, sediment, and soil. Following optimization, the method attempted to detect the same five PFAS analytes but in 12 different oils; these analytes were detected in the oils at concentrations ranging from 0.2129 to 3.053 ppb. This approach represents a novel design that targets the function-oriented deep eutectic-solvent-based materials for PFAS analysis. While this approach provides good relative sensitivity and selectivity, it requires specialized equipment, sample preparation, extraction, and expertise for analysis. Therefore, the scalability to field--ready or rapid analysis remains a large source of uncertainty.

4. Discussion

4.1. Comparisons Among Emerging PFAS-Sensing Technologies

The majority of the emerging PFAS analytical methods reviewed in this study (n = 28) were focused on PFOS (71%) and PFOA (50%), with fewer studies investigating their efficacy for other analytes (e.g., PFBA, PFHxA, PFHxS) (i.e., [36,37,38,39,45,55]). The LODs for these emerging techniques were generally higher than those achievable using conventional methods. However, several of the optical, electrochemical, and nonconventional approaches were able to detect environmentally relevant concentrations of PFOS and PFOA in water and soil (Figure 1 and Figure 2).

Optical colorimetric methods (paper-based or porphyrin-based) are very promising for achieving low-cost, field-deployable, and rapid capabilities, but the sacrifice is the relatively high LOD ranging from 3000 to 10,000 ppb PFOS (see [25,26]). Optical fluorescence methods, such as cationic siloxane and erythrosine B-based fluorescence, have low upfront costs, as they primarily require a spectrophotometer, and provide moderate detection limits (e.g., 3.3 to ~1350 ppb PFOS). However, newer approaches, such as those using optical methods paired with conjugated polymers, allow for bright fluorescence at lower concentrations to show improved detection limits (i.e., 5–7 ppb LOD for PFOS; [34,35]) while maintaining similar costs and enhanced selectivity against common ions (e.g., Na, K, Cl). Advanced optical techniques, such as imprint-and-report fluorescence using dithiol dynamic combinatorial libraries (DCLs), distinguish mixtures of up to six PFASs with high specificity at detection limits of 10 ppb PFOS (20 nM) but require increased expertise and capability to perform. Other specialized approaches in genetically engineered bacterial biosensors, which demonstrate exceptional sensitivity (10 ppt for PFOA/PFOS) but remain in early development stages, with challenges related to field implications. Lastly, highly sensitive technologies like surface-enhanced Raman scattering (SERS) provide low detection limits in the ng/L range. For example, the SERS optical method developed by McDonnel and co-authors [41] had reported LODs of 0.0005 and 0.07 ppb for PFOS and PFOA, respectively. Additionally, given the relatively stable shelf life of 9 months for the sensors, and field-portable capabilities, the SERS sensor has potential for supporting field-based efforts for delineating PFOA in water samples. Overall, methods with strong selectivity, such as imprint-and-report fluorescence and aptamer-based sensors and surface-enhanced Raman scattering, promising aspects of this technology for application to field investigations are the low limit of detections, making its use potentially feasible, and selectivity in complex mixtures.

Many of the electrochemical-based methods showed promising LODs for PFOS, with detection limits ranging from 0.0017 to 20 ppb (n = 6 studies; Figure 2 and Figure 3). There are advancements in the use of platinum electrochemical sensors paired with selective perfluorinated anion exchange ionomers coatings which offer a balance of moderate sensitivity (LOD 6.21 ppb for PFOA; [43]) with the potential future application of a deployable in situ sensing method. Yet, current drawbacks of this method are limitations regarding a need for additional redox active probes and chemical reagents. For higher specificity, ion transfer electroanalytical methods using microinterface arrays [46] achieve detection limits as low as 0.497 ppb for PFOA while maintaining selectivity against chloride and sulfate ions. Similarly, nano-electrochemical approaches with silver nanoparticles achieve high sensitivity (0.04 ppb PFOS; [50]) with a medium cost, highlighting their potential in highly sensitive applications. Techniques such as molecularly imprinted polymers (MIPs) demonstrate exceptional sensitivity, detecting PFOS at 0.0075 ppb [44] and 0.0017 ppb PFOS [47]. For example, Clark et al. [47] applied a unique approach of using oxygen in river water as a mediator to calibrate PFOS concentrations on a MIP-modified carbon substrate and achieved low LODs in the presence of matrix interferences (Cl, humic acid). Lu et al. [44] detected PFOS in tap water at a LOD of 0.0075 ppb with an ultra-sensitive voltametric sensor that used a glassy carbon electrode (GCE) with a thin coating of gold nanostar (AuNS) and electropolymerized MIP.

