EPA 600/R-22/066 I August 2022 i www.epa.gov/research
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Introduction to PFAS in Groundwater
by
Lee K. Rhea
Center for Environmental Solutions and Emergency Response
Groundwater Characterization and Remediation Division, Subsurface Remediation Branch
919 Kerr Research Drive, Ada, OK 74820
Tae K. Lee
Center for Environmental Solutions and Emergency Response
Water Infrastructure Division, Chemical Methods and Treatment Branch
26 West, Martin Luther King Drive, Cincinnati, OH 45268
MallikarjunaN. Nadagouda
Center for Environmental Solutions and Emergency Response
Water Infrastructure Division, Chemical Methods and Treatment Branch
26 West, Martin Luther King Drive, Cincinnati, OH 45268
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DISCLAIMERS
Any mention of trade names, manufacturers or products does not imply an endorsement by the
United States Government or the U.S. Environmental Protection Agency. EPA and its
employees do not endorse any commercial products, services, or enterprises.
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TABLE OF CONTENTS
LIST OF ACRONYMS 4
EXECUTIVE SUMMARY 8
INTRODUCTION 11
Description 11
Manufacturing 12
Synthesis 12
Impurities 12
Products 13
History 13
Classification 14
Polymers 14
Non-polymers 14
Manufacturing Method 15
Properties 15
Backbone and Hydrophobic Interactions 15
Functional Groups and Electrostatic Interactions 15
Partitioning to Media Interfaces 16
Environmental Transformation of Precursors 16
Uses 17
Aqueous Fire-Fighting Foam (AFFF) 17
Environmental Occurrence 19
Toxicity and Regulation 19
Environmental Fate and Transport 20
Subsurface Transformations 20
Vadose Zone 20
Phreatic Zone 22
Modeling 24
SITE CHARACTERIZATION 26
Conceptual Site Model (CSM) 26
Release Sites 26
Subsurface Characterization 27
Analytical Suite 28
Data Reduction 28
ANALYTICAL METHODS 29
Methods for Individual Analytes 29
US EPA Method 537.1 29
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US EPA Method 533 29
US EPA SW-846 Methods 3512/8327 29
US EPA Method 1633 30
ASTM D-7968 and D-7979 30
Other Analytical Methods in Development 30
Total Oxidizable Precursors Assay 31
Combustion Ion Chromatography Methods 31
Nuclear Magnetic Resonance 32
Fluorine Specific Electrode 32
SAMPLING 33
Equipment 33
Quality Assurance/Quality Control 35
US EPA Approved Analytical Methods 36
Method 537.1 36
Method 533 36
Method 8327 37
Draft Method 1621 37
Draft Method 1633 37
REMEDIATION 38
Comparative Reviews 38
Dominant Remedial Methods 39
Developing Remedial Methods 39
Short-chain PFAS 40
Handling of Remedial By-Products 40
Adsorbents 40
Sorption Materials 41
Sorption Mechanisms 42
Influence of Environmental Conditions and Co-contaminants 43
Regeneration of GAC and IXR 43
Reverse Osmosis (RO) and Nanofiltration (NF) 44
Biodegradation 45
Oxidation 47
Reduction 48
Other Technologies 49
Treatment Train Processes 50
SUMMARY 51
REFERENCES 52
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LIST OF ACRONYMS
6:2 FTSA - 6:2 Fluorotelomer sulfonic acid
7:3 FTCA - 7:3 Fluorotelomer carboxylic acid
AC - activated carbon
AEC - anion exchange capacity
AFFF - aqueous firefighting foam
A1 - aluminum
AOF - adsorbable organic fluorine
AOP - advanced oxidation process
AR-AFFF - alcohol-resistant AR-AFFF
ARP - advanced reduction process
ASTM - American Society for Testing and Materials
BOHP - petitjeanite Bi30(0H)(P04)2
Br - bromine
BTEX - benzene, toluene, ethylbenzene, and xylenes
Ca - calcium
CAC - colloidal activated carbon
CAS - chemical abstracting service
C-C - carbon-carbon bond
CEC - cation exchange capacity
C-F - carbon-fluorine bond
C-H - carbon-hydrogen bond
CIC - combustion ion chromatography
CI" - chloride ion
CI - chlorine
CMC - critical micelle concentration
CMT - critical micelle temperature
CNT - carbon nanotubes
C-0 - carbon-oxygen bond
CSM - conceptual site model
CVOC - chlorinated volatile organic compound
DO - dissolved oxygen
DOC - demonstration of capability
DoD - Department of Defense
DOM - dissolved organic matter
eaq- - aqueous electrons
ECF - electrochemical fluorination
EO - electrochemical oxidation
EOF - extractable organic fluorine
ETFE - ethylene tetrafluoroethylene
F" - fluoride anion
F - fluorine
FEP - fluorinated ethylene-propylene
FFFP - foam forming fluoroprotein
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foe - fraction organic carbon
FTOH - fluorotelomer alcohols
FTSA - fluorotelomer sulfonates
GAC - granular activated carbon
H* - hydrogen radical
HA - humic acid
HDPE - high-density polyethylene
HF - hydrogen fluoride
HFPO-DA - hexafluoropropylene oxide dimer acid (GenX)
HO - hydroxide radical
H02- - hydroperoxyl radical, also known as the hydrogen superoxide
I" - iodide anion
I - iodine
IUPAC - International Union of Pure and Applied Chemistry
IXR - ion exchange resin
JP 4 - jet fuel 4
K - potassium
Kd - partitioning coefficient between solid phase media and groundwater
KOC - organic carbon partition coefficient
LC - liquid chromatography
LDPE - low-density polyethylene
LLOQ - lower limit of quantitation
LOQ - limit of quantitation
MDEQ - Michigan Department of Environmental Quality
Mg - magnesium
MTP - molecularly imprinted polymer
MNA - monitored natural attenuation
MRM - multiple reaction monitoring
MS - mass spectrometry
MTBE - methyl tert-butyl ether
MWCO - molecular weight cut-off
Na - sodium
Na+ - sodium ion
NAPL - non-aqueous phase liquid
NF - nanofiltration
NFDHA - nonafluoro-3,6-dioxaheptanoic acid
N-MeFOSAA - N-methylperfluorooctane sulfonamidoacetic acid
NOM - natural organic matter
NTU - nephelometric turbidity units
nZVI - nano-zerovalent iron
OC - organic carbon
OECD - Organization for Economic Co-operation and Development
OH" - hydroxyl anion
ORP - oxidation-reduction potential
OW - Office of Water (US EPA)
PAC - powdered activated carbon
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PBSF - perfluorobutanesulfonyl fluoride
PCE - perchloroethylene (tetrachloroethyene)
PFAA - Perfluorinated alkyl acids
PFAS - per- and polyfluoroalkyl substances
PFBA - perfluorobutyric acid
PFBS - Perfluorobutane sulfonic acid, perfluorobutanesulfonate
PFC - perfluorocarbon
PFCA - perfluorocarboxylic acid
PFDA - perfluorodecanoic acid
PFDS - perfluorodecanesulfonic acid
PFEI - pentafluoroethyl iodide
PFHxS - perfluorohexanesulfonic acid, perfluorohexane sulfonate
PFNA - perfluorononanoic acid
PFOA - perfluorooctanoate
PFOS - perfluorooctane sulfonate
PFPA - perfluorooctane sulfonamide
PFPrA - perfluoropropanoic acid
PFSA - perfluorosulfonic acid
PIGE - particle-induced gamma emission
pKa - acid dissociation constant
PMS - peroxymonosulfate
POSF - perfluorooctanesulfonyl fluoride
PPE - personal protective equipment
PRB - permeable reactive barrier
PTFE - polytetrafluoroethylene
PVC - polyvinyl chloride
PVDF - polyvinylidene fluoride
QA/QC - quality assurance / quality control
RO - reverse osmosis
SAP - sampling and analysis plan
SC - specific conductance
SERDP - Strategic Environmental Research and Development Program
SO32" - sulfite anion
SOP - standard operating procedure
SPLP - synthetic precipitation leaching procedure
SSL - site screening level
TCE - trichloroethylene
TFC - thin-film composite
TFE - tetrafluoroethylene
Ti - titanium
TOFA - total organofluorine assay
TOF-CIC (or TOF) - total organofluorine-combustion ion chromatography
TOP - total oxidizable precursors assay
UF - ultrafiltration
UFP-QAPP - Uniform Federal Policy for Quality Assurance Project Plan
US EPA - United States Environmental Protection Agency
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UV - ultraviolet
VOC - volatile organic compound
VUV - vacuum ultraviolet
WWTP - wastewater treatment plant
Zn - zinc
ZVI - zerovalent iron
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EXECUTIVE SUMMARY
The purpose of this document is to introduce topics relevant to management of
groundwater contaminated with per- and poly-fluoroalkyl substances (PFAS).
Perfluorinated substances are based on a fully fluorinated carbon chain and
polyfluorinated substances are based on a partially fluorinated carbon chain. A modern
definition of PFAS is:
PFAS are defined as fluorinated substances that contain at least one fully
fluorinated methyl or methylene carbon atom (without any H/Cl/Br/I atom
attached to it), i.e., with a few noted exceptions, any chemical with at least
a perfluorinated methyl group (-CF3) or a perfluorinated methylene
group (-CF2-) is a PFAS.
The simplest PFAS to describe are the perfluoroalkyl acids (PFAAs). These prototypical
PFAS are comprised of a linear carbon-chain with most or all the bonding sites occupied
by a fluorine (F) atom and one of the functional groups from organic chemistry attached
to one end. Functional groups can also be attached at any branches in the carbon chain.
The fluorinated carbon-chain is referred to as the backbone, the fluorinated end of the
backbone is referred to as the tail, and the functional group at the opposite end of the
backbone is referred to as the head. The backbone and tail are physically durable,
thermally stable, chemically inert, and water- and oil-repellant (hydrophobic/oleophobic).
If an attached functional group is hydrophilic then the PFAS is amphiphilic and can
function as a surfactant.
PFAS manufacturing methods provide useful information for identifying sources. They
are predominantly manufactured using either electrochemical fluorination (ECF) or
telomerization. ECF produces many impurities with branched backbones, or even cyclic
backbones, and has been the only process used to create the perfluorosulfonic acid
(PFSA) subclass of PFAA. Telomerization has been used to produce the
perfluorocarboxylic acid (PFCA) subclass of PFAA.
There are several classification systems for PFAS. The most basic division of PFAS is
polymer versus non-polymer. PFAS are also classified as short chain versus long chain.
PFCA with eight or more carbons (seven or more carbons are perfluorinated) and PFSA
with six or more carbons (six or more carbons are perfluorinated) are considered long
chain PFAS by US EPA.
There are thousands of PFAS because they have a myriad of applications. Some
examples are use in aqueous fire-fighting foams (AFFF), fabric protectants such as
Scotchgard™, coatings on nonstick kitchen cookware and food packaging, papermaking,
oil production, mining, metal plating, electronics, and additives to cleansers, polishes,
waterproofing agents, tanning agents, wax, lubricants, ink, and paint.
The production and widespread use of products containing PFAS have resulted in their
introduction to the environment through releases from primary and secondary
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manufacturing facilities, releases from industrial and manufacturing facilities, releases of
products containing PFAS such as aqueous fire-fighting foams (AFFF) at airports and
military bases, escape of landfill leachate, release of wastewater treatment plant (WWTP)
effluent to water bodies, irrigation using WWTP effluent, and land application of WWTP
biosolids and sludges for soil improvement. PFAS are present in environmental media
including air, soil, sediment, surface water, and groundwater, and in locations as remote
as the Arctic. PFAS are of concern because they bioaccumulate in plants and animals and
they have been linked to conditions such as low birth weight, thyroid hormone disruption,
low activity sperm, elevated cholesterol, diabetes, and cancer. Regulatory thresholds are
presently in the parts per trillion range but vary between government entities and are still
evolving.
Some PFAS are known as "precursors" because they are susceptible to transformation in
the environment. Precursors include non-fully fluorinated PFAS species that contain
carbon-hydrogen (C-H) or carbon-oxygen (C-O) bonds such as fluorotelomer alcohols
(FTOHs) and fluorotelomer sulfonates (FTSAs). Precursors also include side-chain
fluorinated polymers with the potential to form PFCAs. Many can degrade either
biotically or abiotically in the subsurface to form very environmentally stable
endmembers such as the PFAAs, although they may pass through intermediary daughter
products. PFAAs include the widely studied PFAS perfluorooctanoate (PFOA) and
perfluorooctane sulfonate (PFOS). Precursors tend to be more common in shallow soils
and PFAAs tend to be more common in groundwater, but the relative location of PFAS
that are intermediate breakdown products of precursors is less predictable.
Subsurface transport of PFAS is not yet fully understood, but it is influenced by the
physiochemical characteristics of the environmental media, individual PFAS species, and
co-contaminants in both dissolved- and separate-phases. Their subsurface transport is
retarded by adsorption of their tails and backbone to other hydrophobic materials such as
organic matter (organic carbon). This process is termed hydrophobic interaction. PFAS
with one or more polar functional groups are amphiphilic and behave as surfactants. In
addition to hydrophobic interactions, subsurface transport of amphiphilic PFAS is
retarded by accumulation at media interfaces and electrostatic interactions with charged
surfaces such as some clays and minerals. Longer-chain and straight-chain PFAS tend to
be more retarded than short-chain and branched isomers, apparently due to the greater
hydrophobicity of longer and straight chains and possibly greater hydrophilicity
conferred by any additional polar functional groups capping branches. Cationic and
zwitterionic PFAS are more retarded that anionic PFAS. The mechanisms by which non-
PFAS can interact with PFAS to alter subsurface transformations and transport are not
well understood, but adsorption to NAPL appears to be more significant that absorption
into NAPL. The complexities of PFAS retardation and the difficulty of quantitating the
contributions render the use of the traditional methods for estimating retardation of
hydrophobic organic contaminants unsuitable for PFAS.
Developing a site conceptual model (CSM) for a PFAS site can be challenging because
their behavior in the environment is not fully understood and available laboratory
analyses can only identify a small fraction of the individual compounds. Consequent of
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the state of knowledge about the environmental behavior of PFAS, site characterization
should include assessment for both total and specific PFAS, all potential co-contaminants
in both dissolved and separate-phase form, media characteristics, and groundwater
geochemistry. The atypical retardation of PFAS is directly affected by media
characteristics such as grain size, organic carbon, clays and minerals, pH, and cation and
anion exchange capacity (AEC). Assessing PFAS retardation and the representativeness
of water samples also necessitates characterization of parameters such as individual
dissolved ions, alkalinity, turbidity, pH, oxidation-reduction potential (ORP), specific
conductance (SC), and dissolved oxygen (DO).
Special consideration is needed when planning and performing field activities at PFAS
sites because of their broad use in manufactured goods and their widespread occurrence,
including the ground surface due to atmospheric deposition. Sampling materials and
equipment potentially containing PFAS, or fluorinated materials, should be avoided. Trip
blanks and equipment blanks must always be included in the sampling and analysis
program.
The US EPA has validated four analytical methods for measuring PFAS in water
samples. The US EPA SW-846 Method 8327 quantifies a sampling of 24 PFAS analytes
in non-potable water such as wastewater. US EPA Method 537.1 quantifies 18 PFAS in
drinking water, and US EPA Method 533 focuses on "short chain" PFAS and quantifies
25 PFAS in drinking water, 11 of which are not included in Method 537.1. US EPA
Method 1633 quantifies a larger sampling of 40 PFAS species in wastewater, surface
water, groundwater, soil, biosolids, sediment, landfill leachate, and fish tissue. Other
analytical methods for individual and total PFAS are available or in development but
have not yet been approved by US EPA.
