Treatment of per- and polyfluoroalkyl substances in landfill leachate: status, chemistry and prospects

Zongsu Wei ab, Tianyuan Xu ac and Dongye Zhao *a
aEnvironmental Engineering Program, Department of Civil Engineering, Auburn University, Auburn, AL 36849, USA. E-mail: zhaodon@auburn.edu
bCentre for Water Technology (WATEC), Department of Engineering, Aarhus University, Hangøvej 2, DK-8200 Aarhus N, Denmark
cKey Laboratory of Pesticide & Chemical Biology of Ministry of Education, Institute of Environmental & Applied Chemistry, College of Chemistry, Central China Normal University, Wuhan 430079, People's Republic of China

Received 27th July 2019 , Accepted 3rd October 2019

First published on 4th October 2019


Per- and polyfluoroalkyl substances (PFAS) have been widely detected in municipal landfills due to the massive uses of PFAS in everyday consumer products. PFAS in landfill leachate can contaminate the neighbouring soil and groundwater and pose serious health concerns to human and ecosystems. Yet, information is lacking on the distribution, transformation and fate of various PFAS in landfills, and on the treatment technologies to remove or degrade PFAS in landfill leachate. As the relevant regulations are rapidly evolving, the research and technology development have been gaining momentum in recent years. Here, we present a comprehensive review of the state of the science on the occurrence and treatment technologies of PFAS in landfills. Specifically, this review aims to overview the following aspects: 1) the occurrence and transformation of PFAS under typical landfill environmental conditions, 2) the chemistry and state-of-art technologies for treating PFAS in landfill leachate, including adsorption, biodegradation, photo-degradation, and membrane processes, and 3) the key knowledge gaps and future research/technology needs for controlling PFAS from landfills.



Water impact

Municipal landfills are a major source of per- and polyfluoroalkyl substances (PFAS). This work overviews the state of the science on the occurrence, chemistry and treatment technologies for PFAS in landfill leachate, and identifies some knowledge gaps and future research needs. The information is valuable for improving our understanding of PFAS in aquatic systems and for guiding development of cost-effective treatment technologies.

1. Introduction

Per- and polyfluoroalkyl substances (PFAS) have been intensively manufactured since the 1940s for industrial and residential applications, including fluoropolymeric surfactants, aqueous film-forming foams, metal plating, textile, household products, etc.1 Because of their unique molecular structure and strong C–F bond energy (536 kJ mol−1), PFAS show some unique properties (e.g., both hydrophobic and lipophobic) and extremely high chemical stability.2 There are more than 3000 PFAS in the global market, most of which are overlooked with unknown human exposure, environmental fate, and toxicological information.3 From 2009 to 2017, 455 PFAS in the forms of anions, cations, zwitterions, and neutral molecules were detected in the aquatic environment.4 Perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) are two of the most detected PFAS in aquatic systems and drinking water worldwide.5 PFAS are also ubiquitously present in sediments, soil, activated sludge, and landfill, raising a global concern due to their toxicity, persistence and bioaccumulative potential.6,7

Due to the ubiquitous presence of PFAS, humans can be exposed to them through multiple routes. In addition to drinking water, PFAS can reach humans through food (e.g., fish), consumer products, and even indoor air/dust.8 PFAS, notably PFOA and PFOS, were detected at up to 200 ng L−1 in human tissue or blood serum.2 PFOA and PFOS accumulated in the livers of marine mammals were >1200 ng g−1 w.w. in the Arctic area.9 Exposure to PFAS and their accumulation in the human body have been associated with various adverse health effects such as immunotoxicity, neurotoxicity, reproductive and developmental deficits, hepatomegaly and hepatic peroxisome proliferation, and pancreatic (acinar cell) tumours.1,10

In response to the rising concern about PFAS, the U.S. EPA has urged the industry (e.g., 3 M company) to reduce the production of PFOS and related compounds in the early 2000s.11 In Europe, PFOS and PFOA were both listed on the Stockholm Convention on Persistent Organic Pollutants (POPs) in 2009.12 At the beginning of the 2010s, the EPA further worked with eight leading chemical companies to reduce PFOA by 95% in a PFOA Stewardship Program.13 In 2012, EPA added six PFAS on the list of The Third Unregulated Contaminant Monitoring Rule (UCMR 3).13 The monitoring results from UCMR 3 led to the regulatory action on May 19, 2016, when the EPA set a drinking water health advisory level of 70 ng L−1 for combined PFOA and PFOS concentrations.14 Meanwhile, several states in the US have promulgated drinking water standards for PFAS, as the EPA is moving aggressively toward regulating these chemicals.

In the life cycle of PFAS, landfills are considered as the final stage, where PFAS are released from products and wastes derived from residential, commercial and industrial sources.15 However, PFAS can leach out from landfills and affect surrounding soil and water systems. PFAS have been detected in landfill leachate, including perfluoalkyl acids (PFAAs), fluorotelomer polymers (FTPs), perfluoroalkyl sulfonamide derivatives, and polyfluoroalkyl phosphate esters (PAPs).16 In particular, PFAAs, such as perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkyl sulfonic acids (PFSAs), have been the most frequently detected PFAS in landfill leachate in the ng L−1 to mg L−1 range.5 In general, short-chain PFAAs (C4–C7) are predominant over long-chain PFAAs (>C7), since the former are more soluble and more persistent in water and thus more prone to leaching out from solid wastes.17 In the U.S., PFCAs in landfill leachate were reported to range from 10 to 8900 ng L−1, while PFSAs were found to be from 50 to 3200 ng L−1.18,19 The broad concentration range was attributed to the heterogeneity of solid wastes, as well as the landfill age, climate conditions, and biochemical activities.

Treating PFAS in landfill leachate is very challenging because of the presence of highly complex matrices. Typically, landfill leachate is treated ex situ through conventional activated sludge processes in wastewater treatment plants (WWTPs). However, biological processes were ineffective for degrading PFAS.20–22 While membrane bioreactors (MBRs) have been commonly employed in treating landfill leachate, they are inefficient to remove PFAS from landfill leachate.23 Therefore, additional treatment methods are needed to deal with PFAS, including adsorption of PFAAs by activated carbon (AC)24 and membrane processes, i.e., nanofiltration (NF) and reverse osmosis (RO).25 Although ozone- and ultraviolet (UV)-based advanced oxidation processes (AOPs) have been pursued as options for leachate treatment, their effectiveness is severely compromised due to the high matrix strength of leachate, particularly the high total organic carbon (TOC) contents.26 Combined physico-chemical and biological methods have been investigated to further polish treated leachate.27 For instance, Yan et al.28 studied combined biological oxidation and membrane filtration to treat PFAAs in the leachate from five municipal landfills in China. While NF or RO yields a high PFAS rejection rate, the concentrates containing PFAAs from the membrane process still need further treatments.

While many treatment technologies have been explored in recent years to treat PFAS in drinking water systems, much less work has been reported pertaining to treating PFAS in high strength waste streams (e.g., landfill leachate, membrane concentrates, and ion exchange brine). While landfill leachate has become a major source of PFAS, there is an urgent need to develop cost-effective technologies to treat PFAS in landfill leachate. To this end, a systematic review of the state of the technology is highly desired. Specifically, this review aims to: 1) briefly overview some relevant fundamentals on PFAS and physicochemical characteristics of landfill leachate, 2) critically review the occurrence, impacts, and transformation of PFAS in landfill leachate, as well as current treatment technologies, such as adsorption, reduction/oxidation, membrane filtration, and biological processes, and 3) discuss the related engineering challenges and future research needs for cost-effective PFAS removal from landfill leachate.

2. Fundamental properties of PFAS

As shown in Table 1, PFAS are classified into perfluoroalkyl and polyfluoroalkyl groups based on the saturation level of fluorine on the carbon backbone.29,30 Perfluoroalkyl substances feature fully fluorinated (perfluoro-) alkane (carbon-chain) chemical structures, with a carbon chain tail and a functional head group on the other end. For instance, perfluoroalkyl carboxylates (PFCAs) and perfluoroalkane sulfonates (PFSAs), the most extensively produced PFAS, have carboxylic and sulfonic functional groups, respectively. Polyfluoroalkyl substances are not fully fluorinated, with a non-fluorine atom (typically hydrogen or oxygen) attached to carbon atoms but with at least two of the remaining carbon atoms fully fluorinated. Compared to the C–F bond, the C–H bond (337.2 kJ mol−1) in polyfluoroalkyl molecules is a “weaker” link that is susceptible to oxidation or reduction cleavage. The C–C bond is also weakened when an F atom is substituted by an H atom, e.g., the dissociation energy of CF3–CF3 (406 kJ mol−1) is larger than that of CH2F–CH2F (368 kJ mol−1).
Table 1 Classes of common PFAS and their molecular makings
Family Class Group Chemical structure
Perfluorinated Perfluoroalkyl acids (PFAAs) Perfluoroalkyl carboxylates (PFCAs) F3C(CF2)nCOO
n = 2, PFBA; n = 3, PFPeA; n = 4, PFHxA; n = 5, PFHpA; n = 6, PFOA; n = 7, PFNA; n = 8, PFDA; n = 9, PFUnDA; n = 10, PFDoDA; n = 12, PFTDA
Perfluoroalkane sulfonates (PFSAs) F3C(CF2)nSO3
n = 3, PFBS; n = 5, PFHxS; n = 7, PFOS; n = 9, PFDS
Perfluoroalkane sulphonamides (FASAs) Perfluorooctane sulfonamide CnH2n+1–SO2NH2
N-Alkyl perfluoroalkane sulfonamides CnF2n+1–SO2N(R′)H
Where R′ = CmH2m+1 (m = 0, 1, 2,4)
Polyfluorinated Fluorotelomer substances n[thin space (1/6-em)]:[thin space (1/6-em)]2 Fluorotelomer alcohols (n[thin space (1/6-em)]:[thin space (1/6-em)]2 FTOHs) F3C(CF2CF2)n–CH2CH2OH
n[thin space (1/6-em)]:[thin space (1/6-em)]2 Fluorotelomer sulfonic acids (n[thin space (1/6-em)]:[thin space (1/6-em)]2 FTSAs) F3C(CF2CF2)n–CH2CH2SO3H
Fluorotelomer carboxylic acids (FTCAs) F3C(CF2CF2)n–CH2COOH (n[thin space (1/6-em)]:[thin space (1/6-em)]2)
CnF2n+1–CH2CH2COOH (n[thin space (1/6-em)]:[thin space (1/6-em)]3)
Perfluoroalkane sulfonamido substances Perfluoroalkane sulfonamido ethanols and N-alkyl perfluoroalkane sulfonamido ethanols CnF2n+1–SO2N(R′)CH2CH2OH
Where R′ = CmH2m+1 (m = 0, 1, 2, 4)
Perfluoroalkane sulfonamido acetic acids and N-alkyl perfluoroalkane sulfonamido acetic acids CnF2n+1–SO2N(R′)CH2COOH
Where R′ = CmH2m+1 (m = 0, 1, 2,4)


Polymeric PFAS are a group of polymerized macromolecules with repeating smaller molecules (i.e., monomers), including fluoropolymers, polymeric perfluoropolyethers, and branched fluorinated polymers (Table 1). Based on the number of carbon atoms, PFAS are categorized into long-chain and short-chain PFAS.31 For example, PFCAs with eight or more carbons are considered long-chain and those with seven or fewer carbons are considered as short-chain; PFSAs with six or more carbons are categorized as long-chain and those with five or fewer carbons are categorized as short-chain.