Barriers for both analytical detection methods are the selectivity of the mechanism, both in terms of the differentiation for specific PFAS compounds in the presence of other PFASs (e.g., PFOA versus PFOS) and selectivity in terms of matrix interferences (e.g., organic matter, salts, etc.). Several optical- and electrochemical-based analytical methods reported PFAS selectivity in the presence of other PFAS mixtures (Table 2 and Table 3). In a notable example, Harrison and Waters [36] successfully distinguished among mixtures of six PFASs (i.e., PFOS, PFOA, PFHxA, GenX, PfPeA, PFHpA) using an optical “imprint-and-report fluorescence” approach. However, for many of the emerging technologies, there remains much uncertainty regarding the selectivity of PFASs as the complexity of the matrix increases from laboratory to field evaluations. Common matrix interferences evaluated included cations and anions (e.g., Na, K, Ca, Cl; e.g., [28,32]) or metals (e.g., Cu, Mn, Fe, Cr, Mg, Zn; [32]). However, fewer studies evaluated method selectivity for common organic matrix interferences. For example, many of the reported approaches used laboratory “spiked” water (Table 2 and Table 3). In one of the few studies that used a complex matrix, Pierpaoli et al. [48] demonstrated that molecularly imprinted polymers (MIPs) can be designed with molecular cavities to target specific PFAS molecules, which can improve detection among complex mixtures (e.g., wastewater, landfill leachate).

PFASs are pervasive at low concentrations in the environment globally, with many studies reporting ng/L detections found in ambient concentrations of rainwater, soil, sediments, biota, freshwater, estuarine, and marine sources [6,32,73,74,75]. Reported PFOS ambient concentrations range from non-detectable to 0.138 ppb in surface waters [71], non-detectable to 20 ppb in groundwater [76], <0.04 to 0.07 ppb in drinking water [77], 0.03 to 1.95 ppb in soils [78], 0.005 to 1.13 ppb in sediment [79], and 6.6 ppb in fish tissue [80]. In the context of emerging regulatory criteria and screening values for PFASs, Ruffle et al. [81] conducted a recent review of surface water quality criteria (SWQC) or screening values for PFASs by region (e.g., Australia, Canada, European Union (EU), and USA) and found that numerical criteria for PFOS span approximately five orders of magnitude from 0.0047 to 600 ng/L. The broad range of regionally specific criteria highlight the existing uncertainties and evolving science associated with the exposure, fate, and toxicity associated with PFASs [81]. To better inform our understanding of exposure (and risk) associated with PFASs, there have been recommendations for the development of more comprehensive monitoring programs (e.g., fate and transport predictive models, bioaccumulation; [22]). The challenge of ever-decreasing promulgated regulatory criteria and the pervasiveness of PFASs in the environment drive an urgent need for a broad array of analytical testing methodologies to support expanded environmental monitoring and risk assessment programs.

Figure 1. Reported PFOA detection limits as compared to ambient background concentrations and conventional analytical detection limits. Note: LC-MS/MS method detection limit (MDL) for PFOA based on USEPA method 537.1 (PFOA = 0.00053 ppb (µg/L)) [82]; USEPA legally enforceable maximum contaminant levels (MCLs) for PFOA = 0.004 ppb (µg/L), ¹ Reported ambient concentration ranges for PFOA for surface and groundwater based on the work of Javis et al. [71] and Johnson et al. [76]; ² Reported ambient PFOA concentration ranges for soil and sediment based on the work of Rankin et al. [78] and Vedagiri et al. [79]. Optical references: [26,27,30,31,34,35,36,37,38,41]; electrochemical references: [43,45,46]; nonconventional references: [52,53].