PFAS typically do not respond well to traditional remedial techniques, such as chemical
oxidation or bioremediation, due to their unusual properties such as the armoring of the
carbon-chain backbone by very strongly bonded fluorine atoms. Currently, ex-situ
granular activated carbon (GAC) and ion exchange resins (IXR) are the primary means of
removal for PFAS from groundwater. However, these materials have limitations such as
interference by co-contaminants, preferential adsorption of some PFAS species or limited
adsorption ranges, and issues with management of spent media. Reverse osmosis (RO)
and nanofiltration (NF) are also proven technologies for PFAS separation but have shown
to be comparatively expensive due to high energy consumption. Innovative methods and
adaptions of existing methods are being explored.
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INTRODUCTION
The purpose of this document is to provide an overview of topics relevant to management
of groundwater contaminated with per- and poly-fluoroalkyl substances (PFAS).
Perfluorinated PFAS are based on a fully fluorinated carbon chain and polyfluorinated
PFAS are based on a partially fluorinated carbon chain. Broader overviews for PFAS are
provided by Meegoda et al., (2020) and Evich et al., (2022), and a very thorough but
lengthy treatment is provided by ITRC, (2022). Sima and Jaffe, (2021) list recent reviews
focused on subtopics including occurrence, fate, migration, and remediation.
Description
PFAS are a family of thousands of man-made organic chemicals (OECD, 2018). The
widely used definition of PFAS provided by Buck et al., (2011) has recently been
superseded by an update intended to encompass a broader range of structurally related
chemicals (ODEC, 2021):
PFAS are defined as fluorinated substances that contain at least one fully
fluorinated methyl or methylene carbon atom (without any H/Cl/Br/I atom
attached to it), i.e., with a few noted exceptions, any chemical with at least
a perfluorinated methyl group (-CF3) or a perfluorinated methylene
group (-CF2-) is a PFAS.
The simplest PFAS to describe are the perfluoroalkyl acids (PFAAs). These prototypical
PFAS are comprised of a linear carbon-chain with most or all the bonding sites occupied
by a fluorine atom and one of the functional groups from organic chemistry attached to
one end. If the carbon chain is branched, functional groups can also be attached there.
The fluorinated carbon-chain is referred to as the backbone, the fluorinated end of the
backbone is referred to as the tail, and the functional group at the opposite end of the
backbone is referred to as the head. Example PFAS are illustrated on Figure 1.
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Manufacturing
Knowledge of PFAS synthesis methods and manufacturing history are useful for the
developing practice of source identification and separation, known as PFAS forensics.
PFAS forensics is a challenging process, and consequently still an active area of research.
Synthesis
PFAS can be synthesized from perfluoroalkyl building blocks via ECF, telomerization (a
chain-transfer oligomerization), other methods of oligomerization (e.g., radical
oligomerization, oligocondensation, ionic oligomerization and ring-opening reactions),
direct fluorination, or photooxidation (Buck et al., 2011; Savu, 2000; Dams and Hintzer,
2016). Only ECF and telomerization are commonly known in the U.S. ECF, also known
as the "Simons process" (3M 1999; Kempisty et al., 2018), typically proceeds by
electrolysis in liquid anhydrous hydrogen fluoride (HF) of an acyclic, hydrogen-saturated
hydrocarbon with a functional group attached to one end. The ECF process replaces the
hydrogen atoms bonded to the carbon-chain with fluorine atoms and under suitable
reaction conditions all the hydrogen will be replaced by fluorine (3M, 1999). A common
example of ECF is treatment of 1-octanesulfonyl fluoride to form
perfluorooctanesulfonyl fluoride (POSF), which is used as a feedstock to produce final
products such as PFOS. Another example is ECF of octanoyl fluoride to ultimately
synthesize PFOA.
Telomerization typically reacts a perfluoroalkyl iodide telomer such as pentafluoroethyl
iodide (PFEI) with a taxogen such as tetrafluoroethylene (TFE) to yield longer chain
perfluoroalkyl iodides (Buck et al., 2011). The perfluoroalkyl iodides are then reacted
with a chemical such as ethylene to form further lengthened fluorotelomer iodides. The
fluorotelomer iodides are in turn used as reaction intermediaries to eventually produce
final products such as PFOA, fluorotelomer surfactants, and polymer products.
Impurities
PFAS source identification is facilitated by the fact that ECF and telomerization
substantially differ by the characteristics of the PFAS building blocks and suites of
impurities they produce (Buck et al., 2011). ECF often causes fragmentation and
rearrangement of the starting carbon-chain, which results in creation of impurities
including homologs of varying chain lengths, up to 20% to 30% branched isomers, some
cyclic structures, and perfluorocarbons (PFAS without a functional group, or PFCs) (3M,
1999). Consequently, PFAS mixtures produced from ECF have constituents with both
odd and even numbers of carbons in their backbones. In contrast, telomerization does not
produce branched isomers of PFAS unless a branched telomer is intentionally used, and
only produces a small fraction of odd-numbered carbon-chains, if any. However, like
ECF, telomerization can produce PFAS that are not entirely chemical inert in the
environment which can be transformed to other PFAS with odd-numbered carbon chains.
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Products
PFAS source identification is also facilitated because PFSAs such as PFOS were only
produced using building blocks generated using ECF. Also, ECF can produce both
perfluorinated and polyfluorinated compounds whereas telomerization produces
polyfluorinated fluorotelomers, although these can be used as feedstocks to produce
perfluorinated products (3M, 1999; ITRC, 2022). The products from both synthesis
methods can be used as feedstocks to produce the most studied group of PFAS, the
PFAAs. However, as discussed below, only ECF is known to have been used to produce
both PFAA subcategories, the PFCAs and the (PFSAs). Telomerization is known to have
been used to manufacture PFCAs but not PFSAs. Thus, presence of PFSAs such as PFOS
is an indication of source material produced by ECF.
History
PFAS have been in continuous commercial production from nearly the time they were
first discovered until the present. The first polymerizations of fluoroethenes, such as
polytetrafluoroethylene (PTFE), were studied at IG-Farbenindustrie in Germany during
the early 1930's (Wetzel, 2005) and the first patent application for a fluoropolymer was
filed in 1934 by Schloffer and Scherer, (1934). PTFE was subsequently synthesized by a
DuPont de Nemours (DuPont) chemist investigating fluorinated refrigerants in 1938
(Science History Institute, 2017). Processes to commercially produce PFAS using ECF
were developed in the 1940s. DuPont commercialized PTFE in 1946 under the name
Teflon™ (Ebnesajjad, 2000). The Minnesota Mining and Manufacturing Company (3M)
licensed the Simons ECF process from Dr. Simons of Penn State University in 1945 (3M,
1999) and reportedly began production of PFOA in 1947 (Prevedouros et al., 2006),
although they report they built the first manufacturing-scale pilot ECF process in 1949
(3M, 1999). 3M began industrial-scale production via ECF of PFOA, PFOS, and PFAS
containing products such as Scotchgard™ in the 1950s (Banks et al., 1994) and it is
widely reported that DuPont began purchasing PFOA from 3M circa 1951 to improve
their production of Teflon™. Telomerization was later invented by DuPont (Munoz et al.,
2019) and its use to produce PFOA began in the 1970s, ameliorating DuPont's need to
purchase PFOA from 3M for manufacturing Teflon™.
Environmental concerns brought changes to PFAS manufacturing. During early 2000 3M
was the sole producer of perfluorooctane sulfonic acid (PFOS) in the U.S. (US EPA,
2003). However, early in the decade and at the urging of US EPA, 3M began phasing out
ECF production of perfluorooctanoic acid (PFOA) and POSF based PFAS, including
perfluorohexane sulfonate (PFHxS), PFOS, perfluorodecanesulfonic acid (PFDS), and
related compounds (Buck et al., 2011). The phase-out was reportedly completed by 2008,
but due to the purported lower toxicity of short-chain PFAS 3M continued to produce the
shorter-chain perfluorobutanesulfonyl fluoride (PBSF)-based PFAS such as
perfluorobutane sulfonic acid (PFBS) via ECF. PFAS production by ECF also continues
in other countries, including China, India, and Russia (ITRC, 2022). PFOA production by
telomerization ramped up for a period by other producers in the U.S. after the 3M
phaseout, but production of PFOA, some longer chain homologs, and some related
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compounds was phased out by most producers by 2015 (US EPA, 2017). Replacements
for PFOA include the fluoroalkylether carboxylates GenX and Adona, and a replacement
for PFOS is the chlorinated polyfluoroalkyl ether sulfonate F53-B (Munoz et al., 2019).
GenX is a trade name for the DuPont spinoff Chemours' process to manufacture
fluoropolymers without the use of PFOA, typically hexafluoropropylene oxide dimer acid
[HFPO-DA], Adona is 3M's trade name for dodecafluoro-3H-4,8-dioxanonanoate. F53-B
has been used in China since the 1970s and has been reported as the trade name for
chlorinated polyfluorinated ether sulfonate, 9-chlorohexadecafluoro-3-oxanone-l-
sulfonic acid, and ll-chlororeicosafluoro-3-oxaundecane-l-sulfonic acid.
The types of PFAS manufactured expanded over time as new uses were found, eventually
including fluoropolymers, fluorotelomer-based substances, perfluoro(poly)ether-based
substances, perfluoroalkane sulfonyl-based substances, perfluoroalkane carbonyl-based
substances, and polyfluorinated substances (Buck et al., 2011; OECD, 2018; Gliige,
2020). There may be others, as some PFAS mixtures are trade secrets. A list of PFAS that
are known to have been manufactured was produced by OECD, (2007) and US EPA
maintains a database of PFAS accessible on the web at: CompTox Chemicals Dashboard
(epa.gov).
Classification
Several classification systems have been developed for PFAS, including polymer versus
non-polymer, partially fluorinated "poly-" versus fully fluorinated "per-" backbone, and
combinations of these. A combined classification system for "environmentally relevant"
PFAS was presented by Buck et al., (2011) in a paper that also attempted to standardize
PFAS nomenclature. A recent publication by OECD, (2021) provided recommendations
on reconciling terminology for PFAS, and their updated definition of PFAS included
forms with side-chain aromatics.
Polymers
Polymer PFAS were divided by Buck et al., (2011) into three primary groups:
fluoropolymers, polymeric perfluoropolyethers, and side-chain fluorinated polymers.
Fluoropolymers have a carbon-only backbone but perfluoropolyethers have some oxygen
atoms included in the carbon backbone. Side-chain fluorinated polymers have non-carbon
substitutions included in the backbone and can be further subdivided into fluorinated
acrylates and methacrylates, fluorinated urethanes, and fluorinated oxetanes.
Non-polymers
Non-polymer PFAS were divided by Buck et al., (2011) into perfluoroalkyl and
polyfluoroalkyl substances. They further classified perfluoroalkyl substances into
aliphatic (typically straight-chain) perfluorocarbons, PFAAs, perfluoroalkane sulfonyl
sulfides, perfluoroalkane sulfonomides, perfluoroalkane iodides, and perfluoroalkane
aldehydes. The polyfluoroalkyl substances were subclassified into perfluoroalkane
sulfonamido derivatives, fluorotelomer-based compounds, and semifluorinated //-alkanes
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and alkenes. The PFAAs have long been the primary focus of environmental research on
PFAS. The US EPA defines short- versus long-chain PFCA as having less than eight
carbon atoms and short- versus long-chain PFSA as having less than six carbon atoms.
Other definitions based on chain length, such as those used in Europe, can vary slightly.
Manufacturing Method
Evich, et al. (2022) provided lists of PFAS groups based on manufacturing method.
PFAS groups from direct fluorination included PFCAs, PFSAs, hydrofluorocarbons,
hydrofluoroethers, hydrochlorofluoroolefins, hydrofluoroolefins, side-chain fluorinated
aromatics, perfluoroalkyl-tert-amines, and perfluoroalkanoyl/perfluoroalkanesulfonyl
fluorides. PFAS manufactured using oligomerization (telomerization) included
fluoropolymers, perfluoropolyethers, fluorotelomers, perfluoroalkyl (ether) carboxylic
and sulfonic acids, and perfluoroalkene derivatives.
Properties
The physiochemical properties of PFAS vary and are difficult to directly measure (Wang
et al., 2021; ITRC, 2022). Example PFAS and physiochemical properties are provided in
Table 1. Most PFAS have low volatility and at room temperature are solid, although
shorter-chain PFAS may be liquid.
Backbone and Hydrophobic Interactions
The strong C-F chemical bonds and armoring of the carbon-chain by highly
electronegative fluorine atoms render the backbone and tail of PFAS physically durable,
thermally stable, chemically inert, nonpolar, and water- and oil/lipid- repellent
(hydrophobic/oleophobic) (O'Hagan, 2008; Gagliano et al., 2020)1. Hydrophobic
materials such as PFAS tails and organic matter ("organic carbon", or OC) appear to be
attracted when immersed in a polar solvent such as water. The apparent attraction is the
result of the tendency for hydrophobic areas on molecules to avoid contact with the
solvent and is termed "hydrophobic interaction". The hydrophobicity and propensity for
hydrophobic interactions of a PFAS backbone increase with increasing carbon-chain
length and decrease with increasing carbon-chain branching (Park et al., 2019).
Functional Groups and Electrostatic Interactions
Functional groups such as carboxylates, sulfonates, sulfates, phosphates, betaines, or
amines are added to a fluorinated carbon-chain backbone to create a finished PFAS
molecule. Functional groups can render PFAS nonionic, anionic, cationic, or zwitterionic
(Xiao et al., 2019). Ionic functional groups confer the ability to undergo electrostatic
1 There are studies that have demonstrated some PFAS can partition to the phospholipid bilayers of
bacteria, indicating that they may not exhibit lipophobic (oleophobic) tendencies when dissolved in water
(Fitzgerald et al., 2018).
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interactions with other charged materials to PFAS. However, electrostatic interactions are
significantly more influenced than hydrophilic interactions by the physiochemical
properties of the solid matrix materials and the geochemistry of the water solution.
Partitioning to Media Interfaces
The existence of both hydrophobic and hydrophilic regions on a molecule renders it
amphiphilic and confers surfactant properties. Surfactants can be divided into nonionic,
anionic, cationic, and zwitterionic classes and there are PFAS of each of these types. All
types of surfactants reduce surface tension by adsorbing to the interface between two
phases such as air and water. The surface tension typically decreases nonlinearly at low
surfactant concentrations but otherwise decreases log-linearly until it plateaus near the
concentration at which micelles (bubbles) or hemimicelles (bubbles attached to a surface)
can form. This concentration is termed the critical micelle concentration (CMC) and is
associated with the concentration at which the surfactant has reached saturation at the
fluid interface. Additional surfactant merely creates more micelles. The CMC is
characteristic of individual surfactants but varies as function of surfactant type,
temperature, the alkyl chain length, the ionic head group, and the strength of the binding
between the head group and the electrolytes/counterions (complementary ions) available
in solution. Surfactants have a minimum temperature termed the critical micelle
temperature (CMT) or Krafft temperature, below which micelles will not form.
The CMC generally decreases with increasing alkyl chain length but increases with
increasing hydrophilicity of the polar head group. Addition of electrolytes such as salt
decreases the CMC and addition of alcohol increases the CMC.