At the molecular level, the fluorine atom has a very high electronegativity and small size, resulting in a very strong C–F bond (dissociation energy = 536 kJ mol−1),32 which accounts for the high resistance of PFAS to degradation by oxidative/reductive processes. Meanwhile, fluorine has a low polarizability that leads to weak intermolecular interactions (i.e., van der Waals interactions and hydrogen bonding). The unique properties of fluorine endow many PFAS with their hydrophobic and lipophobic properties and extremely high thermal and chemical stability. Many PFAS are strong acids due to the presence of acid functional groups. For example, the pKa value for common PFAAs was reported to range from −0.5 to 3.8.33 Therefore, PFAS are usually present in fully dissociated anionic forms at environmentally relevant pH. It should be noted that PFAS in protonated and anionic forms vary greatly in physical and chemical properties. For example, anionic PFOA is highly water soluble, while the undissociated acid has a very low solubility.34 It should be noted that the octanol/water partition coefficient (log[thin space (1/6-em)]KOW) value usually refers to the acid form, and increases with the carbon-chain length or molecular weight.34 Therefore, these log[thin space (1/6-em)]KOW values are not relevant to anionic forms.

Understanding the physio-chemical properties of PFAS is critical to understand their fate and transport behaviour in the environment. However, reliable measurements of the physical and chemical properties of PFAS are difficult (e.g., vapor pressure, Henry's law constant, and octanol/water partition coefficient), and there have been large variations in published data due to difficult conditions and methods used.30 In addition, our knowledge on the chemistry of PFAS, especially under landfill leachate conditions, has been very limited.

3. PFAS in landfill leachate

3.1. Characteristics of landfill leachate

Landfills are the end-point for disposal of municipal, commercial and industrial wastes, which undergo various physical, chemical, and microbial processes to produce a variety of by-products.35–38 When rainwater percolates through the wastes, the produced pollutants accumulate in the water-based solution, i.e., the leachate. Typically, four groups of pollutants may be produced: dissolved organic matter (DOM), e.g., fulvic-like and humic-like compounds, xenobiotic organic compounds (XOCs), heavy metals, and inorganic salts.37Table 2 summarizes the compositions of a typical landfill leachate. The concentrations may vary broadly due to the different landfill sites, ages, weather conditions, and waste sources. There are several unique features for landfill leachate: 1) high organic content (e.g., TOC may reach as high as 29.0 g L−1, 2) high salt concentration (62.4 g L−1), which is several orders of magnitude higher than that in groundwater, and 3) a variety of XOCs or heavy metals.39
Table 2 Characteristics of a typical municipal landfill leachate (adapted from Kjeldsen et al.)37
Parameter Range
Note: Values are from ref. 37. The inorganic macrocomponents include ammonium (NH4+), sodium (Na+), potassium (K+), calcium (Ca2+), iron (Fe2+), magnesium (Mg2+), manganese (Mn2+), chloride (Cl), hydrogen carbonate (HCO3), total phosphorous (P), sulphate (SO42−), and silica (SiO2). Heavy metals include cadmium (Cd2+), cobalt (Co2+), copper (Cu2+), lead (Pb2+), mercury (Hg2+), nickel (Ni2+), zinc (Zn2+), chromium (Cr3+), and arsenic (As).
pH 4.5–9
Conductivity (μS cm−1) 2500–35[thin space (1/6-em)]000
Total solids 2000–60[thin space (1/6-em)]000
Total organic carbon (TOC) 30–29[thin space (1/6-em)]000
Biological oxygen demand (BOD5) 20–57[thin space (1/6-em)]000
Chemical oxygen demand (COD) 140–152[thin space (1/6-em)]000
BOD5/COD (ratio) 0.02–0.80
Organic nitrogen (mg L−1) 14–2500
Inorganic macrocomponents (mg L−1) 985.13–62[thin space (1/6-em)]363
Heavy metals (mg L−1) 0.08115–1032.56


The XOCs in landfill leachate may span across aromatic hydrocarbons, phenolic compounds, halogenated hydrocarbons, pesticides, phthalates, and pharmaceuticals.40 Heavy metals (0.08115–1032.56 mg L−1) such as cadmium, chromium, copper, lead, mercury, and zinc could reach problematic levels in the leachate, particularly when complexed with high level organic matter. Other inorganic macrocomponents (985.13–62[thin space (1/6-em)]363 mg L−1) include iron, manganese, calcium, sodium, potassium, ammonia, etc., among which ammonium nitrogen was detected at a much higher concentration in landfill leachate.

The composition of the leachate is usually an indicator of the waste types and the processes occurring within the anaerobic environment. Both organic and inorganic pollutants maintain high level concentrations in the leachate and their potential release to surface and groundwater would have a major impact on the safety of the water source and ecosystem, especially for those landfill sites without engineered liners and collection systems. Furthermore, long-term evaluation of the leachate composition is critical for implementing practical treatment processes for treating PFAS in landfill leachate because the leachate matrix can strongly interfere with treatment performances.

3.2. Source and distribution of PFAS in landfill leachate

Both legacy and emerging PFAS, in particular PFOA and PFOS, have been widely detected in landfill leachate due to disposal of consumer products from domestic (e.g., food contact paper, microwave popcorn bags, carpets, outdoor clothing, etc.) and industrial sources (textiles, surfactants, insecticides, aqueous film-forming foams, etc.) containing PFAS.41,42 PFAS in landfills were found to be mainly distributed in the aqueous phase and in a wide range of concentrations. For example, the concentration of PFOA ranged from 0.15 to 9.2 μg L−1 in 13 U.S. landfills.19,43 In accordance with many contaminated sites, PFCAs and PFSAs are the most frequently detected PFAS in landfills worldwide (Table 3). The carbon chain length varied from C3 to C14 and in concentrations ranging from ng L−1 to μg L−1. In addition, fluorotelomer-based compounds (e.g., n[thin space (1/6-em)]:[thin space (1/6-em)]2 and n[thin space (1/6-em)]:[thin space (1/6-em)]3 fluorotelomer carboxylic acids and n[thin space (1/6-em)]:[thin space (1/6-em)]2 fluorotelomer sulfonates), perfluoroalkyl sulfonamide derivatives (e.g., ethyl-perfluorooctane sulfonamidoethanol (EtFOSE)), and polyfluoroalkyl phosphate esters (e.g., di-substituted fluorotelomer phosphate esters and EtFOSE-based polyfluoroalkyl phosphate diesters) have also been detected in landfill leachate.16
Table 3 The global distribution of PFAS in landfill leachate. Unit is ng L−1
Compound Region
North America Europe Australia China
Note: N.R. = not reported; <LOD = less than limit of detection. a Values are from ref. 19, 41–43, 45, 53–55. b Values are from ref. 24, 39, 44–48 and 56. c Values are from ref. 45 and 50. d Values are from ref. 45 and 51.
PFBS (C4) 28–3200 <0.39–1356 <840 1600–41[thin space (1/6-em)]600
PFPeS (C5) N.R. N.R. N.R. N.R.
PFHxS (C6) 45–1100 <0.20–1800 <1900 <479
PFHpS (C7) N.R. N.R. N.R. N.R.
PFOS (C8) <9.5–4400 0.01–1500 <1100 1150–6020
PFDcS (C10) N.R. <1–0.28 <3 N.R.
FTCA 10–15[thin space (1/6-em)]000 N.R. N.R. N.R.
PFPrA (C3) N.R. N.R. N.R. 638–10[thin space (1/6-em)]000
PFBA (C4) 69–660 <3.36–2968 <1600 1100–9270
PFPeA (C5) 54–3200 <829 N.D. 609–6530
PFHxA (C6) 190–8900 <2900 12–5700 146–4430
PFHpA (C7) 62–3100 <600 2.2–3500 75.4–5830
PFOA (C8) 42–5000 < 4200 19–2100 281–214[thin space (1/6-em)]000
PFNA (C9) 11–450 < 680 <89 <381
PFDA (C10) 0.3–1100 <410 <57 <18.8
PFUnA (C11) <120 <430 <18 N.R.
PFDoA (C12) <1.4 –16 <25 <28 N.R.
PFOSA 3.4–220 <0.15–14.0 N.R. N.R.
FTSA 0.3–300 N.R. N.R. N.R.
Sum PFSAs <82.5–8700 <1.6–4656 <3843 <48[thin space (1/6-em)]099
PFCAs <433.1–22[thin space (1/6-em)]766 <3.51–13[thin space (1/6-em)]056 33.2–13[thin space (1/6-em)]092 2849.4–250[thin space (1/6-em)]459.8


As shown in Table 3, PFCAs are in general more abundant than PFSAs, whilst short-chain PFAS (<C8 for PFCAs and <C6 for PFSAs) are observed to be more prevalent than their long-chain homologues. For example, Busch et al.24 reported that short-chain PFAS, perfluorobutanoic acid (PFBA) (∼27% in mass) and perfluorobutane sulfonate (PFBS) (∼24% in mass), were dominant in the untreated leachate from 22 landfill sites in Germany. This is expected because: 1) short-chain PFAS are more hydrophilic and more mobile, and thus are preferentially released from the wastes, 2) shorter-chain PFAS are often intermediate products from the breakdown of longer-chain PFAS or other precursors such as FTPs, and 3) the industrial production of PFAS has shifted from PFOA and PFOS to shorter-chain compounds since 2000.