Figure 2. Reported PFOS detection limits as compared to ambient background concentrations and conventional analytical detection limits. Note: LC-MS/MS method detection limit (MDL) for PFOS based on USEPA method 537.1 for PFOS = 0.004 ppb (µg/L) [82]; USEPA legally enforceable maximum contaminant levels (MCLs) for PFOS = 0.004 ppb (µg/L), ¹ Reported ambient concentration ranges for PFOA for surface and groundwater based on the work of Javis et al. [71] and Johnson et al. [76]; ² Reported ambient PFOA concentration ranges for soil and sediment based on the work of Rankin et al. [78] and Vedagiri et al. [79]. Optical references: [25,27,28,30,32,33,34,35,36,37,41]; electrochemical references: [44,45,47,48,50]; nonconventional references: [69].

4.2. Application to Risk-Based Assessments

There are several potential applications for these emerging analytical tools within the risk assessment and risk management process. Initial site investigations rely on results of sampling and analysis to identify potential PFAS sources, spatially delineate contamination, and inform likely exposure routes (e.g., fate and transport; [17,18,83]). At this phase, there are often practical challenges of planning and prioritizing sample locations across environmental media. The initial step is identifying the potential receptors (e.g., humans, aquatic organisms, wildlife, plants), defining the nature and extent of potential exposure and developing a conceptual site model to describe the sources, exposure pathways, and potential receptors of concern. In the analysis phase, data on PFAS concentrations in environmental media (e.g., soil, water, sediment) are collected, and an exposure assessment is conducted to determine the extent to which ecological receptors may be exposed to PFASs. During this phase of the evaluation, there are opportunities to use non-standardized, cost-effective, rapid, onsite methods for initial (screening level) characterization to optimize collection of samples for more detailed (and expensive) analyses (e.g., USEPA 1633 [21]) and reduce uncertainty in site management decisions. These tools could be particularly helpful in larger, more complex, and high-cost risk assessments requiring large numbers of samples over space and time (Figure 3).

In a typical ecological risk assessment framework, screening level assessments are aimed at developing readily available data and information to decrease key uncertainties and better inform decisions. Emerging sensing technologies could be coupled with standardized approaches to both inform when and where to sample in a tiered process (Figure 3). Early application of these technologies would have a secondary benefit of enabling the development, validation, and commercialization of these emerging techniques.

In the context of the risk characterization phase, there are some proponents for adopting more holistic approach to informing risk of PFASs based on classes or groups of compounds with similar chemical structures, toxicological profiles, and mechanisms of action (e.g., the relative potency factor approach, mixture-based approaches; [84]). As these approaches evolve, there maybe opportunities for application of technologies that provide more holistic and correlative measures of PFAS exposure [22]. However, to have utility in supporting risk assessments, these methods would first need to be validated against effects data to ensure that what is being measured is correlated to biological response and to quantify uncertainty. Finally, as the assessment process moves into the risk management or remedial action phase, there are exciting opportunities to integrate these emerging sensing opportunities in tandem with treatment technologies to enable the optimization of treatment processes and facilitate adaptive management [11].

Addressing future challenges posed by PFAS contamination, the integration of traditional analytical methods with emerging detection technologies potentially offers a balanced approach to improving site investigation through informed sample design by reducing uncertainty and optimizing resource allocation (i.e., sample collection and analysis quantity). While conventional methods, such as LC-MS/MS, offer lower limits of detection and higher accuracy, they are time-intensive and costly, limiting their application for more spatially and temporally comprehensive environmental monitoring. Development of sensing approaches may offer a cost-effective alternative to offset the limitations of current analytical tools [85]. Coupling of these more novel, emergent methods with traditional analytical tools has the potential to strike a balance between cost, time and uncertainty to facilitate more comprehensive and efficient assessments of PFASs over a range of environmental settings.

Figure 3. Conceptual integration of emerging technologies within risk assessment processes. Examples of ecological risk analysis framework (A), tiered framework for screening chemical and advanced materials (B); and proposed hybrid relationship between emerging and standardized approaches for PFAS analysis for risk-based processes (C). Note: panel (A) based on the USEPA ecological risk framework [17,18]; panel (B) based on Moore et al.’s [86] tiered process flowchart to improve assessment, monitoring and adaptive management of emerging contaminants.