Amphiphilic PFAS accumulate at media interfaces, such as between air and water, non-
aqueous phase liquids (NAPL) and water, or soil and water. The propensity for
accumulation at media interfaces is a concentration-dependent function of the change in
surface tension between two fluids, but generally increases with increasing backbone
length (Psillakis et al., 2009). Also, the surface tension and affinity for media interfaces
decrease with increasing PFOA and PFOS concentrations in a nonlinear way, possibly
indicating that factors other than PFAS concentration (e.g., solution pH and ionic
strength) affect surface tension (Costanza et al., 2019).
Environmental Transformation of Precursors
PFAS that can undergo transformations in the environment are known as precursors.
Precursors are non-fully fluorinated PFAS species that contain C-H and potentially
carbon-oxygen (C-O) bonds. Precursor transformations may be natural or influenced by
various remedial actions (McGuire et al., 2014). Transformations can occur abiotically or
biogenically, aboveground or belowground, and aerobically or anaerobically, but cannot
mineralize PFAS because of chemically inert sections within their backbone (ITRC,
2022). Precursors transform either directly or through intermediary daughter products
into "terminal" or "endmember" PFAS that are very environmentally stable or fully inert
(Suthersan et al., 2016). For example, polymer PFAS such as (FTOH) and (FTSA) can be
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transformed into PFAA2. Proposed reaction pathways vary but are an area of active
research. Biotransformation of PFAS was recently reviewed by Zhang et al., (2022).
Uses
PFAS have a myriad of applications. Over 200 use and sub-use categories were identified
by Gliige et al., (2020). Example uses include Class B aqueous fire-fighting foams
(AFFF), fabric protectants such as Scotchgard™, coatings on nonstick kitchen cookware
and food packaging, papermaking, oil production, mining, metal plating, electronics, and
additives to cleansers, polishes, waterproofing agents, tanning agents, wax, lubricants,
ink, and paint (Moody and Field, 2000; Prevedouros et al., 2006; Paul et al., 2009; Wang
et al., 2017; Tokranov et al., 2019; OECD, 2020).
Aqueous Fire-Fighting Foam (AFFF)
AFFF use accounts for many of the sites where PFAS have been released to the
subsurface environment and migrated to groundwater. AFFF are associated with complex
contaminant mixtures in soil and groundwater at facilities such as chemical plants,
petroleum refineries, airports, and military installations. Groundwater impacted by PFAS
has the greatest concentrations at AFFF sites (Backe et al., 2013).
Composition
AFFF formulations are mostly proprietary (Place and Field, 2012) but they are known to
contain hydrocarbon surfactants such as sodium alkyl sulfate and PFAS surfactants such
as carboxylates, perfluoroalkyl sulfonates, perfluoro betaines, perfluoro sulfonamides,
perfluoro sulfonamidoethanol, perfluoro thioamido amino carboxylates, perfluoro
sulfonamido amines, and fluorotelemer sulfonates (Suthersan et al., 2016). They also
contain materials such as magnesium sulfate, sodium octyl sulfate, sodium decyl sulfate,
ethylene glycol, propylene glycol t-butyl ether, diethylene glycol monobutyl ether, and
ethanol. Polyfluorinated precursor compounds are a major fraction of the PFAS detected
in vadose soils at AFFF release sites (Sharifan, 2021).
Legacy AFFF
AFFF manufacturers and formulations have changed over time (Kempisty et al., 2018).
After their development by the U.S. Navy in the 1960s 3M manufactured a "legacy"
(manufactured pre-2001) PFOS-based AFFF branded Lightwater™ from the late 1960s
through 2002 (DoD, 2014). Lightwater™ contained PFOS and several precursors that
could break down to PFCAs such as PFOA (Backe et al., 2013). Other brands of
fluorotelomer foams were manufactured from the 1970s through 2016 that were
comprised mostly of C6-PFAS, some longer-chained PFAS, and polyfluorinated
2 Until very recently it was thought that PFAAs were completely recalcitrant in the environment
(Prevedouros et al., 2006; Ferrey et al., 2012), but recent studies raise the possibility that this might not
be the case (e.g., Huang and Jaffe, 2019).
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precursors but did not contain PFOA except as a potential impurity (Schultz et al., 2004;
Place and Field, 2012; Backe et al., 2013). Not unlike Lightwater™, however, their
impurities could break down to PFOA and other PFCAs and as such, are considered
precursors (Weiner et al., 2013; Harding-Majanovic et al., 2015). Annunziato et al.,
(2020) identified more than 100 PFAS in a legacy foam sample dominated by PFOS
(31%). Other specific PFAS identified were perfluorohexanesulfonic acid (PFHxS) (5%)
and a mixture of other PFAS totaling 3%.
Modern fluorotelomer AFFF
Fluorotelomer foams have been in use since the 1970s and after 2001 became the
dominant foam (TRB, 2017; ITRC, 2022) because long-chain foam manufacture (by
ECF) was discontinued. These newer AFFF contain shorter chain (predominantly C6)
fluorotelomers that are expected to be less bioaccumulative and toxic but may still
contain trace amounts of PFOA and PFOA precursors (Scheringer et al., 2014; US EPA,
2018). Nonetheless, a review by Ateia et al., (2019) found that the newer short-chain
(four to seven carbon) and ultra-short-chain (two to three carbon) PFAS, like their longer
chain homologs, are environmentally persistent and can break down to form stable PFCA
and PFSA end products (Hurley et al., 2004; Renner, 2006; Lee et al., 2010a; Liou et al.,
2010; Ritter, 2010; Butt et al., 2014). Also, the newer short-chain PFAS are typically
used in greater concentrations because they are somewhat less effective for fire-fighting
applications than longer chain PFAS (Lindstrom et al., 2011).
Alternative formulations
There are alternatives to traditional PFAS-containing AFFF formulations, so it is possible
that some AFFF release sites may not be impacted by PFAS. Fire-fighting foams are
grouped into several classes that have distinct characteristics which may be helpful when
conducting release forensics (Chemguard, 2005):
• Traditional AFFF - contain synthetic foaming agents (hydrocarbon surfactants),
solvents, fluorochemical surfactants (PFAS), stabilizers, and salts.
• Alcohol-resistant AFFF (AR-AFFF) - comprised of a traditional AFFF and a high
molecular weight polymer.
• Synthetic detergents - comprised of hydrocarbon surfactants and solvents.
• Wetting agents - similar to Class A foams.
• Protein - based on a hydrolyzed protein mixed with foam stabilizers and
preservatives.
• Fluoroproteins - contain protein and fluorocarbon surfactants.
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• Foam Forming Fluoroprotein (FFFP) - a derivative of AFFF and fluoroprotein, based
on fluoroprotein formulations with increased amounts of fluorocarbon surfactants.
Environmental Occurrence
The production and use of products containing PFAS has resulted in their introduction to
the environment through releases from primary and secondary manufacturing facilities,
releases from industrial and manufacturing facilities, use of PFAS products such as AFFF
at airports and military bases, escape of landfill leachate, release of WWTP effluent to
water bodies, irrigation using WWTP effluent, and land application of WWTP biosolids
and sludges for soil improvement (Wang et al., 2017; OECD, 2018; Gliige, 2020;
Sharifan et al., 2021). PFAS occur in environmental media including air, soil, sediment,
surface water, and groundwater, and in locations as remote as the Arctic (Young et al.,
2007).
Toxicity and Regulation
Environmental management of PFAS is motivated by their propensity for
bioaccumulation and their high toxicity. PFAS bioaccumulate in both plants and animals
and are toxic to both animals and humans (OECD, 2002; D'Hollander et al., 2010; US
EPA, 2016a and b; OECD, 2018; AT SDR, 2021; ITRC, 2022). Branched PFAS isomers
preferentially bioaccumulate in humans and linear isomers preferentially bioaccumulate
in most other animal species (Schulz et al., 2020), but both types of isomers preferentially
adsorb to albumin in blood as well as other proteins (Forsthuber et al., 2020). PFAS have
been linked to low birth weight, thyroid hormone disruption, low activity sperm, elevated
cholesterol, diabetes, and cancer (Wang et al., 2013, Domingo and Nadal, 2019).
Permissible exposures to PFAS in the U.S. vary across government entities and are
expected to continue evolving as more information accumulates. Although there are
many chemical groups of PFAS the most widely studied have been the PFAAs. These
include the PFCAs such as PFOA, and PFSAs such as PFOS. Permissible human
exposures to PFAS have been developed based on exposure routes such as air, airborne
dust, drinking water, food, food-contact materials (boxes, papers, and wrappers), and
breast milk (D'Hollander et al., 2010), and available toxicity evaluations for PFAS are
detailed in ASTDR, (2021). Based on the information available at the time, in 2016 the
US EPA issued a Lifetime Health Advisory of 70 parts per trillion (ppt) for PFOA and
PFOS in drinking water, individually or combined (US EPA, 2016a and b). The US EPA
also provides screening levels for its Regions available at:
https://www.cpa. gov/ri sk/regional-screening-1 evel s-rsl s-generic-tables, which include
levels for PFBS. The risk-based SSL for PFBS in groundwater is 1,900 ppt and for
K+PFBS in groundwater is 3,000 ppt. In 2021 the US EPA issued final Human Health
Toxicity Values for PFBS and its potassium salt, potassium perfluorobutane sulfonate
(K+PFBS) (US EPA, 2021a). This document indicates that PFBS and K+PFBS are
almost an order of magnitude less potentially harmful than PFOA and PFOS.3
3 This may be because PFBS have a shorter retention time in the body than PFOS.
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The 2016 advisories were replaced in June 2022 by interim advisories issued by the US
EPA Office of Water (OW) (US EPA, 2022a) for PFOA and its replacement GenX, and
PFOS and its replacement PFBS. The new, interim drinking water advisories are 0.004
ppt for PFOA, 10 ppt for GenX, 0.02 ppt for PFOS, and 2,000 ppt for PFBS
(https://www.epa.gov/svstem/files/documents/2Q22-Q6/drinking-water-ha-pfas-factsheet-
communities.pdf).
Environmental Fate and Transport
The fate and transport of PFAS in the subsurface are influenced by the physiochemical
characteristics of the individual PFAS species and the site's characteristics. These
properties combine to influence PFAS surface chemistry, surfactant properties, solubility,
sorption, stability, interactions with solvents, and other properties and behaviors. ITRC,
(2022) indicated that relevant PFAS characteristics included chain-length, functional
groups and their charge state, and extent of fluorination. They also indicated relevant site
characteristics may include atmospheric conditions and precipitation, surface water and
groundwater flow rates, soil permeability, surface charge, soil and sediment organic
carbon (OC) content, pH, anion exchange capacity (AEC), cation exchange capacity
(CEC), mineralogy, water content, depth to groundwater, pH, redox conditions, presence
of co-contaminants such as fuel hydrocarbons and halogenated solvents, competitive
inhibition amongst PFAS, presence of non-PFAS surfactants and stabilizers typically
included in AFFF formulations, and presence of NAPL. However, recent work in the
literature indicates that some contributing factors probably need to be subdivided to
establish their relationships with PFAS transport, such as measuring individual ion
concentrations rather than solution ionic strength (Pereira et al., 2018).
Subsurface Transformations
Transformations of precursor PFAS have been previously discussed. They can be
biotically or abiotically mediated and occur under aerobic and anaerobic conditions, but
proposed reaction pathways are still being evaluated.
Vadose Zone
Most PFAS releases are to the ground surface, with the exception of landfill leachate.
Substantial concentrations of PFAS often remain beneath the location of the release in the
vadose zone and represent a long-term source for groundwater contamination (Brusseau
and Van Glubt, 2019; Barzen-Hanson et al., 2017).
PFAS entering the vadose zone beneath a source area can volatilize, be absorbed by
biota, accumulate at media interfaces, adsorb to media solids and NAPL via hydrophobic
or electrostatic interactions, transform, or leach to groundwater (Sharifan et al., 2021).
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These processes are influenced by competitive inhibition of adsorption, ion exchange,
and the presence of both dissolved and NAPL4 co-contaminants.
Although mixtures of PFAS with varying chain lengths and branching have been
observed in shallow soil horizons (Sepulvado et al., 2011; Xiao et al., 2015; Hale et al.,
2017; Nickerson et al., 2020) there are trends in vertical stratification. Generally, longer
chain, linear, and cationic or zwitterionic PFAS are retained more strongly and occur at
greater concentrations shallow in the vadose zone. In contrast, shorter-chain, non-linear,
and anionic PFAS are more mobile, occur deeper in the vadose zone, and can more easily
migrate into the phreatic zone (Backe et al., 2013; Hatton et al., 2018; Li et al., 2018;
Wang et al., 2021; Adamson et al., 2022). Also, precursors tend to be more common in
shallow soils (Sharifan et al., 2021) and their end products tend to be more common in
groundwater. However, groundwater concentrations of intermediate products are
sometimes comparable to concentrations of end products (Nickerson et al., 2021).
Most work characterizing the retardation of PFAS in soil has focused on anionic species
that exist in a charged state within the normal range of subsurface pH. However, Xiao et
al., (2019) characterized the subsurface behavior of cationic and zwitterionic PFAS
relative to that of anionic PFAS. They found:
• Cationic and zwitterionic PFAS were more retarded than anionic and neutral PFAS,
although the differences diminished as concentrations increased.
• Hydrophobicity was a poorer predictor of sorption for cationic and zwitterionic PFAS
than anionic PFAS.
• Sorption of cationic and zwitterionic PFAS was highly nonlinear.
• Sorption of cationic and zwitterionic PFAS apparently increased as their
concentrations decreased.
• Sorption of cationic and zwitterionic of PFAS at low concentrations is dominated by
electrostatic interactions with negatively charged soil constituents.
• Sorption of cationic and zwitterionic of PFAS at high concentrations is dominated by
hydrophobic interactions.
• Sorption of cationic PFAS is highly correlated to OC and is reversible.
• Sorption of zwitterionic PFAS at low OC showed concentration-dependent hysteresis,
indicative of adsorption irreversibility.
4 Although adsorption at NAPL interfaces is significant, recent work indicates partitioning into bulk NAPL is
probably not a significant concern (Glubt and Brusseau, 2021).
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• The maximum electrostatic potential of PFAS ions, computed using density
functional theory, was found to be a useful predictor of the sorption of ionic PFAS
species.
Phreatic Zone
Many PFAS fate and transport processes of the vadose zone are of less concern in the
phreatic zone. Volatilization is primarily a concern in vadose soils within the source area.
Accumulation at fluid interfaces (Lyu et al., 2018) including any NAPL interfaces (Glubt
and Brusseau, 2021) is important in unsaturated soils but probably less so in groundwater
plumes (Brusseau et al., 2019), although possible in the capillary zone. Adsorption and
solvation interactions with co-contaminants and NAPLs are thought to be complex and
not yet well understood (Chen et al., 2009; Guelfo et al., 2013). However, they are
apparently a function of the media OC and surface charges, PFAS chain length, and
PFAS concentration. Competitive inhibition of sorption amongst PFAS is a function of
the differences in physiochemical properties of individual PFAS, which are generally not
available because existing laboratory methods can only identify a small fraction of the
known PFAS.
Groundwater transport of individual PFAS is predominantly retarded by electrostatic
interactions and adsorption to organic carbon, likely with extended concentration decay
due to matrix diffusion in compositionally heterogenous materials (Adamson et al.,
2022). The degree of retardation is a function of the ambient geochemistry and the
physicochemical properties of the PFAS and the porous medium (Brusseau, 2018).
electrostatic interactions
PFAS with ionizable functional groups can undergo electrostatic interactions with other
materials such as charged mineral surfaces and ions in solution (ITRC, 2022).