In U.S. and Canada, PFOA was detected in the range of 42 to 5000 ng L−1 in landfill leachate, whilst PFOS was from <9.5 to 4400 ng L−1 (Table 3). In the U.S., shorter-chain PFCAs and PFSAs were predominant over PFOA and PFOS, consistent with the transition of C4-based production.19,43 In Europe, particularly Northern European countries (i.e., Denmark, Finland, Norway, Sweden, and Germany), the reported values in landfill leachate for PFCAs and PFSAs fell in the range of <3.51–13[thin space (1/6-em)]056 and <1.6–4656 ng L−1, respectively,24,44–48 which are smaller than those in North America. Interestingly, PFOA and PFOS were more abundant compared to short-chain PFAS in most landfill sites of Europe. In the landfill leachate of Australia, Gallen et al.49,50 observed that the PFCA and PFSA concentrations were up to 13[thin space (1/6-em)]092 and 3843 ng L−1, respectively. The reported values are similar to those in European countries, but slightly smaller than those in North America. In comparison, in the largest developing country, China, much higher concentrations of 2849.4–250[thin space (1/6-em)]459.8 ng L−1 and <LOD – 48[thin space (1/6-em)]099 ng L−1 were observed for PFCAs and PFSAs in seven landfills, respectively.51 The high concentration profile is attributed to active manufacturing of PFAS-containing products such as construction materials, carpets, clothing, electronics, etc. Although China has restricted the production of long-chain PFAS, the past production still impacted the abundance and distribution of PFAS in the landfill leachate, with long-chain PFOA and PFOS as the prevalent compounds.

Several factors can affect the type and concentration of PFAS in landfill leachate, including the type and amount of waste landfilled, age of the landfill, sampling locations and years, climate, and industrial discharges. Typically, the PFAS concentration decreases with landfill aging due to biological activities.15 The production shift to short-chain PFAS in recent years may also contribute to the reduced concentration of long-chain PFOA and PFOS.52 Landfills located in wet weather regions may show much higher PFAS leaching than those in temperate and arid regions, because abundant precipitation facilitates the leaching process.15

3.3. Leaching and transformation of PFAS in landfill systems

While PFAS are generally considered recalcitrant to microbial degradation in the environment, biological processes may play an important role in the leaching of PFAS from various consumer products in landfill systems. For instance, in bench-scale studies, researchers found that the total leached PFAS concentration in live anaerobic reactors filled with municipal solid wastes was up to one order of magnitude higher than that in biologically inactive reactors in which abiotic processes, e.g., desorption, were dominant.20,57 Common biodegradation intermediates, including methylperfluorobutane sulfonamide acetic acid and n[thin space (1/6-em)]:[thin space (1/6-em)]2 and n[thin space (1/6-em)]:[thin space (1/6-em)]3 fluorotelomer carboxylates, increased steadily under methanogenic conditions, with the 5[thin space (1/6-em)]:[thin space (1/6-em)]3 fluorotelomer carboxylate becoming the most concentrated PFAS in live reactors; the low-level leaching in the abiotic reactors was primarily due to PFCAs ≤C8.20 The PFAS leaching was dependent of the type of waste. When waste carpets were fed into the bioreactors, short-chain (<C7) PFAS were dominant, whereas for clothing wastes, both long- and short-chain PFCAs were accumulated in the bioreactors.20

The water matrix compositions, landfill age, and environmental conditions including the climate conditions can also affect the fate and transport of PFAS in landfill systems.16,42 NOM may complex with PFAS via hydrophobic interactions, and thus may affect the leachability and degradability of PFAS. Inorganic anions may compete with anionic PFAS for adsorption sites and thus increase the mobility of anionic PFAS. On the one hand, inorganic cations may enhance adsorption by suppressing the surface negative potential of solid surfaces and by increasing the electrostatic interactions via the bridging effect.58

However, detailed information has been lacking on the mechanisms governing the biologically mediated PFAS leaching and biodegradation of PFAS under landfill conditions. In particular, as PFAS are generally considered not biodegradable, understanding the biological transformation of PFAS in landfill systems is of great significance in developing engineered biological treatment of PFAS in contaminated soil and water. To this end, knowledge is needed on leaching and transformation of PFAS from various source products and under different landfill conditions (type, age and environmental conditions).

4. Treatment technologies for PFAS in landfill leachate

With mounting concern about PFAS contamination of soil and water, and as landfill leachate is one of the important sources of PFAS, it becomes increasingly evident that PFAS in landfill leachate should be treated before PFAS spread into the surrounding environment. However, due to the unusual recalcitrance of PFAS, conventional treatment technologies, such as reduction/oxidation and biodegradation, are not effective to degrade PFAS. While AC and ion exchange are effective to adsorb PFAS, their applications are limited due to poor regeneration, production of large volumes of waste residual and subsequent disposal of the process wastes. The treatment becomes even more challenging for PFAS landfill leachate because of the complex and strong matrix effects. Treatment approaches specifically designed for PFAS removal from the leachate have been lacking though some limited studies have been reported. This section not only reviews technologies that have been tested for landfill leachate, but also those that can potentially be employed for treating PFAS in landfill leachate, including adsorption, ion exchange, photochemical, membrane and biological processes. Furthermore, the effects of operating parameters (e.g., pH, matrices, etc.) on the PFAS treatment efficacy are also reviewed wherever available. Table 4 provides an overview of the common treatment technologies that are then reviewed in the following sections.
Table 4 Overview of PFAS treatment methods reviewed in this work
Method Mechanism Advantage Disadvantage
Physical separation Adsorption Electrostatic attraction, hydrophobic interaction Easy to scale up Contaminants are not degraded, costly regeneration
Resin Ion-exchange reactions Fast and selective removal High material cost, costly regeneration
Membrane filtration Size exclusion, electrostatic interaction High removal rate, easy to scale up High operational cost, further disposal of the concentrate
Oxidation ˙OH Electron transfer, decarboxylate, CF2 elimination Easy activation of the radical precursor Not effective
SO4˙ Electron transfer, decarboxylate, CF2 elimination Easy activation of the radical precursor Slow degradation, partial mineralization
h+ Decarboxylate, unimolecular decomposition Energy-effective Easily scavenged
Reduction e, CO2˙ Electron transfer, unimolecular decomposition Faster degradation High chemical dosage, easily scavenged
Thermolysis Incineration, sonolysis Thermal dissociation Complete mineralization High energy demand, unscalable


4.1. Adsorption

Adsorption has been widely studied for removing PFAS from water or wastewater.59–61 In addition to carbon adsorption in engineered systems, natural adsorbents (e.g., clay, or iron oxide) can also adsorb PFAS in natural systems, thereby affecting the distribution, fate and transport of PFAS in natural systems.5 Typically, organic contaminants are adsorbed by various adsorbents through hydrophobic interactions, electrostatic interactions, hydrogen bonding, ion exchange, π–π bonding, and/or van der Waals force.62 For PFAS adsorption, because of the absence of π electrons, the low polarizability, and the small molecular size, the π–π bonding and van der Waals force are generally considered negligible.63 On the other hand, the strongly polar functional groups (COO and SO3) of anionic PFAS can induce strong electrostatic interactions and complexation with mineral surfaces. Besides, long-chain PFAS are more hydrophobic, and thus show higher adsorption capacities towards hydrophobic adsorbents such as AC and polystyrene-based resins.62 It should be noted that the matrix effect exerted by landfill leachate, which often contains high levels of NOM and salts, can alter the adsorption rate and extent of PFAS. This review focuses on carbon-based materials, minerals, and polymeric resins.
4.1.1. Carbon-based adsorbents. Carbon based adsorbents, such as activated carbon (AC) and carbon nanotubes (CNTs), have been shown effective in adsorption of PFAS (especially long-chain PFAS such as PFOA and PFOS) through hydrophobic interactions, electrostatic interactions and/or hydrogen bonding.62 In particular, AC has been one of the widely used adsorbents in water treatments due to its high adsorption capacity, low cost, and easy operation/maintenance.

Based on particle size, AC is divided into granular activated carbon (GAC, size ≥0.2 mm) and powdered activated carbon (PAC, size ≤0.1 mm).64 For adsorption of PFAS, especially long-chain PFAS, PAC offers a faster adsorption rate and higher capacity than GAC. For instance, Yu et al.64 reported that the equilibrium time for adsorption of both PFOA and PFOS by PAC was only 4 h, while it took over 168 h for GAC. The point of zero charge pH (pHpzc) was ∼7.5 for both PAC and GAC, both being coal-based ACs. As such, both electrostatic and hydrophobic interactions were involved in the adsorption of PFOA/PFOS at acidic or circumneutral pH (Fig. 1). However, the pHpzc value may vary remarkably depending on the source material. For example, Deng et al.65 reported a pHpzc value of 3.2 for a bamboo-derived GAC, in which case adsorption of anionic PFAS becomes less favourable due to surface repulsion.


image file: c9ew00645a-f1.tif
Fig. 1 Schematic illustration of adsorption of PFOA and PFOS on activated carbon via: electrostatic interactions (A) and hydrophobic interaction (B).64

AC has been used to treat landfill leachate in Germany,24 and the concentration of PFAS in the untreated landfill leachate ranged from 30.5 ng L−1 to 12[thin space (1/6-em)]922 ng L−1. The major PFAS in the leachate were PFBA (accounting for 27%), PFBS (24%), PFHxA (15%), PFOA (12%), PFPA (6.0%), PFHpA (4.0%), 6[thin space (1/6-em)]:[thin space (1/6-em)]2 FTS (3.7%), PFOS (2.7%), and PFHxS (2.3%). After treatment by AC, the concentration of PFAS in the leachate ranged from 9.26 to 4079 ng L−1, which was lower than that with conventional wet air oxidation (1992–4610 ng L−1) or biological treatment (4023–8059 ng L−1) for the same raw leachate.24 However, Yu et al.64 pointed out that low-molecular-weight NOM in landfill leachate seriously reduced the adsorption capacity and rate of PFAS by AC due to competition for adsorption sites, whereas large-molecular-weight NOM (>30 kDa) posed much less effect. It is noteworthy that NOM in landfill leachate can vary widely in both physical forms and chemical compositions. For instance, Campagna et al.66 reported that 37% of the COD in a landfill leachate was particulate or colloidal (>100 kDa) matter, whereas 63% was soluble (<100 kDa) with a broad distribution: 100–10 kDa (5%), 10–1 kDa (3%), 1 kDa–500 Da (11%) and >500 Da (44%). The fraction of smaller NOM accounted for >50% of the soluble COD, which may significantly inhibit PFAS adsorption by AC.