4.3. Knowledge Gaps and Recommendations

Despite significant advancements in recent years, several knowledge gaps remain in the development of emerging sensing methods for PFASs. First, the selectivity of these sensors across complex environmental matrices requires further improvement and evaluation, particularly in distinguishing PFASs from other anions, cations, and organic contaminants commonly present in environmental samples. Second, while many sensing methods demonstrate low detection limits (e.g., ng/L) in controlled laboratory settings, their field applicability remains largely unexplored, including challenges related to sensor stability, reproducibility, and performance in diverse and dynamic conditions (e.g., varying pH, temperature, and salinity). Finally, scalability and standardization are critical bottlenecks, as most of these methods are still in the early stages of development and lack validation against standardized techniques (i.e., EPA method 537.1 and 1633; [19,21]). This includes uncertainties regarding the cost effectiveness, reusability, and simplicity (ease-of-use) of sensors to support widespread deployment in environmental monitoring and risk assessment efforts. Addressing these gaps is essential to advance the practical application of optical and electrochemical sensing methods for PFAS detection in diverse environmental contexts.

Despite these uncertainties, it should be noted that there remains tremendous practical value in qualitative and semi-quantitative environmental monitoring data and even data generated using non-validated tools to use in conjunction with validated chromatography-based methods (i.e., a hybrid analytical testing approach). If these methods can enable more consistent and frequent monitoring, this would decrease uncertainty regarding exposures over time and space. These data could support delineating spatial or temporal trends at sites and “trigger” decisions for more intensive sampling using validated methods used for regulatory decisions. Additionally, qualitative or semi-quantitative sensing technologies could integrate with the rapidly expanding suite of remedial PFAS destruction technologies in development to inform treatment performance in terms of relative rates and extents of removal (relative changes pre- and post-treatment).

5. Conclusions

With the greater understanding of the potential human and ecological risk of PFASs, there has been a dramatic increase in the development of systematic monitoring programs, site investigations, risk assessments, and development of treatment technologies. Commensurate with these efforts is the need to evolve the current suite of analytical tools to support the greater volume and diversity of required data needs. There has been substantial progress in developing conventional techniques (e.g., LC-MS/MS) to achieve low (ng/L) detection levels. However, there is a concomitant need to develop non-traditional PFAS sensors or approaches to hybridize the analytical testing options. This review highlights several examples of promising technologies that could support risk assessment and risk management decision making. Key opportunities exist for supporting screening-level assessment whereby detection capabilities of in situ sensors can qualitatively delineate sites with PFAS contamination. Both optical and electrochemical based methods are promising as they have demonstrated environmentally relevant detection limits, while enabling cost-effective, rapid and onsite results. We proposed a conceptual approach whereby a tiered evaluation process aided by these emerging techniques could ultimately save time, money, resources in the extensive environmental monitoring plans and programs on the horizon.

Author Contributions

A.M.: Conceptualization, Supervision, Writing—Original Draft, Methodology, Investigation, Formal Analysis. A.K.: Writing—Review and Editing, Writing—Original Draft, Methodology. P.K.: Writing—Review and Editing, Writing—Original Draft, Visualization, Methodology, Investigation, Formal Analysis. A.L.: Writing—Review and Editing, Formal Analysis. A.C.-E.: Writing—Review and Editing, Methodology, Investigation, Formal Analysis. D.M.: Writing—Review and Editing, Methodology, Funding Acquisition. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the PFAS risk reduction applied research program (PE 622144DI7).

Data Availability Statement

The data that support the findings of this study are available from the corresponding author upon request.

Acknowledgments

The authors are grateful for the support of William Gardiner (USACE Seattle District), Marlowe Laubach (USACE Seattle District), and Austin Scircle (USACE ERDC) for their assistance reviewing earlier versions of this manuscript.

Conflicts of Interest

The authors declare no conflicts of interest. The views and opinions expressed in this article are those of the individual authors and not those of the US Army Corps of Engineers, US Army Engineer Research and Development Center, or other sponsor organizations.