Electrostatic interactions are greatly affected by geochemical conditions because they
predominantly determine whether polar materials are in a neutral or charged state.
The pH affects the ionization state of polar materials and interacts with other
geochemical parameters (Nguyen et al., 2020). For example, organic matter has the effect
of increasing CEC as pH increases. Also, Tang et al., (2010) found that PFOS adsorption
to the mineral goethite increased with decreasing pH, which they attributed to increased
electrostatic interactions between anionic PFOS and increased net positive charges on
goethite.
Retardation of ionic PFAS by electrostatic interactions is a function of the availability of
suitably charged surfaces. Both positively charged and negatively charged surfaces are
always present in the subsurface, but negative charges predominate. The density of
negatively charged sites on subsurface particles is measured using CEC and the density
of positively charged sites on subsurface particles is measured using AEC. Negative
charge sites are associated with materials such as organic matter, humic substances, and
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most clays (e.g., smectite/montmorillonite). Positive charge sites are associated with
materials such as kaolinite clay, and hydrous oxides of iron and aluminum (Al) such as
goethite. However, some materials such as kaolinite have "variable" charge states
affected by pH; kaolinite takes on a negative charge at high pH and a positive charge at
low pH.
Dissolved ions may alter retardation of charged PFAS via direct competition for
complementary charged sites on mineral surfaces. However, there is also evidence that
dissolved ions may interfere with hydrophobic interactions, but the interference is
probably dependent on the valence state or species of ion (Adamson et al., 2022).
hydrophobic interactions (adsorption to organic carbon)
PFAS hydrophobic interactions are strongly influenced by the conformation of their
carbon-chain backbone. Hydrophobicity increases with backbone lengthening and
decreases with backbone branching (Park et al., 2019). Branching can also be associated
with increased hydrophilicity because there are more heads to which polar functional
groups can be attached. Consistent with these relationships, PFAS adsorption to OC
decreases with decreasing backbone chain-length (Sepulvado et al., 2011) and increasing
backbone branching (Schulz et al., 2020). Also, individual PFAS may have significant
differences in their propensity for interacting with individual fractions of OC based on
the length of their backbone. Pereira et al., (2018) found that longer-chain PFAS may
preferentially interact with humin and shorter-chain PFAS may preferentially interact
with humic and fulvic acids.
combined effects
Although hydrophobic interactions are generally insensitive to solution chemistry, PFAS
adsorption to organic matter is known to be affected by factors that influence electrostatic
interactions including solution pH, the ionization state of PFAS functional groups, and
the ionic strength of the solution (Sima and Jaffe, 2021). There is a general trend for
adsorption of anionic organic contaminants to increase with decreasing pH and increasing
cation concentration (Jafvert, 1990). Anionic PFAS exhibit a similar but modified
behavior, whereby sorption to organic carbon increases with decreasing pH but
increasing concentration of only divalent cations (Higgins and Luthy, 2006; Johnson et
al., 2007; You et al., 2010; Wang and Shih, 2011; Kwon et al., 2012; Kwadijk et al.,
2013; Zhou et al., 2013; Du et al., 2014; Pereira et al., 2018). It has been suggested that
divalent cations function as bridges between negatively charged surfaces in soil and
negatively charged heads of anionic PFAS (Higgins and Luthy, 2006; Du et al., 2014).
Interestingly, Pereira et al., (2018) found that the presence of divalent cations was
important to increase adsorption to organic carbon of intermediate-chain but not long-
chain anionic PFAS.
PFAS adsorption to OC via hydrophobic interactions is widely thought to have a stronger
effect on retardation than electrostatic interactions (Higgins and Luthy, 2006; Fabregat-
Palau et al., 2021). For example, OC carries a net negative charge within the normal
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range of subsurface pH because its carboxylic and phenolic acid groups are dissociated
(Kinniburgh et al., 1999). Also, anionic PFAS such as PFOA and PFOS exist in the
subsurface in their (negatively) charged state because the pH of groundwater is almost
always greater than the acid dissociation constant (pKa) of their functional groups.
Despite the electrostatic repulsion, however, anionic PFAS still adsorb to the organic
carbon (Higgins and Luthy, 2006; Zhang et al., 2019). Further, the widely observed
greater retardation of PFOS than PFOA is due to stronger adsorption to OC conferred by
the additional C-F unit in the PFOS backbone, rather than differences in electrostatic
interactions between the soil matrix and the respective functional groups (Zhou et al.,
2010; Milinovic et al., 2015).
Modeling
Site characterization and risk assessment rely in part on estimates of the subsurface
transport of individual contaminants. However, there is not yet a reliable method to
provide these estimates for PFAS because the standard equations for retardation of
hydrophobic organic chemicals do not perform satisfactorily (Brusseau, 2018; Anderson
et al., 2019; Silva et al., 2020). A primary concern is that these equations estimate
retardation solely as a function of adsorption to OC (US EPA, 1990; 1996; 2020a)
whereas PFAS are also significantly retarded by electrostatic interactions, and (primarily
in the vadose zone) at fluid interfaces (Sharifan et al., 2021; Wang et al., 2021). US EPA,
(1996) discusses modifying the standard calculation of the retardation coefficient (Kd) by
adding a term for adsorption to inorganic materials, but problems remain. Many of the
assumptions underlying the standard retardation calculations are violated by PFAS. Any
mechanistic model must be provided coefficients specific to each process, but these are
not yet available, and available laboratory analytical methods can only identify a small
fraction of individual PFAS.
Contaminant retardation equations based on matrix organic carbon content have been
successfully used for decades to model subsurface transport of many organic chemicals,
but they are based on several caveats and assumptions that are problematic for PFAS. US
EPA, (1996) enumerated the assumptions, including: within the range of OC present
there was not significant sorption of the organic contaminant to other materials, such as
minerals or clays; the organic chemical was nonionizing and therefore did not require use
of separate coefficients for the pH-dependent fractionation between its ionized and
neutral forms; there was no contaminant source loss due to volatilization or degradation;
adsorption was linear with concentration; the system was in equilibrium with respect to
adsorption, such that adsorption and desorption kinetics could be ignored; and sorption
was reversible (rather than accounting for the facts that desorption is usually slower than
adsorption and adsorption is sometimes irreversible). These assumptions are problematic
for PFAS because the transformation of PFAS precursors is not fully characterized; some
PFAS may volatilize; PFAS are often ionizing and interact with minerals such as clays
and metal-oxides; their affinity for media interfaces is nonlinear, their adsorption may be
faster than desorption and their sorption may exhibit hysteresis; they have slow sorption
kinetics; and they exhibit some irreversible adsorption (Pignatello and Xing, 1995;
Milinovik et al., 2015; Zhi and Liu, 2018; Xiao et al., 2019; ITRC, 2022). Further, there
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has been increasing recognition over several decades that estimates of organic
contaminant concentrations in groundwater frequently do not decrease as rapidly as
predicted by sorption equilibrium models (US EPA, 1990). Desorption is recognized as
frequently being a slower process that adsorption, and slow adsorption kinetics are now
widely recognized as a significant process, attributable to effects such as pore diffusion
and matrix diffusion5 (Sudicky et al., 1985; Pignatello and Xing, 1995; Li et al., 2018;
You et al., 2020). These processes result in slower decreases in contaminant
concentrations than models based on reversible, equilibrium adsorption to organic carbon
predict.
Although the challenges of modeling subsurface transport of PFAS are considerable,
ongoing research is progressing toward developing suitable models. For example,
Higgins and Luthy, (2007) had encouraging results with a mechanistic model that
estimated the contributions of hydrophobic and electrostatic interactions using Gibbs free
energy terms. More recently, compartment models have been developed for modeling
PFAS retardation in different environments where the relative influences of retardation
mechanisms differ, such as between the vadose and phreatic zones (Brusseau et al., 2019;
Silva et al., 2020). However, adequate information to populate the parameters of these
models for site-specific application is not yet available.
Chemometric research is being conducted to provide the process-specific retardation
parameters needed for mechanistic PFAS transport models. Recent studies have used
methods that assume independent, additive linear relationships to identify the individual
contributions of various processes to observed retardation. Li et al., (2018) compiled a
dataset from the available literature and based on linear regression identified OC, pH, and
clay fraction as significant predictors of retardation. They also noted field based Kd
values were biased high relative to those calculated using laboratory batch method.
Knight et al., (2019) used linear regression on soil properties and partial least squares
regression on infrared spectra of soils to predict sorption of PFOA with some success
using OC, silt + clay content, and pH. Rovero et al., (2021) found that soil sodium and
calcium ions were significant predictors of retardation, but OC was only a significant
predictor at a relatively high (>5%) fraction. They noted a tremendous range in Kd values
in the literature and that Kd from field samples were biased high relative to those from
synthetic laboratory samples. They also cautioned that use of a single Kd value might not
be appropriate for estimating retardation of PFAS and that the standard equations for
estimating subsurface transport of organic chemicals (US EPA, 1996; 2020a) might not
be appropriate for PFAS. As suggested by Rovero et al., (2021), Fabregat-Palau, (2021)
compiled a dataset from the literature but augmented it with their own results to expand
the ranges of the predictive variables, particularly that of OC. They found that PFAS
sorption onto both organic and mineral fractions increased with chain length and that
sorption could be predicted using a parametric method suggested by US EPA (1996) as a
function of PFAS chain length, soil organic content, and silt + clay content.
5 Matrix diffusion typically results from contaminants having differing adsorption propensities for
constituents of compositionally heterogeneous media.
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SITE CHARACTERIZATION
This document focuses on management PFAS contamination in groundwater. However,
planning characterization and remediation of contaminated groundwater require a
wholistic understanding of the context in which it exists.
Conceptual Site Model (CSM)
Characterization of a PFAS site starts with the development of a preliminary CSM, and
sampling data are used to refine the CSM. While developing the CSM and planning site
characterization the input of US EPA Regional risk assessors should be sought in
addition to that from hydrologists, hydrogeologists, ecologists, and engineers, so the
information gathered is adequate to support risk assessment.
A CSM usually includes information on hydrogeology, contaminant sources and
concentrations, fate and transport, and geochemistry. CSMs for PFAS should incorporate
all environmental phases encountered at a site, including liquid (groundwater, soil water,
surface water, and atmospheric water vapor), solid (saturated and unsaturated bedrock,
unconsolidated materials, soil, and sediments), and gas (soil gas and the atmosphere).
Potential contaminant transport pathways and receptors (such as humans, wildlife, and
vegetation) are also important components of the CSM. A CSM for a PFAS site may also
require characterization of plants because PFAS are known to bioaccumulate in them.
Release Sites
All potential sources of PFAS releases to the environment need to be considered when
developing a CSM. Several PFAS source areas are often present on the same site. For
example, AFFF source areas other than fire training areas on U.S. Air Force sites have
been documented by Anderson et al., (2019). They found the most common PFAS in
groundwater at these sites were PFOS and PFHxS followed by PFOA. Frequencies of
detection for most PFAS were similar at high-volume "testing and maintenance" and
medium-volume "hangars and buildings" locations but lower at the low-volume
"emergency response" locations where only a one-time release of AFFF occurred.
Site Classifications
The four commonly recognized major PFAS sites are fire training sites, industrial sites,
landfills, and WWTPs, but metal treatment operations, runoff to surface water
contaminated by atmospheric deposition, recharge to groundwater by surface water, or
off-site upgradient locations are examples of sources that are sometimes overlooked (Hu
et al., 2016; Meegoda, 2020; ITRC, 2022). Examples of concerns typical for
characterizing each of these types are provided by ITRC (2022). However, fire-fighting
practice areas are briefly described herein to provide an example of the information
needed to develop a CSM for PFAS sites, because of their relative frequency and
complexity of contaminant mixtures (Schultz et al., 2004; Place and Field., 2012; Houtz
et al., 2013; D'Agostino and Mabury, 2014; Nickerson et al., 2021). Although the aspects
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of development of a CSM for a fire training site can be extended to other types of PFAS
sites, significant differences from other "major" PFAS site types exist. Releases at
industrial and manufacturing sites are associated with atmospheric deposition, stormwater
and wastewater discharges, disposal of solid wastes, leaks, spills, and potentially fire
training. Leachate from landfills directly enters surface water or groundwater. Releases
associated with WWTP include effluent discharges, inadvertent releases from liquid
containment structures, air emissions, and biosolids disposal or use as soil amendment.
Fire Pits
Fire pit PFAS source areas and their characteristic contaminants are described by
Meegoda et al., (2020) and ITRC, (2022) (Figure 2). Fire training pits are often located
at chemical manufacturing facilities, refineries, airports, and military installations. During
fire training, an aircraft carcass has often placed in a bermed area, dowsed with a
flammable liquid, ignited, and then extinguished using AFFF. The range in flammable
liquids applied to the fire training pits has been broad and includes fuels such as aviation
gasoline, jet fuel 4 (JP4), diesel, etc., all of which could contain BTEX compounds. In
addition to hydrocarbon fuels and the relatively broad range of materials included in
AFFF, other organic compounds measured in groundwater beneath and downgradient of
fire pits include the chlorinated volatile organic compounds (CVOCs) perchloroethylene
(PCE) and trichloroethylene (TCE) (Meegota et al., 2020). Transport of contaminants
into the subsurface from fire training areas occurs by infiltration of precipitation and
water used during fire training.
Subsurface Characterization
Although not the focus of this document, sampling and analysis of soil and solid aquifer
media during development of the CSM provides critically important information for
assessment of groundwater plumes. In addition to providing data needed to
biogeochemically characterize the subsurface media, these samples provide contaminant
concentration data that is necessary to understand the horizontal and vertical distribution
of contaminants in both the source area and the downgradient plume. They are also useful
for assessing the potential for source area soils to continue to leach contaminants to
groundwater and for back-diffusion to prolong the presence of PFAS in groundwater after
source removal. This information contributes to the technical basis for assessing risk and
developing, designing, and deploying remedial technologies to address PFAS
contamination. However, each solid media sample only characterizes conditions at
discrete points, or if vertically composited, over selected depth intervals at discrete
points. Also, these samples cannot be practicably replicated at later times to discern the
mobility of contamination in the solid phase materials.
Unlike analyses of solid media, including vertically composited samples, analyses of
groundwater provide a vertically and aerially integrated measure of contamination in the
subsurface that can readily be repeated through time to assess plume stability.
Groundwater samples integrate the effects of time-varying groundwater flow directions
such as those frequently observed intra-annually within water-table (e.g., overburden)
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aquifer matrices containing fine-grained materials. Lack of sufficiently frequent
groundwater monitoring to detect all significant shifts in groundwater flow direction can
result in inflated estimates of hydrodynamic dispersion (and therefore diffusion) relative
to advection in groundwater flow models. Such a mis-estimation can significantly alter
projected cleanup times that are influenced by matrix diffusion.