CNTs are a class of emerging carbon materials with a high surface area and strong adsorption potential for hydrophobic organic compounds. Deng et al.63 investigated the adsorption of PFOA on pristine and functionalized CNTs with hydroxyl and carboxyl groups (pHpzc < 3.5). The pristine CNTs showed over 3 times higher adsorption capacity for PFOA than the functionalized CNTs, where the hydroxyl and carboxyl groups were mostly deprotonated at the experimental pH (4–10), indicating that hydrophobic interactions dominated the adsorption and the deprotonated hydroxyl and carboxyl groups on CNTs were repulsive to PFOA anions. The PFOA adsorption on functionalized CNTs decreased with increasing pH in the range of 4–10 due to elevated electrostatic exclusion between the increasingly deprotonated hydroxyl/carboxyl groups on CNTs and PFOA anions.

To enhance interactions with the functional groups, the AC surface may be modified with, for instance, metal oxides. For example, in a recent work, Zhang et al.67 synthesized a coal-based magnetic activated carbon (MAC) for PFAS adsorption in a landfill leachate effluent that was pre-treated by anaerobic/aerobic (A/O) biological processes. The impregnated magnetite not only facilitated interactions with the head groups of anionic PFAS through both electrostatic and surface complexation, but allowed for magnetic separation of the spent PAC from treated water. At a MAC dosage of 53 g L−1, the composite was able to remove 72.8–89.6% of the following PFAS in 120 min: PFHxA (1435 ng L−1), PFHpA (423 ng L−1), PFOA (2683 ng L−1), PFNA (221 ng L−1) and PFDoA (53 ng L−1) in leachate samples (numbers in the brackets indicate initial concentration). The spent MAC was then recycled using a high-intensity magnetic separator, and then regenerated by ultrasonic treatment for 4 h, which was able to recover 76% of the adsorption capacity.

While AC has been widely accepted in drinking water treatment, CNTs are yet to be further investigated to be practically viable. Moreover, carbon-based materials are challenged with poor regeneration efficiency when used for PFAS removal. In fact, conventional thermal regeneration is less effective for PFAS, whereas uses of incineration and/or organic solvents are often cost-prohibitive. The regeneration is even more difficult in landfill leachate treatment due to the complex interference of the leachate matrix, especially organic fouling. Moreover, conventional regeneration processes do not degrade PFAS, and often result in large volumes of process waste residual that needs further handling and disposal, such as landfill disposal and incineration. While ultrasonic regeneration has shown some effectiveness, its practicality is yet to be demonstrated due to operating issues and potential damage to the adsorbents. To overcome these drawbacks, modifications of carbon-based materials should aim at both effective adsorption and subsequent regeneration. It is also highly desirable to develop a “Concentrate-&-Destroy” strategy, where PFAS are first effectively adsorbed/concentrated on the carbon surface and then degraded in situ, which also regenerates the materials without invoking other costly regeneration treatments.

4.1.2. Minerals. In the natural environment, minerals may interact with PFAS, and thus affect their fate and transport.68,69 In engineered systems, mineral-based adsorbents may be used as low cost adsorbents for PFAS. Clay minerals (e.g., kaolinite and montmorillonite), which are often rich in alumina and silica, are one of the most common minerals that participate in many important environmental processes due their unique structures.61 Clay minerals have also been used for removing organic and metal cations from water,70,71 because of their low cost, large specific surface area and high cation-exchange capacity. Zhang et al.72 investigated PFOS adsorption on kaolinite and montmorillonite at the molecular level using ATR-FTIR and EXAFS, and they found that PFOS could quickly form an outer-sphere complex on kaolinite and montmorillonite through ligand exchange.

Sodium bentonite, whose major constituent is montmorillonite with a high swelling capability, has been often used as a low-permeability containment barrier in landfill liners. However, Li et al.73 found that anionic PFAS in landfill leachate may not effectively bind to sodium bentonite at pH 8.4 due to the repulsive effect of the surface negative potential (pHpzc for bentonite is <7.5). Likewise, Johnson et al.69 observed that increasing the solution pH resulted in decreased PFAS adsorption on kaolinite and silica sand.

Boehmite (AlOOH) is a common form of hydrated aluminium oxide that exists in soil and sediments, and it has been an important industrial mineral adsorbent because of its high surface area (>300 m2 g−1).74 Boehmite has a rather high pHpzc (between 7.7 and 9.4), and thus shows a positive surface potential under circumneutral conditions.63 Hence, boehmite can favourably adsorb anionic PFAS through electrostatic attraction (eqn (1) and (2), where L indicates PFAS molecules). In addition, boehmite can also take up PFAS through ligand exchange with the hydroxyl groups, which are abundant on the boehmite surface (eqn (3)). Generally, increasing the solution pH from 4 to 7 has a moderate impact on adsorption of PFAS on boehmite, owing to an increase in ligand exchange reactions but a decrease of electrostatic interactions.63 The presence of chloride ions (0.1–100 mM) cause significant reductions in PFOS and PFOA adsorption on boehmite due to competitive adsorption on the active adsorption sites.63 Like AC, regeneration of exhausted boehmite is costly, and the spent material needs to be further handled or disposed of.

 
Al–OH + H+ → Al–OH2+(1)
 
Al–OH2+ + L → Al–OH2–L(2)
 
Al–OH + L → Al–L + OH(3)

Hematite is one of the most abundant iron (oxyhydr)oxides in weathered soil or sediments,75 and has been used to remove Cr(VI) from landfill leachate.76 Although they have not been tested for treating PFAS in landfill leachate, iron oxides including hematite are known to show higher PFAS adsorption capacities and rates than other clay minerals, owing to the more available ligand exchange sites on the iron (oxyhydr)oxide surface.68 PFOA may form inner-sphere Fe–carboxylate complexes with hematite, whereas the PFOS sulfonate group may form outer-sphere complexes and possibly hydrogen-bonds at the mineral surface.77 The presence of high concentrations of NOM, ammonia, bicarbonate and other competing ligands in landfill leachate can severely suppress PFAS adsorption on hematite.78,79 For the low material cost, hematite may be used as a disposable adsorbent, with the spent material disposed of in landfills.

While detailed information has been lacking on the use of various minerals for treating PFAS in landfill leachate, these natural or modified natural adsorbents show potential to be used as a class of disposable, low-cost and effective adsorbents for removing PFAS from the leachate. This is of particular practical significance given the costly regeneration of PFAS-laden conventional adsorbents. However, the adsorption efficiency can be severely curbed by the high strength matrix of landfill leachate, especially NOM. Therefore, other treatment approaches (e.g., biological or abiotic oxidative pre-treatment of NOM or more advanced adsorption) may be combined to achieve the treatment goals.

4.1.3. Polymeric resins. Polymeric ion exchange resins have attracted increasing attention for concentrating and removing PFAS, primarily because of their high adsorption capacity, good selectivity, and regenerability (e.g., using a mixture of NaCl and methanol solution).80 However, most studies on ion exchange methods have been focused on drinking water treatment, and information is lacking about the use of ion exchange resins for treating PFAS in landfill leachate.

Based on results from water treatment studies, adsorption of anionic PFAS is affected by the resin matrix, porosity, and functional groups. Deng et al.81 compared the adsorption of PFOS using six resins representing different matrices (gel vs. macroporous, polyacrylic vs. polystyrene) and functional groups (polyamine vs. quaternary amine), and found that IRA67 (polyacrylic and gel matrix with polyamine functionality) and IRA958 (polyacrylic and macroporous matrix with quaternary amine functionality) offered the highest adsorption capacity, with a maximum adsorption capacity of 5.70 and 4.86 mmol g−1, respectively. Both these capacity values exceeded the resins' original ion exchange capacities (5.6 and 3.7 meq. g−1), indicating that hydrophobic interactions enhanced the PFOS adsorption. Moreover, the authors claimed that the resin pore size had little influence on the adsorption kinetics of PFOS on the polyacrylic resins, which can be ascribed to the relatively small molecular size of common PFAS, and the effect of the basicity of functional groups was also minor. Consequently, the resin's specific surface area and ability to facilitate concurrent hydrophobic and electrostatic interactions governed the ion exchange effectiveness. Zaggia et al.82 compared another series of ion exchange resins and found that resins of higher hydrophobicity offered a higher adsorption capacity for both short- and long-chain PFAS, for example, the adsorption capacities of PFBA, PFOA, PFBS and PFOS followed the order: A532E (highly hydrophobic, bifunctional quaternary amine) > A520E (fairly hydrophobic, triethyl quaternary amine) > A600E (non-hydrophobic, trimethyl quaternary amine). Based on these findings, both electrostatic and hydrophobic interactions are important mechanisms in adsorption of PFAS by ion exchange resins. It is also noteworthy that transmission electron analysis by Zaggia et al.82 revealed the presence of large molecular macro-aggregates of PFAS in the intraparticle pores, suggesting that surface adsolubilization is also operative (a common phenomenon for surface accumulation of common surfactants).

Compared with AC, ion exchange resins offer higher adsorption capacity towards PFAS.83–86 For instance, AW-F900 (a weak base resin) offered a maximum Langmuir capacity of 1930.9 mg g−1 for PFOS, compared to 1320.6 mg g−1 for an AC.83 The higher adsorption capacity of polymeric resins is attributed to their concurrent columbic and hydrophobic interactions, whilst only hydrophobic interaction is the dominant adsorption mechanism for AC.

Overall, these engineered adsorbents are more effective than clay or iron/aluminium oxide minerals, which, however, come with a much higher price. Moreover, the strong matrix effects of landfill leachate should be taken into account in selecting the adsorbents for PFAS removal from landfill leachate. The high levels of organic and inorganic components in landfill leachate (Table 3) are expected to deteriorate the PFAS removal capacity and regenerability of engineered materials, and then the uses of low-cost minerals such as activated alumina, silica, zeolite, and montmorillonite may become more practical as disposable adsorbents.

The major downside with the adsorption method is that it merely transfers PFAS from the aqueous phase to the solid phase without degradation, and regeneration of the adsorbents are often costly, and may result in large volumes of PFAS-laden wastes, which require further treatment, such as landfilling. Often times, large amounts of organic solvents (e.g., methanol) and elevated temperature are required to regenerate AC and ion exchange resins, and energy-intensive processes like incineration are needed to treat the resulting process wastes.82 Zaggia et al.82 used reverse osmosis coupled with under-vacuum evaporation to reduce the volume of regeneration effluents requiring incineration, but the practicality needs to be further investigated.