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Table 1. Summary of emerging PFAS detection methods, descriptions of approach, and example applications.

Detection Mechanism	Description	Technique	Materials/Approach	Reference
Optical-based methods	These methods rely on the interaction of PFAS molecules with specialized substances or surfaces that enhance optical signals and quantify target analyte through either a chemical reaction or a mechanism.	Colorimetric	Atomic copper-carbon nitride; paper-based devices using methylene green; porphyrin-based detector; optical emission of biphasic oil–water droplets	[25,26,27]
		Fluorescence	Bacterial gene expression; fluorescein-modified aptamer; cationic siloxane and erythosine B; perylene dimide; conjugated polymer; imprint-and-report; ligand-binding; aptamer-based; dual DNA and lysozyme fiber biosensors	[28,29,30,31,32,33,34,35,36,37,38,39]
		Surface-enhanced Raman spectroscopy (SERS)	Nanoparticle–nanorod sandwich structure; aerosol jet printing silver nanoparticles and graphene inks on Kapton films	[40,41]
Electrochemical-based	Emerging electrochemical-based analytical approaches for PFAS detection focus on exploiting interactions of PFAS molecules with modified electrodes. PFASs are generally electrochemically inert and do not transfer electrons in normal environmental pH conditions. Therefore, novel approaches have been developed to block signals, which results in a quantifiable reaction proportional to the concentration.	Electrochemical impedance spectroscopy (EIS)	Bare platinum electrode; modified electrode with selective perfluorinated anion exchange ionomer	[42,43]
		Cyclic voltammetry (CV), differential pulse voltammetry (DPV)	Gold nanostars (AuNS) on glassy carbon electrodes (GCE); micropipette-based interfaces between two immiscible electrolyte solutions	[44,45,46]
		Molecularly imprinted polymers (MIPs) paired with EIS or DPV	MIP-based carbon electrode	[47,48]
		Self-assembled monolayers (SAMs) paired with voltammetry	SAMs using bonding between gold substrates and thiol groups	[49]
		Single particle collision electrochemistry (SPCE)	nano-electrochemistry with silver nanoparticles	[50]
Nonconventional	Novel technologies applying methods of detection outside of electrochemical or optical methods of detection.	Hydrogen swelling	Microgel structure change	[51]
		Flow rate analysis	Lateral flow immunoassay (LFIA)	[52]
		Thermal detection	MIP paired with heat transfer method (HTM)	[53]
		Nuclear magnetic resonance (NMR)	Fluorine NMR spectroscopy	[54]
		Metal–organic framework (MOF) paired with deep eutectic solvents	Deep eutectic solvents paired with magnetic solid-phase extraction system immobilized on a MOF	[55]

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McQueen, A.; Kimble, A.; Krupa, P.; Longwell, A.; Calomeni-Eck, A.; Moore, D. Review of Emerging and Nonconventional Analytical Techniques for Per- and Polyfluoroalkyl Substances (PFAS): Application for Risk Assessment. Water 2025, 17, 303. https://doi.org/10.3390/w17030303

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McQueen A, Kimble A, Krupa P, Longwell A, Calomeni-Eck A, Moore D. Review of Emerging and Nonconventional Analytical Techniques for Per- and Polyfluoroalkyl Substances (PFAS): Application for Risk Assessment. Water. 2025; 17(3):303. https://doi.org/10.3390/w17030303

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McQueen, Andrew, Ashley Kimble, Paige Krupa, Anna Longwell, Alyssa Calomeni-Eck, and David Moore. 2025. "Review of Emerging and Nonconventional Analytical Techniques for Per- and Polyfluoroalkyl Substances (PFAS): Application for Risk Assessment" Water 17, no. 3: 303. https://doi.org/10.3390/w17030303

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McQueen, A., Kimble, A., Krupa, P., Longwell, A., Calomeni-Eck, A., & Moore, D. (2025). Review of Emerging and Nonconventional Analytical Techniques for Per- and Polyfluoroalkyl Substances (PFAS): Application for Risk Assessment. Water, 17(3), 303. https://doi.org/10.3390/w17030303

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