Analytical Suite
In addition to PFAS, analyses of samples from PFAS sites should include possibly co-
occurring contaminants such as hydrocarbon fuels, organic solvents, and non-PFAS
surfactants. Analyses for both individual and total PFAS should be performed and
include precursors and intermediates. Example precursor and intermediate PFAS include
chemicals such as 6:2 FTSA, 7:3 Fluorotelomer carboxylic acid (7:3 FTCA), andN-
MeFOSAA6. Precursors and PFAS intermediates are not monitored as frequently as
PFAAs, but these chemicals may have concentrations similar to those of PFAAs (Houtz
et al., 2013; Robel et al., 2017; Martin et al., 2019; Liu and Avendano, 2013). It is also
important to include geochemical parameters in the sampling and analysis program
because subsurface geochemistry can influence PFAS distribution and treatment (Weber
et al., 2017). Due to the evolving understanding of PFAS transport (Rovero et al., 2021)
the authors recommend that at a minimum, supplementary soil and sediment sample
analyses should include pH, foe, clay content, and CEC, and field screening during
surface and groundwater sampling should include pH, ORP, SC, turbidity, and DO.
Data Reduction
Forensics for identifying and unconfounding PFAS release sources are rapidly
developing but not yet comparable in capability to those in use for contaminants such as
hydrocarbons and CVOCs (Charbonnet et al., 2021). Current methods include
comparisons of analytes to those expected for manufacturing process, comparisons of
analyte distributions between samples using bar charts and radar plots, diagnostic ratios,
ordinal methods principal component analysis (PCA), and machine learning methods
such as clustering, neural networks, and Bayesian inference (Kibbey et al., 2020; Ruyle et
al., 2021; ITRC, 2022). Commercial packages are already available to support forensic
analysis, but ongoing research is exploring the utility of modern data reduction
techniques such as machine learning methods.
Regardless of the environmental media analyzed, presentation of the PFAS contaminant
concentration data should also include mapping and contouring in both plan-view and
cross-sectional views. For a better understanding of PFAS presence, fate, and transport, it
may be necessary to map and contour the contamination by individual target compound
(e.g., PFOA, PFOS, Benzene, TCE), by sums of PFCAs and PFSAs, by PFAS precursors,
and/or by total organic fluorine. Mass-flux calculations may be useful (e.g., Adamson et
al., 2020).
6 6:2 FTS and N-MeFOSAA are measured by 8327/3215. N-MeFOSAA is measured by 537. All 3 are
measured by ASTM D7979.
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ANALYTICAL METHODS
A common challenging problem with PFAS release sources is analytical characterization.
Source materials and release areas contain multiple PFAS for which available analytical
methods do not provide full coverage for individual analytes. Species-specific analytical
methods for PFAS include EPA methods 537.1 (US EPA, 2020b), 533 (US EPA, 2019a),
3512 (US EPA, 2021b), 8327 (US EPA, 2021c), and 1633 (US EPA, 2021d) (Table 2),
ASTM D-7979 (ASTM, 2019), and ASTM D-7968 (ASTM, 2017). Information
regarding the origins of analytical techniques for PFAS is provided by Kissa, (2001);
Schultz et al., (2004); and Higgins et al., (2005). US EPA provides a reference webpage
for analytical methods at PFAS Analytical Methods Development and Sampling
Research 1 LIS EPA. Summary tables for a wide range of analytical methods are provided
on the web by the ITRC at: FT :tionll.2 AnalyticalMethods Jan2022.xlsx
(live.com).
Methods for Individual Analytes
Currently four analytical methods are recognized by the US EPA for the detection of
specific PFAS in water. If other methods are approved for use at a site, then the
laboratory should provide the most recent Demonstration Of Capability (DOC) and Limit
Of Quantitation (LOQ) or Lower Limit Of Quantitation (LLOQ) for US EPA review. The
DOC should include all target analytes as well as isotopic surrogates.
US EPA Method 537.1
US EPA Method 537.1 (US EPA, 2020b) is a method for the determination of 18 selected
PFAS in drinking water. PFAS sampled from several classes are included in this analysis,
including PFCAs, PFSAs, analytes with ether linkages (Adona, GenX/HPFO-DA),
analytes containing both fluorine and chlorine atoms, and PFOS intermediates. The
method was multi-lab validated and includes data demonstrating performance in reagent
water, groundwater, and surface water.
US EPA Method 533
US EPA Method 533 (US EPA, 2019a) complements EPA Method 537.1 and is a method
for the determination of 25 selected "short chain" PFAS (i.e., those with carbon chain
lengths of 4 to 12) in drinking water, 11 of which are not included in the analyte list for
US EPA Method 537.1.
US EPA SW-846 Methods 3512/8327
US EPA SW-846 Method 3512 (US EPA, 2021b) is a rapid sample preparation method.
US EPA SW-846 Method 8327 (US EPA, 2021c) is a direct injection method that detects
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24 individual PFAS in various environmental waters (i.e., non-potable water). The PFAS
analytes are drawn from several classes, including sulfonic acids (e.g., PFOS), carboxylic
acids (e.g., PFOA), and precursors and intermediates such as fluorotelomer sulfonic acids
(e.g., 6:2 FTSA), and N-MeFOSAA. The method was multi-lab validated and includes
data demonstrating performance in four matrices of reagent water, groundwater, surface
water, and wastewater effluent.
US EPA Method 1633
US EPA draft Method 1633 (US EPA, 2021d) was developed in conjunction with the
U.S. Department of Defense (US DoD) and is a liquid chromatography and mass
spectrometry (LC-MS/MS) method in the multiple reaction monitoring (MRM) mode in
wastewater, surface water, groundwater, soil, biosolids, sediment, landfill leachate, and
fish tissue. This method quantifies 40 individual PFAS compounds sampled from
families including perfluoroalkyl carboxylic acids, perfluoroalkyl sulfonic acids, 6:2
FTSA, perfluorooctane sulfonamides, perfluorooctane sulfonamidoacetic acids,
perfluorooctane sulfonamide ethanols, per- and polyfluoroether carboxylic acids, ether
sulfonic acids, and 7:3 FTCAs.
ASTM D-7968 and D-7979
Two additional methods for analyzing PFAS in environmental media have been accepted
by the American Society for Testing and Materials (ASTM) but not the US EPA. These
are ASTM D7968 for soil and D7979 for water, sludge, influent, effluent, and wastewater
(ASTM, 2017; ASTM, 2019). These methods are single lab validated. ASTM D-7979
includes data demonstrating performance in reagent water, surface water, and WWTP
influent and effluent. ASTM D-7968 includes data demonstrating performance in four
ASTM soils. Both ASTM D7968 and D7979 quantitate the same 21 individual PFAS
compounds.
Other Analytical Methods in Development
Various methods are in different stages of development that are intended to provide
additional data regarding the presence of PFAS compounds in environmental media. The
US EPA maintains a webpage that tracks the status of method research:
https://www.epa.eov/chemical-research/status-epa-research-and-development-pfas
Methods in development include the total oxidizable precursors assay (TOP), extractable
organic fluorine (EOF) and adsorbable organic fluorine (AOF) assays, the total
organofluorine assay (TOFA), the particle-induced gamma emission (PIGE) spectroscopy
method, Nuclear Magnetic Resonance (NMR), and the fluorine specific electrode (FSE).
These methods provide general quantification of PFAS including unknown PFAS that are
most likely derived from precursor compounds in aqueous or solid phase samples. They
may be used to evaluate the potential presence of polyfluorinated compounds as
precursors to perfluorinated compounds (NGWA, 2017), or as a measure of the total
fluorine content of a sample (Suthersan et al., 2016).
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Total Oxidizable Precursors Assay
PFAAs commonly present in PFAS formulations, including PFOS and PFOA, can be
measured using US EPA methods 537.1, 533, 8327, and 1633, but PFAS precursors used
to formulate PFAS products and PFAS intermediates formed during oxidation of
precursors may also be present in environmental matrices. Some PFAS precursors and
intermediates are measured with these methods, but others are not. The TOP assay,
sometimes referred to as TOP A, was designed to estimate precursor concentrations
present and the potential for the PFAS mixture to eventually degrade into PFAAs. When
using the TOP assay, at least two samples are collected. One is analyzed for PFAS
without an oxidation treatment. The second sample is subjected to oxidation with
hydroxide, persulfate, and heat. After oxidation, the second sample is analyzed for PFAS
(Houtz and Sedlak, 2012; Houtz et al., 2013, and 2016; Martin et al., 2019). The
difference between the unoxidized and oxidized sample measurements represents the
precursor and intermediate PFAS present in a sample. During oxidation, the backbone of
some PFAS react to form shorter chain length compounds. For example, 6:2 FTSA may
form perfluoropropanoic acid (PFPrA), PFBA, and perfluorooctane sulfonamide (PFPA)
(Martin et al., 2019). Consequently, TOP assay data must be interpreted carefully.
Precursor and intermediate amounts can be estimated but the chain length or specific
species cannot be identified. When analyzing post-oxidation TOP samples PFPrA should
be included to effectively estimate precursor and intermediate concentrations (Martin et
al., 2019). Unfortunately, this compound is not included in many PFAS analytical
methods.
Combustion Ion Chromatography Methods
In remediation studies the unique properties of PFAS can make it difficult to distinguish
sorption and other losses from transformation and mineralization. Total organic fluorine
assessment (TOFA) methods attempt to address these concerns. TOFA is not specific to
precursors chain length or of the end point compounds; it is an estimate of the total
organic fluorine content in a sample. TOFA can be used where there is uncertainty as to
whether a US EPA method adequately measures all the PFAS likely to be present. TOFA
considers the total mass of fluorine which may be present as PF AS chemicals. Fluorine
mole balances are often used to describe the performance of remedial systems, and
Combustion Ion Chromatography (CIC) methods are often used to measure fluoride
(Miyake et al., 2007).
TOFA methods include EOF and AOF assays. EOF is a capture and combust technique.
It uses CIC to measure organic fluorine and fluoride. The EOF assay is synonymous with
TOF-CIC assay and has been applied to seawater, blood, freshwater, sediments, soils,
protein pellets, fish tissue, and liver tissue (McDonough et al., 2019). The total organic
fluorine can be calculated by subtracting the total inorganic fluoride from the total
fluoride (D'Agostino and Mabury, 2017).
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An adsorbable organic fluorine (AOF) assay has been described that uses elution through
polystyrene-divinylbenzene based organic carbon rather than combustion. Residual
fluoride is then removed using sodium nitrate and the sample is analyzed using CIC
(Wagner et al., 2013). This method has only been applied to water (McDonough et al.,
2019). US EPA draft method 1621 (USEPA, 2022b) is an AOF based on CIC that is
intended for detection of organofluorines in water at the parts-per-billion level. In
addition to PFAS, this method can detect fluorinated pesticides and pharmaceuticals.
Organic fluorine is removed via a nitrate wash. Reliance on contaminant adsorption to
GAC introduces a dependence on the size of the PF AS molecules that can be detected,
however, given the propensity of GACs to preferentially adsorb to longer-chain PFAS.
PIGE spectroscopy is a new, rapid, and effective AOF method to quantify total fluorine
in aqueous samples and consumer products such as papers and textiles (Ritter et al., 2017;
NGWA, 2017). This method relies on adsorption of PFAS to GAC. Due to the high
concentrations of fluoride in natural waters, there is some skepticism that this approach
will be useful in measuring total organic fluorine attributed to PFAS in water samples,
unless methods are developed to differentiate between inorganic and organic fluorine in a
manner compatible with PIGE (Hoque et al., 2002).
Nuclear Magnetic Resonance
Nuclear Magnetic Resonance (NMR) is being developed to detect a greater proportion of
total PFAS than detected by LC/MS methods (e.g., Gauthier and Mabury, 2022). NMR is
being explored as both a screening and PFAS class-specific analysis. The method can
differentiate between PFASs, non-PFAS, and fluoride ions, eliminating the need for
sample clean-up even for complex samples (e.g., Camdzic et al., 2021; McDonough et
al., 2019).
Fluorine Specific Electrode
The fluorine specific electrode (FSE) can be used as an independent measure of the
degradation of per- and polyfluoroalkyl substances (PFAS) at contaminated sites. This is
because it can be used to measure the concentration of the fluoride anion (F") in
groundwater and increased aqueous fluoride concentrations are a strong indicator of
PFAS degradation (Vecitis et al., 2009). Measurements of fluoride concentration have
been used as an indicator of PFAS treatment in treatability studies (Park et al., 2016;
Santos et al., 2016; Yin et al., 2016) but should be used with a full understanding of the
limitations of the method.
Retention of fluoride by solid phase media (i.e., soil, aquifer materials) indicates that
measurements of fluoride in groundwater samples may underestimate the total amount of
fluoride present (Stonebridge et al., 2020). Further, interferences in solution, including
oxidants and reagents to neutralize oxidants, should be taken into consideration. The need
for low detection levels is also a major constraint.
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Experimental FSE methods are under development (Stonebridge et al., 2020) but require
ion-specific electrodes. Ion-specific electrodes are needed because fluorine precipitates
with calcium so fluorine could be misrepresented in hard water using some electrodes.
Nonetheless, in general ion specific electrodes are portable and inexpensive and may be
used for aqueous samples in the field or the lab.
SAMPLING
General guidance and standard operating procedures (SOPs) for sampling PFAS in
groundwater have been provided by the US EPA and others (US EPA, 2019b, and 2019c;
NGWA, 2017; Zintek et al., 2017; MDEQ, 2018, and 2020; Proffitt, 2020). Field et al.,
(2021) summarized commonalities and discrepancies between existing U.S. PFAS
sampling guidelines.
Equipment
Modifications of procedures for collecting samples of subsurface contaminated media are
necessary when sampling them for PFAS. As with other contaminants, analytical
interferences can result from cross-contamination by inadequately decontaminated
drilling tools or well purging and sampling equipment. However, special care is required
because PFAS may be incorporated into the composition and coatings of sampling
equipment and supplies. The summary of sampling guidelines below is intended to help
prevent PFAS contamination of samples in the field but is intended to be used in
conjunction with an approved program- and site-specific sampling SOP.
• All equipment and supplies used for sampling should be scrutinized for potential
PFAS contamination. Equipment should ideally be tested prior to use for PFAS and
fluorine leaching. A soak test in water has been advocated for this determination
(Field et al., 2021) but equipment blanks must be performed regardless of whether
soak test data are used.
• Fluorocarbon or fluorotelomere materials such as Teflon™ PTFE, fluorinated
ethylene-propylene (FEP), ethylene tetrafluoroethylene (ETFE), and polyvinylidene
fluoride (PVDF) must be avoided when sampling because they contain PFAS and
could inadvertently contaminate samples.
• Aluminum foil, fast food wrappers and containers may contain PFAS and should be
avoided on-site, but if present should not be brought into the sampling areas and
hands should be washed after contact. Applied products such as sunscreens and insect
repellent may also contain PFAS. If uncertain about any applied products used,
include them in equipment blank sampling.
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• PPE, including chemically resistant Tyvek® or any other materials made of Teflon™,
Viton®, FEP, etc., should be avoided (uncoated Tyvek is acceptable). Also,
waterproof, water-resistant, or stain-resistant clothing may contaminate the sample
with PFAS used in these products; verify whether these products contain PFAS (for
example, waterproofing using natural rubber, or some waxes may be acceptable). It is
recommended that new clothes are washed 6-10 times to remove PFAS products used
in the textile industry.
• PFAS-free pens such as grease pencil or (if allowed by a mandatory SOPs) ball-point
pen or Sharpee® should be used when recording information in the field and labeling
samples.
• Decontamination should be done with Alconox®, Citrinox®, or Liquinox®.
• Excluding sample bottleware, sampling equipment constructed with stainless steel,
polyvinyl chloride (PVC), polypropylene, and high-density polyethylene (HDPE) or
possibly low-density polyethylene (LDPE) is recommended (NGWA, 2017).