4.2. Photochemical degradation

In recent years, intensive research studies have been reported to transform and mineralize PFAS via enhanced photochemical, sonochemical, electrochemical, or thermochemical treatments.87–90 While sonolysis, electrolysis and thermolysis are deemed to be high-cost and energy-intensive,91–93 photochemical degradation appears to be a more practical technology for PFAS removal from landfill leachate due to its lower cost, easy operation, and effectiveness under ambient conditions. Unfortunately, very little work has been reported on degradation of PFAS in landfill leachate. Consequently, the review in this section is largely focused on fundamental photochemistry associated with PFAS degradation and studies on the photochemical decomposition of PFAS in the context of drinking water treatment, which may provide useful reference and guidance to PFAS degradation in landfill leachate.
4.2.1. Direct photolysis. Ultraviolet light can provide strong energy to initiate various photochemical reactions that can transform POPs. For example, UVC at a wavelength of 185 nm may provide up to 647 kJ mol−1 of photon energy,94 which exceeds the C–F bond energy of 536 kJ mol−1 in PFAS.2 Indeed, it has been reported that PFAS could be excited by 185 nm vacuum ultraviolet (VUV) light, where PFAS are first decarboxylated by forming unstable perfluoalkyl radicals that are then decomposed upon reacting with water to form shorter-chain PFAS with one CF2 unit eliminated. For example, Jing et al.95 reported that 61.7% of PFOA (25 ppm) was degraded or transformed within 2 h in a nitrogen atmosphere under 185 nm VUV light irradiation, with 17.1% of fluorine defluorinated into fluoride. However, the photon energy of UV light decreased with increasing wavelength, e.g., the photon energy of 254 nm UV light is only 471 kJ mol−1, which is insufficient to break the C–F bond.

The homogeneous photolysis process may be enhanced by the use of soluble catalysts. For instance, ferric ions can complex with the carboxylate groups of PFOA, thereby facilitating the direct photolysis of PFOA under 254 nm UV.96,97 In fact, the Fe(III)–PFOA complexes extended the absorption wavelength of PFOA from 180–220 nm to the deep UV region (up to 400 nm), with a peak at around 280 nm.97 Wang et al.97 reported that, in the presence of 80 μM Fe(III), 80.2% of PFOA (48 μM) was decomposed and 47.8% defluorinated within 4 h of UV irradiation (254 nm, 23 W mercury lamp) in an oxic atmosphere. Liu et al.98 demonstrated that PFOA could be decomposed in the presence of Fe(III) under sunlight irradiation, though at a slower rate. In the presence of 480 μM Fe(III), about 97.8 ± 1.7% of 50 μM PFOA was decomposed in 28 days under sunlight, with 12.7 ± 0.5% being defluorinated. However, the process requires extremely acidic pH (∼2), which greatly discounted its practical feasibility.

The direct photolysis method, especially in the presence of Fe(III) under solar light irradiation, is straightforward and potentially cost-effective for treating POPs. However, its effectiveness for landfill leachate has yet to be explored, and may be affected by the complex matrices via light penetration interference, radical scavenging, and Fe(III) complex inhibition. In addition, direct photolysis requires rather acidic pH (<4), while the pH of landfill leachate is usually in the range of 4.5–9 (Table 2). On the other hand, landfill leachate contains abundant dissolved metal cations, and the pH of fresh leachate (in the anaerobic acid phase) can be quite acidic (4.2–5.5).37 Hence, further studies are warranted to assess the feasibility of the direct photolysis method for treating PFAS in landfill leachate.

4.2.2. Radical-mediated oxidation of PFAS.
Hydroxyl radicals. Advanced oxidation based on ˙OH radicals has been widely implemented in water and wastewater treatments, including landfill leachate. The Fenton process is one of the most common approaches to effectively produce ˙OH, and the process can be enhanced under UV and visible light to degrade PFAS. For example, Tang et al.99 found that 53.2% of PFOA (10 ppm) defluorinated in a Fenton system (Fe(II) = 2.0 mM, H2O2 = 30.0 mM, and pH = 3.0) after 5 h of UV irradiation, and the rate decreased slightly from 46% to 44% when the pH was increased from 3.0 to 5.0. A two-stage mechanism was postulated for the PFOA degradation process, including decarboxylation by ˙OH and a direct photolysis process, which is similar to that in the Fe(III)/UV system. As for the catalytic photolysis process, the presence of Fe(III) ions could facilitate electron transfer via the ligand-to-metal charge transfer mechanism under light irradiation,100,101 thus further promoting the degradation of PFOA in the system.

Typically, a Fenton system requires rather acidic pH to generate ˙OH. At neutral or alkaline pH, the system tends to generate more O2˙ and HO2 radicals. While ˙OH cannot effectively break the high-energy C–F bond, O2˙ and HO2 with high nucleophilic reactivity may readily attack electron deficient carbon atoms of PFAS. For instance, Mitchell et al.102 investigated PFOA (100 ppb) degradation using the three single reactive oxygen species in a homogeneous Fenton system without light irradiation, and they observed that no PFOA was degraded by ˙OH, but systems producing only superoxide degraded 68% PFOA (2 M H2O2, 0.5 mM Fe(III)-EDTA with excess ethanol at pH 9.5), and systems producing only hydroperoxide degraded 80% over 150 min, where the solution pH was increased to 12.8. Hence, it's possible that the Fenton process may work for decomposing PFAS at neutral to alkaline pH, which differs from its traditional uses. However, when used for treating PFAS in leachate, it may still need to overcome the matrix interference. To this end, combining Fenton with other approaches such as photochemical processes may better serve the purpose.


Sulfate radicals. Advanced oxidation using sulfate radicals (SO4˙) is an emerging technology for decomposing POPs. SO4˙ radicals offer a high redox potential of 2.6 eV.103 Typically, SO4˙ radicals are generated by activating suitable precursors, such as peroxymonosulfate (HSO5; PMS) and persulfate (S2O82−; PS), through chemical reactions or under UV irradiation (eqn (4) and (5)).
 
HSO5 + → SO4˙ + OH(4)
 
S2O82− + → 2SO4˙(5)

Compared with ˙OH, SO4˙ is more electrophilic and has been widely applied in PFAS degradation. For instance, Hori et al.104 reported that in the presence of 26.8 mM S2O82−, a photochemical system (UV-visible light at 220–460 nm, 35.8% transmittance at 254 nm, 200 W xenon-mercury lamp) was able to completely decompose 1.35 mM PFOA within 4 h, which was 11 times faster than direct photolysis. However, the PFOA degradation was severely suppressed by chloride (0.5–3 mM), and PFOA could not be degraded until Cl was converted ClO3.105 This can be a major drawback since chloride is omnipresent in natural waters. Moreover, as high concentrations of chloride and DOM are present in landfill leachate, further modification of this process is needed to overcome the matrix effects.

In the UV-S2O82− system, the degradation pathway of PFOA by SO4˙ is comparable to that with ˙OH. Briefly, the intermediate of PFOA˙+ is first formed by one-electron transfer from PFOA to SO4˙ (eqn (6)),104 which is then decomposed following the stepwise CF2 elimination reactions mirroring the ˙OH oxidation.

 
SO4˙ + C7F15COOH → SO42− + C7F15COOH˙+(6)

SO4˙ are one of the most reactive species, and when combined with photochemical treatments, they may provide powerful oxidation for PFAS in landfill leachate. However, the high concentrations of co-solutes in the leachate may render the method less efficient due to the lack of selectivity of SO4˙. Moreover, the hydrolysis of persulfate may cause a significant pH drop, which may inhibit the reaction rate and cause other engineering issues especially for leachate treatment. It should be noted that the resulting shorter-chain products are likely more difficult to degrade, and may need an extended reaction time or energy input to reach the final mineralized products F and CO2.


Carbonate radical anion. Carbonate or bicarbonate ions in aqueous solution can be activated by UV light through one-electron oxidation (e.g., by ˙OH in eqn (7)–(9)) to produce carbonate radical anions (CO3˙), with redox potentials of 1.59 V and 1.78 V at pH 12.5 and 7, respectively.106,107
 
˙OH + CO32− → OH + CO3˙(7)
 
˙OH + HCO3 → H2O + CO3˙(8)
 
CO3˙ + CO3˙ → CO2 + CO42−(9)

CO3˙ can work as a selective radical in decomposing PFAS via the electron transfer process. Thi et al.108 used H2O2 (0.075%) and 40 mM NaHCO3 solution under 254 nm UV (400 W) irradiation to produce CO3˙ (eqn (8)). They found that the system completely decomposed 50 ppm PFOA after 12 h, while UV irradiation alone degraded only 52.1%. In addition, the decomposition of PFOA with CO3˙ under UV irradiation is more favourable under moderately alkaline conditions (pH = 8.8), because the concentration of CO3˙ formed is low under acidic conditions. However, under highly alkaline conditions (pH 11), the concentration of CO3˙ could decay quickly through second-order self-combining reactions (eqn (9)). The mechanism of PFOA decomposition in the CO3˙-UV system is similar to that in the UV activated S2O82− system, where PFOA is first activated by CO3˙ to form PFOA˙+, which undergoes the same HF elimination and hydrolysis reactions, resulting in the production of shorter-chain PFAS.

Carbonate radical systems use low-cost and innocuous precursors, and are operated under moderate pH conditions. When properly activated and combined with other methods (e.g., photochemical processes), this approach holds promise for tackling PFAS in landfill leachate in a “green” and cost-effective manner.


Nitrogen dioxide radicals. Nitrogen dioxide radicals (˙NO2) can be produced by photolysis of nitrate (NO3) in aqueous solution under UV light (eqn (10)). ˙NO2 radicals offer a redox potential of 1.03 V, and can decompose organic compounds through an electron transfer process, hydrogen abstraction, and/or radical addition.109 As shown in eqn (10), the reaction is favorable to proceed in the presence of a ˙OH scavenger.
 
NO3 + + H+ → ˙NO2 + ˙OH(10)

Reactions between ˙NO2 and PFOA were first reported by Li et al.,110 where 23.7% of PFOA (initial 5 ppm) was degraded by in situ generated ˙NO2 during the photolysis of NO3 (100 mM) in aqueous solution under 254 nm UV irradiation. For comparison, the direct photolysis degraded only 7.6% of PFOA. Furthermore, the PFOA decomposition efficiency reached as high as 100% in the presence of a ˙OH scavenger (0.012 ppm isopropanol). The estimated reaction enthalpies were calculated by density functional theory (DFT), and results showed that the spin charge transfer from PFOA to ˙NO2 requires an input energy of 0.14 kJ mol−1, which is much lower than that (>2.14 kJ mol−1) for ˙OH attacking PFOA, implying that ˙NO2 is more thermodynamically feasible to withdraw electrons from PFOA compared with ˙OH.

Apparently, ˙NO2 radicals are less effective for PFAS degradation compared to the aforementioned radicals. However, it may find its uses when the leachate contains high concentrations of nitrate and natural scavengers for ˙OH radicals.


Periodate. Periodate (IO4, PI) is a well-known oxidant with a redox potential up to 1.60 V. Highly reactive species would form by photolysis of PI, such as IO3˙, ˙OH, and O˙ (eqn (11)–(13)).111
 
IO4 + → IO3˙ + O˙(11)
 
+ H+ → ˙OH(12)
 
˙OH + IO4 → OH + IO4˙(13)

The main free radical (IO3˙) can effectively decompose PFAS via pathways including abstraction of the fluorine atom or electron transfer from PFAS. Cao et al.112 reported that the addition of 0.5 mM IO4 significantly accelerated the rate of decomposition (70%) and defluorination (17%) of PFOA (0.01 mM) after 120 min of UV (254 nm) irradiation. Without IO4, the defluorination rate of PFOA was only 9%. In the UV/IO4 system, IO3˙ generated by photolysis of IO4 initiated oxidation of PFOA in two parallel reaction pathways: one involves the transfer of an electron from PFOA to yield PFOA˙+, and the other involves abstraction of one F atom from PFOA directly.

Compared with other radicals such as SO4˙ and ˙OH, much less information is available on IO3˙ in advanced oxidation processes. However, given the moderate oxidation potential and the unique defluorination mechanism, and considering the leachate matrix effect, IO3˙ may serve as a more selective radical for degradation of PFAS in landfill leachate.

Overall, information is lacking on the application of combined radical and photochemical processes specially targeting PFAS in landfill leachate. However, the foregoing review reveals that these radicals hold great potential for decomposing PFAS in the leachate due to the high redox potential and mild operating conditions. In addition, the photochemical processes and radical-based reactions may work synergistically. Further investigations are warranted to test the scavenging effects of NOM or salts in landfill leachate, which has been cited as a critical drawback for radical-based AOPs. To mitigate the inhibitive effects, some pre-treatments may be practiced to remove suspended solids and NOM from landfill leachate.

4.2.3. Radical-based reduction of PFAS.
Hydrated electrons. In general, a potential of +2.0 V (vs. NHE) is required to induce PFAS oxidation whilst −1.0 V is required to initiate PFAS reduction.113 Hence, reduction of PFAS is thermodynamically more favourable than its oxidation. A hydrated electron (eaq) is a powerful reducing agent with a standard reduction potential of −2.9 V,87 which has been proven to be effective for PFOA/PFOS defluorination. Unlike in the photo-oxidation process, eaq as a nucleophile can directly cleave the C–F bonds of PFAS in the photo-reductive processes.

Park et al.114 investigated the effects of chain length on PFAS reduction in the presence of KI under 254 nm UV light irradiation. The PFAS degradation rates were observed to decrease with the chain length of PFSAs, e.g., 3.0 × 10−3, 1.2 × 10−3, and 4.0 × 10−4 min−1 for PFOS, PFHxS, and PFBS, respectively. In contrast, the chain-length dependence was insignificant for PFCAs (i.e., PFOA, PFHA and PFBA) reacting with eaq. In addition, the presence of organic matter, such as humic and fulvic acids (HA and FA), can accelerate the rate of PFAS reduction in the UV/I system by acting as electron shuttles to enhance the electron transfer efficiency. For an example, 86.0% of PFOS (0.03 mM, pH 10) was degraded with a defluorination rate of 55.6% in the presence of HA (1.0 ppm) and KI (0.3 mM) after 1.5 h of UV (254 nm) irradiation, compared to only 51.7% of PFOS degradation and 4.4% defluorination in the absence of HA.115

Sulfite (SO32−) as an alternative mediator of eaq has also been investigated to reduce PFAS. Song et al.116 observed that 88.5% of PFOA (20 μmol L−1, pH 10.3) was reductively defluorinated by eaq in a SO32−/UV/N2 system (SO32− = 10 mM, 254 nm) after 24 h. In addition, PFOA degradation was increased with increasing SO32− concentration or solution pH due to accelerated eaq generation. However, regardless of KI or sulfite as the eaq mediator, the presence of dissolved oxygen (DO) could slow down the reduction reactions because DO consumes both eaq and ˙H.

The mechanism of PAFS reduction in the UV-eaq process differs from that in oxidative processes. First, eaq attacks the fluorine atom at the α-position of PFSAs or PFCAs due to their stronger electron affinity.114 For PFCAs of different chain lengths or carbon numbers, the degrees of defluorination are similar because the eaq mediated reduction occurs at a similar initial reaction site near the ionic head group (eqn (14)).114

 
CnF2n+1COO + eaq → CnF2nCOO˙ + F(14)

For PFSAs, however, the photodegradation starts from the cleavage of C–S bonds between the CnF2n+1 and sulfonate group. Thus, the PFAS reduction is more influenced by the type of ionic head group and chain length.114


Carboxyl anion radical. In the system of TiO2, oxalate sorbed on the surface of TiO2 reacts with h+ to form CO2˙.117,118 CO2˙ can be generated by photolysis of ferrioxalate complexes or by reacting oxalic acid or formic acid with photogenerated holes of excited TiO2 (eqn (15)).117,118 In the presence of ferric ions and oxalic acid, PFOA is decomposed via two pathways: 1) photochemical oxidation via Fe(III)–PFOA complexes, and 2) one-electron reduction by CO2˙,118 which is similar to that in the eaq system.
 
C2O42− + /h+ → CO2˙ + CO2(15)

In this role, oxalate acts as a hole-scavenger and suppresses direct hole-mediated oxidation of PFAS. On the other hand, the produced CO2˙ from oxalate (eqn (15)) may compensate for this effect. For instance, CO2˙ can donate an electron to PFOA to induce decarboxylation of PFOA, and then, the resulting C7F15˙ further undergoes hydrolysis and HF elimination to produce shorter-chain PFAS following the step-by-step defluorination mechanism.

Comparing CO2˙ with eaq, the latter is more effective for reductively decomposing PFAS due to its higher reduction potential. However, when used for landfill leachate, eaq may also undergo more fierce competition by side reactions with co-present scavenging species (e.g., H+, DOM, DO, and NO3), resulting in a very low proportion of eaq available for decomposing PFAS.119 Compared to the oxidative processes, much less information is available on reductive degradation of PFAS. However, given the unique degradation mechanisms and effectiveness, further investigations are warranted to polish the radical-based reductive processes to decompose PFAS in water and wastewater. Of particular interest would be combining the abiotic process with anaerobic treatment of landfill leachate.

4.2.4. Photocatalytic degradation. Compared to the homogeneous photolysis, photocatalytic processes offer the advantages of: 1) much more powerful redox potential (i.e., reactivity) and 2) lower energy input (i.e., can be operated under higher wavelength light, i.e., UVA, UVB or visible light). Semiconductors (e.g., TiO2) are widely used in photocatalytic degradation of POPs in water or wastewater.120 When the semiconductor is irradiated by light with an appropriate wavelength, photogenerated electrons (ecb) and holes (h+) are produced in the conduction band (CB) and valence band (VB), respectively.121 The formed ecb in the CB facilitate strong reduction reactions and generate radicals such as O2˙, whereas the h+ in the VB can facilitate direct hole-oxidation and/or react with H2O producing ˙OH.121 It has been demonstrated that ˙OH is ineffective for directly decomposing PFAS,122 because no hydrogen atoms were available for abstraction by ˙OH. Hence, in typical photocatalytic systems using semiconductors, PFAS decomposition primarily depends on direct oxidation by h+ or reduction by ecb and O2˙.

TiO2 has been the most widely investigated semiconductor towards PFAS degradation.87 However, TiO2 alone showed low efficiency of PFAS degradation because ˙OH radicals are the main reactive species. Moreover, Sansotera et al.123 reported that F generated from the PFAS degradation reaction could modify the TiO2 surface, thereby hindering the photocatalytic activity of TiO2 by limiting the charge carriers' mobility. Because the material reactivity increases with reactive surface area of semiconductors, immobilizing TiO2 onto porous supports has been shown to remarkably improve PFAS degradation. For instance, Song et al.124 loaded TiO2 on multi-walled carbon nanotubes (MWCNTs) through a sol–gel method, and observed complete degradation of PFOA (30 ppm, pH 2) after 8 h of UV light irradiation (365 nm). The presence of MWCNTs not only increases the adsorption capacity of PFOA, but also facilitates ecb transfer from activated TiO2 to MWCNTs, resulting in reduced recombination of the ecb–h+ pairs.

In addition, TiO2 modified with transition metals, e.g., Fe, Nb, Cu, Pb, and noble metallic Pt or Pd, showed higher PFAS removal rate than neat TiO2.122,125,126 The transition or noble metals have been demonstrated to successfully trap the photo-induced ecb from TiO2, thereby reducing the recombination of ecb–h+ pairs during photocatalytic processes and thus enhancing the capability of h+ to react with PFAS.100,122 Furthermore, the difference in valence state between Ti4+ and the metallic ions could create oxygen vacancies for charge compensation,122 which is beneficial to separation of the ecb–h+ pairs.

Nanostructured In2O3 has been shown to be more effective for oxidative degradation of PFAS than TiO2 as a photocatalyst.127 In2O3 featuring three nanostructures, i.e., porous microspheres, nanocubes and nanoplates, was obtained by dehydration of the corresponding In(OH)3 nanostructures at 500 °C for 2 h with different mixed solvents.128 The porous In2O3 microspheres presented the highest photo-activity for PFOA degradation (30 ppm, pH 3.9) under 254 nm UV irradiation, and the degradation rate was 74.7 times higher than that of P25. XPS results suggested that PFAS molecules could insert an O atom from their terminal groups, such as COOH or SO3H, into an oxygen vacancy site on the In2O3 surface to form a tight and close contact with In2O3, which is beneficial for direct charge transfer and subsequent photocatalytic decomposition under UV irradiation. Hence, it was concluded that the more effective degradation of PFOA by porous In2O3 microspheres than nanocubes and nanoplates was due to the higher oxygen vacancy defects as well as the higher specific surface area.