• US EPA analytical method 537.1, 533, 8327, and 1633 descriptions explicitly specify
permissible sampling containers. Glass sampling equipment or sample containers
should not be used due to PFAS sorption to the glass and because of the presence of
Teflon™ liners in glass volatile organic compound (VOC) sample vial caps.
• Bailers should not be used for well purging or sampling because they unacceptably
turbidate the water. Rather, low-flow sampling methods should be used.
• Peristaltic pumps are acceptable for performing low flow well purging and sampling.
Peristaltic pumps can collect groundwater samples at water table depths of about 23
feet or less. The flexible tubing used in conjunction with the peristaltic pump should
be silicone; it should not be made of Teflon™, Viton®, FEP, or chemically related
materials as this could impart PFAS or fluorine into the groundwater sample.
• Downhole positive displacement pumps (e.g., Grundfos, Monsoon®) are also
acceptable for performing low flow well purging and sampling. However, these
pumps may contain internal parts and components made of Teflon™, Viton®, FEP,
etc. that could contaminate groundwater samples. PFAS-free water should be pumped
through these pumps and tubing and analyzed as an equipment blank. It is preferable
however to use equipment that does not contain materials that potentially could leach
PFAS.
• If feasible, single use tubing, etc., should be used when well purging or water
sampling for PFAS. Otherwise, decontamination protocols are needed between
sampling events.
• Probes or field meters should be used to track chemical stabilization of the water
purged from the wells. Preferably, the pump outflow should pass through a cell
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equipped with a multiparameter probe such as a YSI Pro Series. However, some field
instruments such as a nephelometer (turbidity meter) will require sampling from the
cell outlet.
• Field screening should be performed at a minimum for alkalinity, turbidity, pH, ORP,
SC, and DO7. Field screening should be intermittently conducted during well purging
until the site-specific sampling plan criteria for parameter stabilization are satisfied,
after which sampling may begin.
• If possible, avoid field filtration of water samples if turbidity is below 50
nephelometric turbidity units (NTU). Instead, attempt to obtain stable field screening
turbidity no greater than 10 NTU. If water sample filtering is deemed necessary, use
polypropylene or glass fiber filters. Avoid nylon and PTFE filters.
• If collecting samples for different contaminant classes (PFAS, CVOCs, BTEX, etc.),
collect samples for PFAS first to avoid cross-contamination by other sample
containers or supplies. Place the PFAS samples in an individual sealed plastic bag and
physically separate them from other sample types. If site-specific procedures permit
the use of LDPE, Ziploc® bags can be used.
• PFAS samples should be transported in coolers and stored at less than 6°C.
Coleman® or Igloo® coolers (or other PFAS-free brands) can be used to transport the
samples. If site-specific procedures permit the use of LDPE, Ziploc® bags can be
used to organize samples for shipment.
Quality Assurance/Quality Control
For US EPA sampling events, specific quality assurance/quality control (QA/QC)
activities that apply to the implementation of these procedures will be listed in the US
EPA Uniform Federal Policy for Quality Assurance Plan (UFP-QAPP) or sampling and
analysis plan (SAP) as prepared for the applicable sampling event (US EPA, 2019c). A
detailed description of QA/QC samples and the steps involved in their collection is also
available from NGWA (2017). Laboratory reporting limits should be checked against
needed detection limits specified by these documents. Individual US EPA Regions will
require QC samples to be collected to assure the quality of the data and might include or
be equivalent to, but are not necessarily limited to, the following:
• temperature blank - a temperature blank is provided by the laboratory and
accompanies the samples throughout the sampling program and back to the
laboratory. One temperature blank should be included in each sample cooler.
7 Note that if sediment or soil samples are collected for PFAS analysis they should also be analyzed for
geochemical parameters including at a minimum pH, foe, CEC, and clay content.
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• trip blank - a trip blank is provided by the laboratory and accompanies the samples
throughout the sampling program and back to the laboratory. One trip blank should be
included in each sample-filled cooler.
• field reagent blank - this sample should be collected in the field to evaluate the
potential for contamination from the overall sample collection process. Two
containers are supplied, one filled with water from the laboratory. The field staff
transfer the water from the filled container into the empty container. One field blank
should be collected on each day of groundwater sampling.
• field duplicate - These are duplicates of samples collected in the field that are blinded
to the laboratory by using some alternative name.
• matrix spike / matrix spike duplicate - these are collected in the field, and the
laboratory adds known amounts of contaminants to them before analysis.
• equipment blank - an equipment blank is collected in the field to determine if
contamination to samples has come from any equipment used during sampling. Water
is poured over or run through equipment into the equipment blank bottle. One
equipment blank should be collected on each day of groundwater sampling.
US EPA Approved Analytical Methods
Method 537.1
US EPA Method 537.1 explicitly lists required sampling containers and sampling
protocol in Section 8 of the Standard. Samples should be collected in 250 mL
polypropylene bottles with polypropylene caps and pre-preserved by the laboratory with
dry Trizma®. Samples do not have to be headspace free. All compounds listed in US
EPA Method 537.1 have adequate stability for 14 days when collected, preserved,
shipped, and stored as described in the method. Water samples should be extracted as
soon as possible but must be extracted within 14 days, and the extracts must be analyzed
within 28 days.
Method 533
US EPA method 533 explicitly lists required sampling containers and sampling protocol
in Section 8 of the Standard. Samples should be collected in polypropylene bottles with
polypropylene caps and pre-preserved by the laboratory or in the field with dry
ammonium acetate. Samples do not have to be headspace free. All compounds listed in
US EPA Method 533 have adequate stability for 28 days when collected, preserved,
shipped, and stored as described in the method. Water samples should be extracted as
soon as possible but must be extracted within 28 days, and the extracts must be analyzed
within 28 days.
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Method 8327
US EPA method 8327 explicitly lists required sampling containers and sampling protocol
in Section 8 of the Standard. Samples should be collected in polypropylene bottles with
polypropylene caps. Samples do not have to be headspace free. All compounds listed in
US EPA Method 8327 have adequate stability for 14 days when collected, preserved,
shipped, and stored as described in the method. Water samples should be extracted as
soon as possible but must be extracted within 30 days, and the extracts must be analyzed
within 30 days.
Draft Method 1621
US EPA draft method 1621 explicitly lists required sampling containers and sampling
protocol in Section 8 of the Standard. Samples should be collected in polypropylene
bottles with polypropylene caps, and aqueous samples must be collected in triplicate but
do not have to be headspace free. Extractions should be performed as soon as possible but
samples may be maintained in the laboratory up to 90 days prior to analysis.
Draft Method 1633
US EPA draft method 1633 explicitly lists required sampling containers and sampling
protocol in Section 8 of the Standard. All sample containers must have linerless HDPE or
polypropylene caps. Other sample collection techniques, or sample volumes may be used,
if documented.
• Aqueous samples: Automatic sampling equipment should be used to collect free-
flowing samples as grab samples. Excepting landfill leachate, two aliquots should be
collected, one in a 500 mL HDPE bottle and one in a 250-mL or 125-mL HDPE
bottle. Use smaller bottles if adequate sample is not available. Landfill leachate also
requires two aliquots, but both may be collected in 100 mL HDPE bottles. Maintain
the samples at 0 - 6 °C from the time of collection until shipped to the laboratory.
Samples must be protected from light in the laboratory. Samples may be held in the
laboratory for up to 90 days from collection if stored at < -20 °C or 28 days if stored
at 0 - 6 °C, but issues may then occur with certain PFPA, perfluorooctane
sulfonamide ethanols and perfluorooctane sulfonamidoacetic acids after 7 days.
• Soil, sediment, and biosolid samples: These should be collected as grab samples in
wide-mouth HDPE jars filled to no more than % capacity. The samples should be
maintained at 0 - 6 °C from the time of collection until receipt at the laboratory. Solid
samples (soils and sediments) and tissue samples may be held for up to 90 days, if
stored by the laboratory in the dark at either 0 - 6 °C or < -20 °C, with the caveat that
samples may need to be extracted as soon as possible if nonafluoro-3,6-
dioxaheptanoic acid (NFDHA) is an important analyte.
• Sample extracts: Sample extracts should be stored in the dark at less than 0 - 4 °C
until analyzed. If stored in the dark at less than 0 - 4 °C, sample extracts may be
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stored for up to 90 days, with the caveat that issues were observed for some ether
sulfonates after 28 days. These issues may elevate the observed concentrations of the
ether sulfonates in the extract over time. Samples may need to be extracted as soon as
possible if NFDHA is an analyte of interest.
REMEDIATION
This section addresses PFAS remediation methods applicable to groundwater, which
remain a topic of active research. For the interested reader, however, a review of PFAS
treatments specific to drinking water is provided by Crone et al., (2019), a review of soil
remedial technologies for PFAS is provided by Mahinroosta and Senevirathna, (2020),
and review of technologies to potentially address back-diffusion is provided by Brooks et
al., (2021).
The outlook for successful management of PFAS in groundwater has been assessed by
Newell et al., (2020). They conducted a study comparing PFAS to other groundwater
contaminants for the purpose of assessing remediation potential. They compared PFAS to
CVOCs, benzene, 1,4-dioxane, and MTBE. They evaluated nine metrics for comparison:
production quantities, number of potential sites, detection frequency, required
destruction/removal efficiency, median plume length, hydrophobic sorption, regulatory
criteria, in-situ remediation capability, and research intensity. They also conducted five
qualitative comparisons: low-level detection capabilities, methods to assess risk of
complex mixtures, nonaqueous phase dissolution, plume length prediction, and monitored
natural attenuation (MNA) protocols. They found that production quantities, number of
potential sites, detection frequency, and required destruction/removal efficiency indicated
that PFAS might be a lesser challenge; that median plume length was comparable to
chlorinated solvent plume lengths; that adsorption was not definitive; and that regulatory
criteria, in-situ remediation, and research intensity indicated that PFAS might be more
troublesome than the comparison contaminants. Their assessment of the qualitative
metrics was that while remediating PFAS sites will be challenging the groundwater
community has the experience to accomplish the task.
Comparative Reviews
Comparative reviews of established and developing remediation processes for PFAS in
water are provided by Merino et al., (2016), Kucharzyk et al., (2017), Ross et al., (2018),
Trojanowicz et al., (2018), Nzeribe et al., (2019), Meegoda et al., (2020), and
Wanninayake (2021). These reviews discussed remedial processes including adsorption
using activated carbon (AC) including colloidal (CAC), powdered (PAC), and granular
(GAC) forms; IXR) and non-ionic exchange resins (XR); biopolymers; molecularly
imprinted polymers (MIP); RO; microfiltration, ultrafiltration (UF), and nano-filtration
(NF); ozonation; Fenton processes; microwave hydrothermal treatment; incineration;
heat-activated persulfate; permanganate oxidation; advanced oxidation processes (AOP)
including electrochemical oxidation (EO) and plasma; advanced reduction processes
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(ARP) including use of aqueous iodide or dithionite and sulfite, vitamin-B12 and Ti(III)-
citrate reductive defluorination; zerovalent iron (ZVI); photolysis; photocatalysis;
sonochemical degradation, sub- or supercritical water; bioremediation using microbes
and fungi; ozonation under alkaline conditions; radiolytic processes using ionizing
radiation; and treatment-train approaches.
Dominant Remedial Methods
Presently, GAC, IXR or less commonly XR, RO, and NF are the commonly available and
demonstrably implementable primary technologies for removal of PFAS from water
(Tang et al., 2006; Espana et al., 2015; Meegoda et al., 2020). Ex-situ GAC with
incineration of the spent GAC is the most common treatment technology. It performs
well for removal of long-chain PFAS but does not perform well on short-chain PFAS or
PFAS precursors. IXR can remove a wider range of PFAS, but older versions of IXR do
not perform well for the shortest chain PFAS and had not yet been evaluated for
precursors. RO and NF have been more successful than GAC and IXR for complete or
nearly complete removal but are considerably more expensive. New GAC and IXR have
recently been developed, however, that are now commercially available and purportedly
are better able to sequester wider ranges of PFAS.
Although many applications of PFAS treatment technologies to date have been ex-situ,
in-situ treatment is also being explored. A current topic of interest is in-situ colloidal
activated carbon (CAC) emplaced either by injection on a grid or as a carbon-based
permeable reactive barrier (PRB). The performance will likely be a function of the ability
of the subsurface injections to infiltrate all migration pathways of significant permeability
so that the contaminant plume fully interacts with the emplaced AC. Sorption capacity
and competition for surface sites between PFAS and other groundwater solutes are site-
specific issues that will likely determine long-term treatment performance. Like other
sorption-based technologies, saturation capacity of the treatment material and
contaminant rebound via matrix diffusion will likely be issues.
Developing Remedial Methods
Methods of PFAS removal from water other than GAC, IXR, RO, and NF are still in
various stages of development or have drawbacks that presently make them less attractive
than the presently dominant methods. Some promising methods have progressed to pilot
testing or small-scale (often mobile) implementation. Others, however, were conducted at
the laboratory bench scale under idealized, sometimes extreme conditions (e.g.,
temperature), and using relatively high PFAS concentrations. They also variously
required uncompetitive residence times, or were less energetically favorable than
currently used technologies, or were likely to increase toxicity of non-target materials
that are common in environmental media. Some examples of such problems are provided
by Horst et al., (2020). For example, plasma, electrochemical treatment, and sonolysis are
expected to be more energy intensive than incineration. Also, in the presence of non-
target materials such as chloride, bromide, arsenic, and trivalent chromium, the hydroxyl
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radical created by some oxidative technologies can react with these to form problematic
materials such as perchlorate, bromate, arsenate, or hexavalent chromium.
Short-chain PFAS
Most research on the remediation of PFAS has focused on relatively long-chain species.
However, Ateia et al., (2019) reviewed remediation of short-chain (four to seven carbons)
and ultra-short-chain (two to three carbons) PFAS. Removal of short-chain PFAS is more
challenging than for longer chain PFAS using carbonaceous sorption. Short-chained
PFAS are hydrophilic and less likely to form aggregates, yet adsorption mechanisms for
carbonaceous materials such as AC are hydrophobic and rely partly on electrostatic
interactions. Electrochemical oxidation (EO), an AOP, has been shown to degrade short-
chain PFAS less efficiently than longer chain PFAS (Niu et al., 2012) and again, can
create undesirable by-products from non-target constituents in the water to be treated.
Previous studies of WWTPs have found higher PFAS in the effluent than the influent,
indicating transformation of precursors occurs at a greater rate than that of endmember
PFAA (Lee et al., 2010a; Pan et al., 2016), and at the time of writing there was no
evidence of reactions of short chain PFAS to the knowledge of the authors. Also,
WWTPs did not treat short-chain PFAS, but it was suggested that hybrid sorption
systems might be effective. It was concluded that at least some of these degradation
methods were promising but further development was needed, including new sorbents for
short chain PFAS and decreased costs such that upscaling costs would not be cost
prohibitive.
Handling of Remedial By-Products
Remediation of PFAS often does not result in complete defluorination or mineralization.
Consequently, management of remedial by-products containing residual or recalcitrant
PFAS is necessary. Therefore, existing literature was reviewed by Horst et al., (2020) to
identify the potential remedial by-products that will require management. Current and
developing remedial methods rely on either thermal destruction or sequential
defluorination. Current technologies typically concentrate PFAS prior to disposal or
attempts at thermal destruction. Treatment technologies such as aerobic or anaerobic
digestion in WWTP or digestion of sludges also transform PFAS precursors to more
recalcitrant species such as PFAAs, but it is likely that some precursors or intermediary
PFAS remain. In addition, PFAS treatment technologies in development can create
problematic by-products when applied to waters from the natural environment in which
non-target materials are typically present. For example, Horst et al., (2020) list
perchlorate, bromate, and hexavalent chromium formed from chloride, bromide, and
trivalent chromium.