Alternatively, PFAS may also be decomposed through photocatalytic reduction with ecb. Zhao and Zhang129 demonstrated that PFOA (pH 4.7) was reductively degraded by a “sheaf-like” β-Ga2O3 nanomaterial in 45 min under UVC irradiation (254 nm). The high conduction band edge potential of β-Ga2O3 (−1.55 eV vs. NHE) allows for a direct heterogeneous reaction of PFOA with ecb. Most notably, much faster PFOA degradation and defluorination were achieved using petitjeanite, Bi3O(OH)(PO4)2 microparticles.130 The rate constant for degradation of PFOA by Bi3O(OH)(PO4)2 was 15 times greater than those of BiPO4 and β-Ga2O3 (∼20–30 times greater when normalized to the surface area). The superior performance of Bi3O(OH)(PO4)2 was primarily related to the positive surface charge that is conducive to adsorption of PFOA, in combination with the favorable redox potential of the Bi3O(OH)(PO4)2 charge carrier.

In fact, the photoactivity of semiconductors towards PFAS degradation can be improved not only by increasing the specific surface area, but also by enhancing the separation of ecb and h+. Graphene is an excellent carrier for semiconductor nanoparticles because of its high specific surface area (2630 m2 g−1) and remarkable electrical conductivity,131 which could significantly accelerate the transfer of photo electrons from the semiconductor to PFAS. Huang et al.132 loaded the semiconductor SiC (conduction band potential = −1.6 eV) to graphene, such that the excited ecb on SiC transferred rapidly to graphene, then further transferred to perfluoroalkyl groups due to the strong electronegativity of the fluorine atoms. The resulting PFOA decomposition rate constant for the SiC/graphene composite was 0.096 h−1, which was 2 and 2.2 times higher than those for SiC and P25, respectively.132

The photocatalytic degradation holds great promise for treating PFAS in landfill leachate for the following advantages: 1) unlike the radical-based oxidative/reductive methods, no precursors are added, thus avoiding potential secondary pollution with the generation of by-products, such as SO42− and I, 2) the long-term stability and reusability of the photocatalysts could reduce the operational cost,133 and 3) adsorptive photocatalysts, e.g., SiC/graphene or TNTs@AC (titanate nanotubes loaded on AC134) may facilitate selective adsorption of the target PFAS from the complex matrix solution such as landfill leachate. Such a “concentrate-and-degradation” strategy is particularly promising for leachate treatment as the materials can be specifically tailored to selectively extract PFAS from the leachate and then get photodegraded. However, further studies are needed to validate the concept, especially when used to decompose PFAS in high-strength wastewater and leachate. Nonetheless, the applications of heterogeneous photocatalysts to landfill leachate will need to overcome the strong matrix effects such as competitive adsorption and blocking or poisoning of the reactive sites. For instance, Li et al.127 reported that the photocatalytic decomposition of PFOA by In2O3 was severely inhibited in treating a secondary effluent from a municipal wastewater plant due to competitive adsorption of NOM. Therefore, tailoring the photocatalysts for selective adsorption of PFAS is a critical step for treating PFAS in landfill leachate, for example, developing metal-impregnated carbonaceous materials may induce concurrent hydrophobic interactions (with the tail of PFAS) and Lewis acid–base interactions (with the head groups), thereby enhancing the selectivity towards PFAS.

4.3. Membrane filtration

Membrane technologies have come a long way in water purification, particularly in the field of seawater desalination. Various membrane-based treatments have been investigated for removing PFAS from contaminated water. While size exclusion is considered the main mechanism for PFAS rejection, other factors such as charge repulsion, hydrophobicity, and dipole moment also affect the process. Tang et al.135 tested the rejection of PFOS in a semiconductor-manufacturing wastewater using four reverse osmosis (RO) membranes (thin-film composite polyamide: ESPA3 and LFC3 from Hydranautics, BW30 from Dow FilmTec, and SG from GE Osmonics) and observed >99% rejection over a wide range of PFOS concentrations (0.5 to 1500 ppm). However, the presence of isopropyl alcohol detrimentally reduced the membrane water flux, suggesting pre-treatment of the leachate may be necessary to avoid membrane fouling by organic compounds in the leachate. In comparison, nanofiltration membranes (e.g., DK, NF90, and NF270 from Dow FilmTec) with pore size slightly larger than the RO membranes rejected 90–99% of PFOS.136 Over a long filtration time, the water flux was gradually reduced due to the entrapped PFOS molecules in the polyamide layer, particularly with rougher surfaces. Steinle-Darling and Reinhard33 further tested nine PFCAs, five PFSAs, and one perfluorooctane sulfonamide (PFOSA) with four NF membranes (DK and DL from GE Osmonics, NF200 and NF270 from Dow FilmTec). They observed that the rejection rates were >95% for PFAS with molecular weight (MW) >300 g mol−1, whilst only 42% for PPOSA (MW = 499 g L−1), which is the only uncharged PFAS. The charge of the solute played a critical role in the rejection process as the membranes are negatively charged.

Appleman et al.137 evaluated the rejection of a suite of PFAAs, i.e., PFBA (M.W. = 214), PFPeA (264), PFHxA (314), PFOA (414), PFNA (464), PFDA (514), PFBS (300), PFHxS (400), and PFOS (500), using a nano-filtration membrane NF270 (Dow FilmTec). For the PFCAs, the rejection rate increased with molecular weight (93% to 99%), suggesting that size exclusion played an important role. Yet, for the three PFSAs, the size-dependence became much less profound. In addition to the size exclusion, the electrostatic exclusion of the negatively charged PFAS by the negative membrane surface further enhanced the rejection. The presence of NOM showed only a slight impact on the PFAS rejection, which is different from the RO process. When the M.W. of NOM (e.g., humic acid) was much larger than the MW cut-off of the NF membrane, the NOM molecules were deterred from being entrapped in the inner pores. The resulting exterior NOM layer may enhance PFAS rejection due to increased steric hindrance and/or electrostatic exclusion effects.

When tested for PFAS removal from landfill leachate, both NF and RO exhibited high removal rates (>90%), yet extensive pre-treatments like biodegradation of organic pollutants and ultrafiltration (UF) of (bio)solids were required.51 Meanwhile, the NF/RO processes yielded 15–20% of the original volume of the PFAS-laden concentrate, which may require further costly handling and disposal (e.g., incineration). Comparing NF and RO, both processes are comparably effective for rejection of both long- and short-chain PFAS, however, NF requires lower external pressure (25–140 psi)137 than RO (>200 psi),135 and appears less vulnerable to organic fouling. Consequently, NF appears more promising for removing PFAS from landfill leachate. However, long-term operating data at the pilot- or full-scale are needed to demonstrate and validate the practical viability (e.g., membrane lifetime, organic fouling and cleaning, and waste disposal) of the technology.

4.4. Biological treatment

Based on our knowledge so far, anionic PFAS, typically PFOA and PFOS, are not susceptible to biodegradation in conventional biological processes (e.g., activated sludge) due to the high strength of the C–F bond.138 For example, Yu et al.22 reported 14.1–638.2 ng L−1 of PFOA and 7.9–374.5 ng L−1 of PFOS in the influent of two conventional activated sludge units, whilst the PFOA and PFOS concentrations in the secondary effluent were 15.8–1057.1 ng L−1 and 7.3 – 461.7 ng L−1, respectively. A similar mass flow trend was also observed for municipal wastewater treatment plants (WWTPs) that employed membrane bioreactors (MBRs).22 Therefore, the effluent of WWTPs is a potential source of PFAS to be released to the environment. Moreover, it should be noted that the activated sludge can adsorb PFOA and PFOS with a capacity range of <5.0–69.0 ng g−1 and 13.1–702.2 ng g−1, respectively. As the PFAS-related regulations continue to mount, the disposal of the PFAS-laden sludge may be constrained to avoid the spread of PFAS.

Various microbial cultures have been tested for their biodegradation potential for PFAS. For example, Liou et al.139 tested five different microbial communities (one from a WWTP, one from industrial site sediments, two from fire training areas, and one from agricultural soil) for degrading PFOA. In the anaerobic incubations, PFOA was injected as an electron acceptor, whilst hydrogen gas, acetate, lactate, or ethanol as an electron donor. However, no significant PFOA degradation was observed for up to 259 days of operation in the presence of either 100 ppb or 100 ppm PFOA. In a recent study, however, Luo et al.140 reported that 40% of PFOA (1.0 μM) in the soil slurry was degraded in 140 days, and the researchers postulated that laccase produced by Pleurotus ostreatus and Pycnoporus sp. SYBC-L3, was able to activate some natural mediators, resulting in organic radicals that attacked the C–C bonds of PFOA, resulting in the formation of shorter-chain perfluorochemicals.

Earlier, Kwon et al.141 reported that Pseudomonas aeruginosa strain HJ4 from aerobic activated sludge was responsible for 67% removal of PFOS (1400–1800 μg L−1) within 48 h. Yet, no fluoride ions were detected in the biodegradation process, suggesting that PFOS was partially broken down into intermediate products without rupturing the C–F bonds. Ochoa-Herrera et al.142 confirmed the high resistance of PFOS and its shorter-chain by-products, e.g., PFBS, trifluoroacetic acid, 6[thin space (1/6-em)]:[thin space (1/6-em)]2 fluorotelomer sulfonic acid (FTSA), to different activated sludge systems under both anaerobic and aerobic conditions in the time span from 25 to 177 weeks.

In comparison, Schultz et al.21 observed decreased concentration for shorter-chain PFAS, i.e., PFHxA and PFHxS, after activated sludge treatment, suggesting that microorganisms may degrade more hydrophilic and mobile PFAS of shorter carbon chains. Other than PFCAs and PFSAs, perfluoroalkyl sulfonamides (PPOSA) were observed to increase in the activated sludge process, which may be attributed to the biodegradation of their precursor compounds. In fact, the degradation of PFAA precursors (high MW derivatives can transform into PFAAs), such as fluorotelomer olefins, could possibly explain why the mass flow of fluorochemicals increased after the activated sludge treatment.3

Yan et al.51 monitored the mass flow of PFAS in the leachate from five landfill sites in China that was subjected to MBR treatment, and found that the amount of PFBA, PFPeA, PFDA, PFNA, PFOA, PFDA, and PFHxS all increased, indicating PFAAs, particularly PFCAs, were not susceptible to the MBR process. Again, the mass increase was attributed to the biodegradation of PFAS precursors including fluorotelomer alcohols or sulfonates.