Adsorbents
Some reviews of adsorption-based remedial processes for PFAS in water are provided by
Wang et al., (2019); Zhang et al., (2019); Gagliano et al., (2020); Vu and Wu, (2020);
and Dixit, (2021). Wang et al., (2019) summarized treatment of PFAS in groundwater
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using carbonaceous materials, numerous IXR, zeolites, minerals, and alumina (commonly
used in water treatment). Zhang et al., (2019) reviewed the technical feasibility of PFAS
adsorption by several materials in solution, including AC, IXR, minerals, MIP, carbon
nanotubes (CNTs), and a wide range of biosorbents. Gagliano et al., (2020) focused their
review on relative performance of adsorbents including AC, biochar, MIP, IXR,
nanoparticles, CNT, and mineral materials, the effect of organic matter, and adsorbent
regeneration. Vu and Wu, (2020) reviewed sorbents including carbonaceous materials,
resins, minerals, biomaterials, and polymers but included somewhat older literature in
their review. Dixit et al., (2021) reviewed PFAS removal by IXR for water treatment
plants but based their review only on the most recent applied developments.
Sorption Materials
Ex-situ adsorption by GAC is presently the dominant remedial technique for PFAS in
water (Schroder et al., 2010) and is considered as providing the best combination of
reliability and cost-effectiveness for PFAS removal (Hansen et al., 2010). As previously
stated, GAC from most sources is known to be more effective at removing long-chained
PFAS than short-chain PFAS (Appleman et al., 2014). GAC is susceptible to earlier
breakthrough of (potentially unmonitored) shorter-chain PFAS and adsorbs PFOS better
than PFOA (Yu et al., 2009), although continuing research on developing GAC with
better performance for short-chain PFAS is promising. PAC is used in potable water
treatment plants but compared to GAC has the disadvantage of limited percent removals
and being difficult to regenerate, so it is typically disposed of as a constituent of spent
treatment sludges. PAC could be used for modest removal needs.
Like GAC, resins are frequently used for PFAS removal. Both anionic and non-anionic
resins have been studied for the remediation of PFAS. An advantage of resins over
carbon-based sorption materials such as GAC is that it can have greater effectiveness at
removing shorter-chain PFAS, particularly for newer resins developed specifically to
adsorb PFAS, but the efficiency of resins varies greatly and background water quality has
an influence (Gagliano et al., 2020). Most PFAS at ambient groundwater pH values are
usually anions so strong base IXR are indicated for their treatment (Gagliano et al.,
2020)8. Deng et al., (2010) reported that polyacrylic resins have shown a higher
efficiency for PFAS removal than polystyrene resins. Also, Dixit et al., (2021) indicated
that polystyrenic resins have a higher affinity for PFAS than NOM, whereas polyacrylic
resins have a higher affinity for NOM than PFAS. Older resins were apparently relatively
specific to PFAS species, and this was considered problematic (Vu and Wu, 2020), but
newer resins (e.g., A592E) have been reported to capture a wide range of PFAS (Dixit et
al., 2021). Several resins developed for PFAS removal are now commercially available
(e.g., Amberlite™, PSR2 Plus, CalRes 2301, Sorbix PURE LC, Resin Tech SIR-110-HP)
but they are intended for single-use rather than regeneration and thus lack one of the key
advantages of most resins, regeneration. In addition, in contrast to the generally observed
trends, at least one supplier (Calgon™) indicates that PFAS-specific GACs can now be
more successful than IXR in removing short-chain PFAS.
8 Obviously, if a mixture of ionic forms of PFAS were present then treatment using a single charge-
dependent material would require augmentation, likely in a treatment-train approach.
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Adsorption alternatives have been thought to be more limited for the newer, shorter-chain
PFAS that have replaced the older PFAS because of their lower hydrophobicity. Vu and
Wu, (2020) in their review of literature, that included some comparatively older resins,
concluded that most sorption studies were conducted in synthetic solutions using AC or
an IXR with rather high specificity, and consequently were successful only in removing
PFOA and PFOS. Dixit et al., (2021) after reviewing the most recent and applied
literature indicated that newer resins were able to remove a broad range of PFAS, but to
restate, lacked the usual advantage of regeneration.
Many sorbents other than AC and IXR have been investigated but are considered
impractical or unsuitable for various reasons. As previously mentioned, certain PACs
may have higher adsorption capacity than certain GACs (Hansen et al., 2010) but is
normally not regenerated for practical reasons including difficulty in separation from
other constituents of treatment sludges (Crone et al., 2019). Biochar and multiwalled
CNTs were found to be incapable of adequate short-chain PFAS removal (Inyang and
Dickenson, 2017; Deng et al., 2010). Chen et al., (2011) found that maize-straw-origin
ash and single-walled CNTs were both effective adsorbents for PFOS but did not
investigate their capability for removal of other PFAS. Ochoa-Herrera and Sierra-
Alvarez, (2008) found that AC showed superior sorption for PFAS than a zeolite and
activated sludge, and Du et al., (2014) illustrated that activated alumina, silica, zeolite,
and montmorillonite were inferior to AC (as well as IXR) for PFAS removal. Other
materials that have been shown to adsorb PFAS have the drawback of not being available
on an industrial basis, including for example: synthesized materials (e.g., a porous
aromatic framework constructed from benzene rings, covalent trizazine-based
framework, hexagonal boron nitride nanosheets, quaternized cotton and aminated rice
husk, AC fibers from polyacrylonitrile fiber, and poly(ethylenimine)-fuctionalized
cellulose microcrystals (Gagliano et al., 2020). Nanoparticle material oxides such as
titania, iron oxides, alumina, and silica also have been shown to possess the ability to
sorb to PFAS via electrostatic interaction and hydrogen bonds (Wang and Shih, 2011; Lu
et al., 2016). A study by Zhou et al. (2016) of magnetite nanoparticles used as a magnetic
nanocomposite found that the material sorbed PFAS, and Gong et al., (2016) eliminated
PFOA using starch-stabilized magnetite nanoparticles. Microplastics, common in some
seawaters, have been shown to weakly sorb PFAS (Llorca et al., 2018).
Sorption Mechanisms
Sorption mechanisms vary by material. The mechanisms for sorption to AC and biochar
are electrostatic and hydrophobic interaction (Zhang et al., 2019) and the hydrophobic
effect has been found to increase with PFAS chain length (Gagliano et al., 2020). Deng et
al., (2010) found that sorption rate and capacity of IXR were mainly a function of
polymer matrix and porosity, but the functional group of the resin also impacts the rate of
PFAS removal (Dixit et al., 2021). Recently available IXR were designed specifically for
PFAS function via ion exchange and hydrophobic effects (Dixit et al., 2021). Other
sorbing mechanisms for various materials include hydrogen and covalent bonding
(Gagliano, et al., 2020) and at adequate concentrations, formation of micelles or hemi-
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micelles as well as ligand and ion exchange and fluorophilic interaction (Vu and Wu,
2020).
Influence of Environmental Conditions and Co-contaminants
Geochemical conditions can significantly affect PFAS adsorption. Tang et al., (2010)
found that PFOS adsorption onto silica was not greatly affected by pH, solution ionic
strength, or calcium concentration. Wang and Shih, (2011) also found that adsorption of
PFOS and PFOA on alumina decreased with increases of sodium, potassium, magnesium,
and calcium cations. The efficacy of adsorbents for PFAS may be reduced by competitive
effects. A review by Zhang et al., (2019) reported that dissolved organic carbon, NOM,
HA, and WWTP effluent organic matter were known to reduce the adsorption affinity of
sorbents for PFAS. However, a recent study by Siriwardena et al., (2019) found that co-
contaminants including kerosene, TCE, and ethanol, and changes in pH, presence of
sulfate, NOM, and iron oxides had little impact on the sorption of PFAS to GAC.
Regeneration of GAC and IXR
Both GAC and IXR can be regenerated for reuse but Gagliano et al., (2020) concluded
that economical regeneration of adsorbents is challenging and in need of additional
research. GAC has an economic advantage over many other treatment technologies
because it can often be relatively inexpensively regenerated several times for re-use using
heat (Baghirzade et al., 2021; Xiao et al., 2020). The GAC is dried at a relatively low
temperature of about 105°C, thermally pyrolyzed at higher temperatures of about 650-
850°C (or, according to some sources, in the range of 1000-1200°C for complete
mineralization of PFAS) and then usually treated with steam or carbon dioxide. The
amount of GAC active surface area recovered varies widely, however, because without
fine control of the regeneration process the micropore structures of the GAC can be
damaged. To date, although demonstrated at the bench scale and possibly completed
without sampling for PFAS, GAC reactivation of PFAS-laden GAC has not been
demonstrated at the full scale.
Regeneration of IXR is less economical than regeneration of GAC. Also, currently
commercially available resins designed for PFAS removal are intended for single use
although research and development continue. Chemical adsorbent regeneration, typically
used for IXR, is usually done using methanol, sodium chloride, or sodium hydroxide.
Stand-alone methanol or sodium chloride or their combination work very well for PFAS
removal, but stand-alone sodium hydroxide performs very poorly. For some resins, a
combination of a base (e.g., sodium hydroxide) with an inorganic (e.g., sodium chloride)
is effective. Unfortunately, all these processes result in significant volumes of
regeneration fluids for which there is no economical way to separate the contaminants,
and the ultimate destruction of the PFAS requires unusually high (perhaps as high as
1600-2000 °C) temperatures (Bolan et al., 2021).
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Reverse Osmosis (RO) and Nanofiltration (NF)
Pressure-driven membrane filtration processes such as RO or NF are well established
separation technologies, which require hydraulic pressure of the feed stream exceeding
the osmotic pressure to generate net flux of the clean water (i.e., the membrane
permeate). RO membranes have lower molecular weight cut-off (MWCO) than NF
membranes for superior rejection efficiencies of dissolved contaminants including
monovalent ions (e.g., Na+, CI"). NF membranes provide higher water permeability and
thus higher energy efficiency than RO membranes due to the larger membrane pore size
distribution. RO and NF have gained a foothold in various water treatment and
desalination applications as a means of municipal water generation from seawater and as
a potable water treatment technology. However, NF and RO have higher operating cost
compared to other dominant remedial technologies for PFAS such as GAC or IXR.
Studies have shown that RO and NF thin-film composite (TFC) membranes with
polyamide barrier layer have excellent rejection efficiencies for PFAS compounds. Tang
et al., (2006, and 2007) found between 90% and 99% removal of PFOS by NF and >90%
by RO membranes. Patterson et al., (2019) investigated point-of-use GAC and RO
systems and found that RO removed about 100% of PFAS. Appleman et al., (2013) also
reported rejection efficiency >93% by NF for 9 different PFAS compounds whose molar
weight ranged between 214 g/mol - 500 g/mol (for PFBA and PFOS, respectively).
Using water samples from water reuse plants in California, Appleman et al., (2014) also
demonstrated superior rejection efficiencies of RO relative to GAC or IXR with 24 PFAS
of various size and functional groups. In a more recent study (Chow et al., 2021) which
investigated PFAS in bottled water products, it was also confirmed that RO-treated
products contained significantly lower summed PFAS concentration than bottled water
products without RO treatment. The above results demonstrated that RO and NF
membranes are both effective for separation of long- and short-chain PFAS as the
membranes reject solutes primarily by size-exclusion (van der Bruggen et al., 2003).
Although the amount of water produced/treated may largely differ depending on the
system size, configuration, and application, NF and RO continuously generate the
residual concentrate stream. However, disposal or management of a large volume of the
membrane concentrate can be a major challenge. The presence of PFAS in the membrane
concentrate may complicate the residual management process due to this concern (Tow et
al., 2021). Accordingly, conventional concentrate management methods such as surface
discharge, sewer, deep well injection, or evaporation ponds may be restricted.
Regulations on PFAS are expected to rapidly increase, and accordingly, the effectiveness
of maximum contamination levels (MCLs) set by different regulatory agencies will need
to be evaluated. Thus, complex, stringent post-treatment or management processes may
be necessary for PFAS-laden membrane concentrate.
Various operational strategies (e.g., concentrate recycling or closed-circuit desalination)
and system configurations (e.g., multi-stage RO with interstage booster pump) have been
explored to minimize the concentrate disposal (Lee et al., 2019; Efraty, 2012). High-
recovery systems, however, not only require costly components, but also increase the
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energy consumption per unit volume of the permeate (i.e., specific energy consumption)
to accommodate a higher concentration factor in the membrane feed channel.
RO and NF systems also suffer from fouling of the membrane due to accumulation of
colloidal particles, scaling or inorganic salts, and growth of microorganisms: liu et al.,
2018a found that PFAS adhere to membrane surfaces or complex with solutes resulting in
surface adsorption via electrostatic or hydrophobic interactions. Formation of a thick
fouled layer above the membrane surface may cause severe decline of the permeate flux,
thereby requiring frequent replacement of the membrane element. Zhao et al., (2016)
demonstrated 70% productivity loss due to membrane fouling induced by PFOS in the
feed solution with magnesium and/or humic acid (HA). In a recent study by Boo et al.,
(2018), development of loose and negatively charged NF membrane was investigated to
allow selective passage of salts to prevent formation of PFAS-salt bridging. To overcome
the challenges from membrane fouling potentially induced by PFAS adsorption,
continued research efforts are necessary to understand the behavior of PFAS under
various fluid conditions and material properties of the membrane.
Biodegradation
Vertical stratification of PFAS species often observed at PFAS release sites could be
explained by aerobic biodegradation of selected PFAS species, predominantly including
PFAS precursors, in addition to differential mobility (Bekele et al., 2020; ITRC, 2022).
At PFAS release sites generally more of the longer chain PFAS and PFAS precursors are
observed in vadose soils than the underlying groundwater, where PFAAs and branched-
chain isomers are more common (Schulz et al., 2020). This could be consistent with
greater mobility of shorter chain PFAS, aerobic transformation of longer chain PFAS to
shorter chain PFAS, and aerobic transformation of PFAS precursors to PFAAs. Thus,
increased groundwater transport could be expected from perhaps common vadose-zone
PFAS degradation to PFAAs. Available studies are congruent with biodegradation as the
predominant mechanism for vadose zone PFAS degradation. Further, in groundwater,
linear PFAS are often closer to the source area and branched isomers are further
downgradient. This is an illustration of differential retardation between linear and
branched PFAS.
WWTP PFAS studies are a good source of information for engineered biological
treatment of groundwater. It is not uncommon for WWTPs, which rely on biodegradation
to degrade incoming contaminants, to have higher concentrations of PFOS and PFOA in
treated water than influent water. This suggests a higher rate of biotransformation of
PFAS precursors to PFOS and PFOA than the rate of any destruction of PFOS and
PFOA, which at present is assumed to not occur. Most WWTPs rely on aerobic digestion
to treat wastewater, although a minority use anaerobic digestion, and both processes are
microbially mediated. In summary, the degradation of PFAS precursors in WWTPs
appears to primarily be biologically mediated and occur under both aerobic and anaerobic
conditions (Lenka et al., 2021).