Based on the limited information, biodegradation of PFAS is dependent upon the compound properties, e.g., carbon chain length and type of microorganism. While most PFAS are considered recalcitrant to biodegradation, it is hopeful to find the right microbial enzymes given the diverse pool of microorganisms. Biological treatment of PFAS is particularly desired for treating PFAS in contaminated soil or landfill wastes including the leachate because of its environmental benign nature and low cost. As many agencies are actively sponsoring such efforts (e.g., the DoD's SERDP program has been actively seeking biological methods for treating AFFF-contaminated groundwater), technical breakthroughs are expected to take place in the field. Alternatively, combining biotic and abiotic processes may facilitate more effective degradation of PFAS. For instance, long-chain PFAS may be first converted into shorter-chain PFAS via photocatalytic processes, which are then subsequently degraded by biological methods.

4.5. Other treatments

Many technologies have been tested to remove POPs from landfill leachate, such as coagulation–flocculation, wetland, and sonolysis. For landfill leachate treatment, the uses of coagulants (e.g., Ferric chloride, aluminium sulphate and lime) and flocculants (e.g., polyelectrolytes N200, K1370, and K506) were able to remove up to 80% of COD (5350 mg L−1 and 70[thin space (1/6-em)]900 mg L−1 for stabilized and young leachates, respectively).143 Up to ≤20% of PFOA/PFOS was removed from drinking water by alum-based coagulation (dosage of 10–60 mg L−1 and pH of 6.5–8.0), where the NOM concentration was low (<13.3 mg L−1) and its impact on the PFOA/PFOS adsorption was insignificant.144 However, as the content of NOM in landfill leachate is several orders of magnitude higher than that in drinking water, PFAS removal by coagulation–flocculation is expected to be largely inhibited due to the competition for the adsorption sites and complexation between NOM and PFAS. However, coagulation–flocculation could be a viable pre-treatment option for the subsequent NF or RO or AOP treatments.

Constructed wetland (CW), which incorporates biodegradation, phytoremediation, and adsorption processes, has been widely employed to treat POPs in waste streams. Yin et al.145 demonstrated treatment of a series of PFAS (7 PFCAs, 4 PFSAs, and 7 PFAA precursors) in landfill leachate in a full-scale CW under the tropical weather conditions of Singapore. The raw leachate was blended with surface water in an equalization tank, and then introduced to aerated lagoons, followed by a sedimentation tank. The effluent from the sedimentation tank was then distributed to five reed beds and then entered five ponds as the final stage. A total of 61% PFAS and 50–96% individual compounds were removed in the CW treatment. Particularly, 42–49% of PFAAs were removed by adsorption onto soil/sediments or plant uptake in the reed bed, whilst PFAA precursors such as 5[thin space (1/6-em)]:[thin space (1/6-em)]3 fluorotelomer were mostly (>55%) biodegraded in the aerated lagoons. However, the removal of short-chain PFAS, i.e., PFHxA, PFHpA, PFBS, and PFHxS, was less effective (50–63%) than long-chain PFAS and PFAA precursors, which can be attributed to the higher mobility and resistance to abiotic processes (e.g., adsorption and photolysis) of shorter-chain PFAS.

Sonolysis can thermally degrade PFAS at the bubble–water interface at the temperature level similar to incineration.146 The extremely high temperature upon the collapse of the cavitation bubbles also dissociates water molecules into ˙OH and ˙H radicals that can further enhance degradation of PFAS. The effect of organic matter was found insignificant during sonolysis of PFOS and PFOA under environmental matrix conditions.147 Given the very high temperature and energy level, sonolysis of PFAS in landfill leachate is likely to work, and further studies are warranted under landfill leachate conditions. Alternatively, sonolysis may be combined with other treatment approaches to maximize the cost effectiveness. For instance, Cheng et al.147 combined ozonation and sonolysis for enhanced degradation of PFOS and PFOA in leachate-contaminated groundwater, and observed complete mineralization of the PFAS. However, the high-energy demand has been hindering the full-scale application of the ultrasound technology. In addition, there is a need to develop more energy effective and environmentally friendly (e.g., low noise) ultrasound equipment that is suitable for field-scale water/wastewater treatment uses.

5. Key knowledge gaps, research direction and concluding remarks

As health and ecosystem concerns over PFAS continue to mount, it is anticipated that stringent regulations for PFAS (PFOA and PFOS) will be stipulated in the near future, and accordingly, more cost-effective clean-up technologies are urgently needed. In February 2019, the U.S. EPA148 initiated a comprehensive Action Plan to target the detection, treatment, and prevention of PFAS contamination.

This review provides a state-of-the-science overview on the occurrence, chemistry and technologies of PFAS in the context of landfill leachate treatment. The review confirms that landfill leachate is a major source of PFAS due to massive uses of PFAS-containing everyday products, and there is a need to control PFAS in the leachate to prevent contamination of the neighbouring soil and groundwater. While the occurrence, distribution, and transformation of PFAS in landfill leachate are relatively better documented, and while various technologies have been widely tested for treating PFAS in drinking water, our knowledge is very limited on the treatability of PFAS in landfill leachate, in particular, pertaining to the effect of the strong leachate matrix.

Carbon adsorption and ion exchange have been found effective for separating anionic PFAS in normal water, especially for longer-chain PFAS. However, regeneration of spent conventional adsorbents or resins is rather demanding, and often requires costly follow-up treatment (e.g., thermal or incineration). This downside is exacerbated when it is used for treating PFAS in landfill leachate, where the adsorbents are confronted with competitive adsorption and organic fouling. Therefore, there is a need to develop selective adsorbents that take advantage of the unique chemistry of PFAS, such as metal–carbon composite materials that can interact with PFAS through concurrent electrostatic and hydrophobic interactions. In addition, low-cost natural minerals may be used as disposal adsorbents for removing PFAS from landfill leachate.

While both NF and RO processes can effectively reject PFAS, NF appears more practical than RO due to the lower fouling potential and lower energy demand of the former. However, further research is needed to demonstrate the effectiveness for landfill treatment. In addition, the handling and disposal of the waste concentrate containing high levels of PFAS need to be further investigated.

Advanced oxidation, reduction, photochemical processes and sonolysis are all able to degrade or transform more than 90% of the PFAS (especially long-chain PFAS like PFOA and PFOS) in drinking water, though only limited mineralization (<50%). However, the effectiveness is remarkably curtailed or even ceased when used for treating landfill leachate due to the severe water matrix effect. Radical-based reduction of PFAS is thermodynamically more favourable than radical-based oxidation, though both showed powerful degradation of PFAS. However, the lack of selectivity of the radicals reduced their efficacy for treating PFAS in landfill leachate due to competition of high concentrations of co-existing solutes especially NOM. Likewise, photolysis and photocatalytic processes also need to overcome the matrix effect (NOM, SS and metals). To this end, heterogeneous adsorptive photocatalysts are the most promising for the following advantages: 1) as adsorbents, the well-tailored functionalities may facilitate selective adsorption (concentrating) of the target PFAS onto the photoactive sites, and 2) as photocatalysts, the pre-concentrated PFAS can be photodegraded in situ, which also automatically regenerates the material. Such materials may be obtained by integrating carbonaceous materials and suitable semiconductors such as the TNTs@AC developed by Liu et al.134 Therefore, such a “concentrate-&-destroy” strategy is particularly suitable for treating landfill leachate and other high-strength wastewater. Converting the stronger C–F bond to the weaker C–H bond via redistribution with the Si–H bond on the SiC/graphene catalyst (Si–H + C–F → Si–F + C–H) offers an interesting option for PFAS degradation.132,149 The modified adsorptive photocatalyst may selectively adsorb PFAS, and the converted C–H bond is more susceptible to radical oxidation than the C–F bond,150 though regeneration of the saturated catalyst remains an issue.

Another innovative research direction is to develop technologies that combine physico-chemical and biological processes to take advantage of the unique features of each and maximize the overall treatment cost-effectiveness. For instance, priming PFAS by photocatalytic processes may convert hardly-biodegradable PFAS to more biocompatible intermediates, which then can be taken over by subsequent microbial processes. It is noted that the technologies for PFAS treatment are still in the rudimentary stage, and even in the proof-of-concept stage when used for leachate treatment. Reliable data on their technical viability and economic effectiveness are yet to be acquired and validated.

Given the impacts of landfill leachate as a major source of PFAS, this review senses an urgent need to seek more cost-effective and environmentally friendly solutions for selective transformation and mineralization of PFAS in the complex matrix of landfill leachate. To this end, the selected state-of-the-art technologies reviewed herein, and the chemistry and knowledge derived from the literature will not only help understand the state of the science, but may guide further research technology development for treating PFAS in landfill leachate as well as in natural, municipal, and industrial (waste)waters.

Nomenclature

ACActivated carbon
AFFFAqueous film forming foam
AOPAdvanced oxidation process
CNTCarbon nanotube
DOMDissolved organic matter
eaqHydrated electron
EtFOSEEthyl-perfluorooctane sulfonamidoethanol
FAFulvic acid
FTFluorotelomer
FTSFluorotelomer sulfonic acid
GACGranular activated carbon
GOGraphene oxide
h+Electron hole
HAHumic acid
HDFHydrodefluorination
MCLMaximum contaminant level
MeFBSAAMethyl-perfluorobutane sulfonamide acetic acid
MWCNTMultiple wall carbon nanotube
NFNanofiltration
NMRNuclear magnetic resonance
PACPowder active carbon
PAPPolyfluoroalkyl phosphate ester
PFAAsPerfluoroalkyl acids
PFASPer- and polyfluoroalkyl substances
PFBAPerfluorobutanoic acid
PFBSPerfluorobutane sulfonic acid
PFCAPerfluoroalkyl carboxylic acid
PFDAPerfluorodecanoic acid
PFDoAPerfluorododecanoic acid
PFHpAPerfluoroheptanoic acid
PFHxAPerfluorohexanoic acid
PFHxSPerfluorohexane sulfonic acid
PFNAPerfluorononaic acid
PFOAPerfluorooctanoic acid
PFOSPerfluorooctane sulfonic acid
PFSAPerfluoroalkyl sulfonic acid
PFOSAPerfluorooctane sulfonamide
POPPersistent organic pollutant
PPCPPharmaceutical and personal care product
FTCAFluorotelomer carboxylic acid
FTSAFluorotelomer sulfonic acid
PZCPoint of zero charge
ROReverse osmosis
SWCNTSingle wall carbon nanotube
TCETrichloroethylene
TNTTitanate nanotube
TOCTotal organic carbon
UFUltrafiltration
VOCVolatile organic compound
VUVVacuum ultraviolet

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

This work was partially funded by the Strategic Environmental Research and Development Program (SERDP) (ER18-1515) and the Auburn University IGP program.

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