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PFAS degradation processes are not universally observed in WWTPs and appear to be
dependent on differences in the bacteria present in individual treatment plants. Ochoa-
Herrera et al., (2016) investigated the microbial toxicity and biodegradability of PFOS
and shorter-chain PFAS. They found that none of the tested compounds were toxic to
methanogenic (anaerobic) activity of wastewater sludge up to 500 mg/L, but all the PFAS
were highly resistant to microbial degradation under both aerobic and anaerobic
conditions. In contrast, Liu and Avendano, (2013) found that sources of activated sludge
inoculant were critical to biodegradation. The same precursors that proved recalcitrant in
the presence of one inoculum were degraded in the presence of another. The same
outcome resulted when aerobic degradation was inoculated using soils. Bacteria cultured
from contaminated groundwater was also able to degrade FTOHs (4:2, 6:2, and 8:2
FTOH) by oxidation, probably cometabolically, and 6:2 fluorotelomer sulfonate (6:2
FTS) by desulfonation and defluorination.
Present data do not show all PFAS can be mineralized using biological mechanisms
(Horst et al., 2020). Aerobic biodegradation of precursors is commonly observed in
several environmental media, and a review by Shasavari et al., (2021) noted that fungal
and bacterial strains have been isolated that are capable of degrading PFAS. Based on a
review of PFAA precursor degradation in the environment by microbes, activated sludge,
plants, and earthworms, Zhang et al., (2021) concluded that environmental
biotransformation mainly involves aerobic oxidation, dealkylation, and defluorination of
non-fluorinated functional groups, and surprisingly the cleavage of carbon-fluorine (C-F)
bonds, to form shorter-chained PFAAs. In a review of microbial degradation of PFAA
precursors in microbial culture, activated sludge, soil, and sediment, Liu and Avendano,
(2013) found the lack of direct detection methods for precursors problematic and that
there was a significant issue with bound residues in soils. Nonetheless, significant
transformation of (FTOHs), fluorotelomer sulfonate, fluorotelomer stearate and citrate
esters, fluorotelomer phosphate esters, fluorotelomer acrylate and methacrylate,
fluorotelomer ethoxylates, and n-ethyl perfluorooctane sulfonamidoethanol were found in
the literature. Noteworthy, some PFAS precursors were found to be anaerobically
defluorinated during biotransformation. Examples of anaerobic bioactive environments
include the less common anaerobic digestion WWTPs and treatment sludge piles.
Degradation of the PFAAs that result from PFAS precursor degradation has not been
observed in the environment to the knowledge of the authors, and they have until recently
been thought to be entirely recalcitrant. However, biodegradation of PFOA and PFOS
were reported in a methodological study of sewage sludge by Schroder, (2003) under
anaerobic but not aerobic conditions. Decreases in concentration of PFOA were much
slower than and subsequent of those of PFOS in the study. Fluoride, indicative of
mineralization, was not detected in either the aerobic or anaerobic reactors.
A study by Yi et al., (2016) sought to optimize degradation of PFOA by Pseudomonas
parafulva and obtained a reduction of 32% after 96 hours, and 48% after 96 hours using
supplemental glucose. A decrease of about 67% PFOS in 96 hours was observed in a
study by Kwon et al., (2014) using Pseudomonas aeruginosa. It was shown by
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Chetverikov et al., (2017) that Pseudomonasplecoglossicida was able to transform 75%
of PFOS in soil to perfluoroheptanoic acid.
A microcosm study by Huang and Jaffe, (2019) investigated the microbial destruction of
PFOA and PFOS using Acidimicrobium sp. strain A6 (A6) under iron-reducing
conditions. A6 is an autotroph that under anaerobic conditions oxidizes ammonium
(electron donor), known as the Feammox process. It is found in the environment where
the pH is less than 7 in iron rich soil. A6 can also use hydrogen as an electron donor and
has the practical benefit of being able to cometabolically degrade trichloroethene (TCE)
and perchloroethene (PCE). The Huang and Jaffe, (2019) study included investigation of
whether PFOA and PFOS can be biodegraded using either pure A6 culture or microbially
enriched A6 culture (i.e., denitrifiers Ralstonia and Bacillus, iron reducers
Acidimicrobium and Aciditerrimonas, and sulfate reducers Desulfosporosinus) using
ammonium or hydrogen under iron-reducing conditions during 60 and 100-day
incubations. They reported the defluorination of up to 60% PFOA and PFOS by
Acidimicrobium sp. strain A6 (A6) during 100-day incubations, with the highest
degradation percentages occurring for the A6 with microbial enrichment using
ammonium as the electron donor. Significantly, there were concomitant evidentiary
increases in fluoride and shorter chain perfluorinated products. Buttressing these results,
Huang et al., (2022) performed anaerobic incubation of biosolids containing PFOA and
PFAS-free lab samples spiked with PFOA. Samples were spiked with Acidimicrobium sp.
Strain A6, or ferrihydrite, or both. Control samples were also used, and incubations lasted
150 days. The only samples that exhibited decreased concentrations of PFAS were those
that were spiked with both Acidimicrobium sp. Strain A and ferrihydrite; PFOA
concentrations decreased in excess of 50% in these samples, with concomitant increases
of shorter-carbon-chain PFCAs and fluoride.
Oxidation
AOPs for PFAS degradation have been extensively tested (Moriwaki et al., 2005). AOPs
are based on free radical oxidants such as hydroxyl that can be generated numerous ways
(Trojanowicz et al., 2018). AOPs for some PFAS such as fluorotelomers and PFCAs
usually proceed by sequential defluorination, but AOP apparently does not destroy
PFSAs (Horst et al., 2020). In fact, PFAS in general are recalcitrant to chemical
oxidation. The resistance to oxidation is believed to be the result of the difficulty of the
relatively large free radicals to access the C-C bonds in the fluorocarbon tails. This is due
to the tight packing of the fluorine atoms around the C-C bonds as well as the strength of
the C-F bonds.
Wang et al., (2019) found mixed results for PFAS treatment using AOPs. Examples of
numerous AOP treatment permutations abound, and results are sometimes contradictory.
Hydrogen peroxide activated using ultraviolet (UV) light, ozone, or ferrous iron were
tested but these trials were unsuccessful in at least one study (Schroder and Meesters,
2005). Ozone with UV has also been trialed and was successful, but relatively slow (Hori
et al., 2004). Ozone and hydrogen peroxide with ozone were successful under alkaline
conditions and pretreatment under acid pH conditions (Lin et al., 2012a). Ozone, UV,
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heat, and photolysis have also been successful, often variously combined with persulfate,
Fenton's reagent, zerovalent metals, and subcritical water (Hori et al., 2005, and 2008a;
Huang et al., 2005; Lee et al., 2010b and 2012; Qu et al., 2010; Tsitonaki et al., 2010; Lin
et al., 2012a; Jin et al., 2014). Suthersan et al., (2016) noted that in-situ oxidation for
PFAS remained a largely un-surmounted challenge. They described a nascent in-situ
remedial technique for PFAS using activated persulfate that reportedly can mineralize
PFOS. This technology is purportedly dependent on a specific activation method.
AOPs have several potential drawbacks. In general, it is problematic when AOPs are
used to treat CVOC plumes that are comingled with PFAS (Merino et al., 2016). AOPs
frequently create PFAS with shorter perfluorinated alkyl chains that are not degraded
(Nzeribe et al., 2019; Trojanowicz et al., 2018), generally increasing PFAS transport.
AOPs also have the liability of forming more toxic oxidized forms of non-target materials
often present in environmental samples.
Reduction
PFAS should be less resistant to reduction than oxidation because the relatively smaller
free radicals should have easier access to the C-C bonds in the fluorocarbon tails. This is
a function of the tight packing of the fluorine atoms around the C-C bonds as well as the
strength of the C-F bonds.
Chemical reduction shows some promise for in-situ destruction of PFAS in groundwater
(Wang et al., 2019) but requires several hours for mineralization of PFAS (Horst et al.,
2020). PFAS have been degraded using ZVI in sub- and super-critical water (Hori et al.,
2006, and 2008b). Nanoscale ZVI has been shown to have improved efficacy due at least
in part to the increased specific surface area and in the presence of stabilizing agents
(Crane and Scott, 2012; Arvaniti et al., 2015) that prevent aggregation which reduces
reactive surface area (Phenrat et al., 2007).
Cui et al., (2020) reviewed destruction of PFAS using ARPs and Trojanowicz et al.,
(2018) reviewed destruction of PFAS using ARPs and AOPs. ARPs have been
understood to sequentially defluorinate the carbon backbone of PFAS; Bentel et al.,
(2019) provide detailed description of several degradation mechanisms. Cui et al., (2020)
found that degradation pathways for PFAS vary as a function of their head groups and in
at least some cases the length of their fluorocarbon chain. Degradation was highly
influenced by solution chemistry factors, such as pH, concentration of sulfate or iodide,
DO, HA, nitrate, and temperature. Degradation increased with increasing temperature,
and nitrate slowed degradation with increasing concentration. Increasing pH favors
reductive degradation, and the optimal pH is within the alkaline range evidently because
there is less hydrogen/hydronium ion (H+/H30+) to scavenge aqueous electrons (eaq~).
When using sulfite (SO32") or iodide (I") as a solute, degradation efficiency increased with
dose because more eaq" becomes available until a critical level was reached, above which
efficiency decreased because scavenging of the eaq" became dominant. The inhibiting
effects of DO on destruction of PFAS are considerable due to scavenging of eaq" by DO.
Also, water matrix constituents can affect destruction of PFAS in water by different
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mechanisms. Humic acid for example is a constituent of dissolved organic matter (DOM)
that can accelerate degradation below a threshold concentration, above which
decomposition slows. The mechanism of action is uncertain.
Other Technologies
Several other groundwater remediation technologies for PFAS are in various stages of
development including field trials, such as photocatalysis, electrochemical treatment,
foam fractionation, plasma, and sonification. Several materials as photocatalysts have
been investigated for the cost-effective treatment of PFAS in groundwater using light
radiation (Xu et al., 2020a). Sahu et al., (2018) synthesized petitjeanite Bi30(0H)(P04)2
(BOHP) microparticles as a photocatalyst and used them with UV light to mineralize
-100% PFOA within one hour of treatment. Mineralization of PFAS also occurred in the
presence of NOM. Liu et al., (2017) used nZVI and visible light, as a cheaper alternative
to UV, and removed 69.7% and 89.7% PFOA in the absence and presence of hydrogen
peroxide, respectively. Xu et al., (2020b) eliminated almost all PFOA using titanium
dioxide with peroxymonosulfate (PMS) and visible light. All these trials required strong
irradiance, however, which is relatively costly.
Electrochemical treatment destroys PFAS at the anode by electron transfer and by
hydroxyl radical generation. Electrons are thought better able to attack chemical bonds in
the fluorocarbon tail of PFAS than oxidants owing to their small size, which affords the
electrons better access between the fluorine atoms tightly packed around the C-F and C-
C bonds. Formation of the hydroxyl radical is likely problematic due to its propensity to
transform relatively innocuous constituents of environmental waters to more toxic forms.
The composition of the anode significantly affects efficiency (Lin et al., 2012b; Zhao et
al., 2013; Schaefer et al., 2015) and boron-doped diamond anodes are often
recommended. Nonetheless, other anode materials have been successful. Yang et al.
(2016) mineralized 90% of PFOA using iron electrodes. Wang et al., (2016) eliminated
99.7%) PFOA using zinc cathode and stainless-steel cathode. Liu et al., (2018b) used an
Al-Zn electrode to degrade PFOA in groundwater and removed 79.4% PFOA within one
hour. Lin et al., (2018) used porous Ti407 ceramic material as an anode for mineralization
of PFOS and PFOA, and within three and two hours of reaction -100% PFOA and 93.1%
PFOS were removed.
Foam fractionation can successfully attain drinking water concentrations of PFAS in
treated water (Meegoda et al., 2020). Air is bubbled through a column of water and foam
fractionate is removed from the surface of the water column. Ozone has been used in
place of air to simultaneously oxidize co-contaminants.
Plasma used in wastewater treatment can be generated using electricity, radiofrequency,
or microwaves (Fridman and Kennedy, 2004; Locke and Thagard, 2017). Unlike
traditional AOPs, plasma from electrical discharges can simultaneously oxidize and
reduce organic molecules (Nzeribe et al., 2019). Efficiency is a function not only of the
plasma source but also the conductivity, temperature, and pH of the water as well as the
chemical structure of the contaminant. Stratton et al., (2017) found plasma was an
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effective treatment technology, but again, oxidation can transform relatively innocuous
constituents of environmental waters to more toxic forms.
Sonification is the use of soundwaves to create bubbles in water, the surface of which
attract PFAS. When the bubbles cavitate extreme heat is generated, destroying PFAS
without formation of by-products. The cavitation however also creates reactive species
like those of ARP. These radicals usually remain within the bubble or on the bubble
surface and include -H and -OH, and if they escape the bubble may recombine to form
hydrogen peroxide, or «H may react with oxygen to form the radical 'HO: (Wood et al.,
2017). Nonetheless sonification has been effective at destroying PFAS (Moriwaki et al.,
2005; Cheng et al., 2008), and can be enhanced by addition of sulfate (Lin et al., 2015).
Co-contaminants can reduce efficiency, probably due to competition with PFAS to
occupy the bubble surface, but Cheng et al., (2008) indicated that DOM has little
deleterious effect.
Treatment Train Processes
Groundwater PFAS plumes often include CVOCs, petroleum hydrocarbons, dissolved
organic carbon, and heavy metals. All these materials can reduce the efficiencies of
PFAS treatment technologies. Consequently, a combined treatment approach including
multiple processes may be advantageous for groundwater remediation (Horst et al.,
2018).
Some current work on treatment train approaches includes an in-situ study by Crimi et
al., (2017) and a successful approach used by Boonya-atichart et al., (2018). Crimi et al.,
(2017) used GAC for adsorption and heat-activated persulfate for in-situ oxidative
degradation of PFAS and found persulfate was ineffective. Boonya-atichart et al., (2018)
demonstrated a combined system of photocatalysis and membrane filtration to treat
PFOA from groundwater. A concentrated stream of contaminants obtained from filtration
was mineralized using photocatalysis with nZVI, and a UF system was utilized to
eliminate the nZVI from the photocatalytic route. By this process, 99.6% PFOA was
removed from groundwater, and 59.6% retentate was degraded using photocatalysis.
Lu et al., (2020) reviewed PFAS remediation treatment train approaches based on a
sampling from 150 publications. They classified treatment trains as either tandem
(removal followed by degradation) or parallel (simultaneous destructive mechanisms).
The tandem treatment trains they reviewed included: NF and electrochemical anodic
oxidation, biochar and ZVI, GAC and activated persulfate, GAC and thermal
mineralization, NF, nZVI and UF, and IXR with electrochemical anodic oxidation. The
parallel treatment trains reviewed included: electro-Fenton with electrochemical anodic
oxidation, hydrogen peroxide and activated persulfate, thermolysis and photolysis
degradation, ZVI/GAC micro-electrolysis with vacuum ultraviolet (VUV)-Fenton, and
electron beam with activated persulfate. They found that many innovative technologies in
laboratory development required extreme operating conditions that were not likely to be
cost effective at scale-up. Based on their review they proposed a tandem combination of
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NF with electrochemical anodic oxidation, and a parallel combination of electro-Fenton
degradation with electrochemical anodic oxidation.
SUMMARY
PFAS are a family of thousands of chemicals defined as: fluorinated substances that
contain at least one fully fluorinated methyl or methylene carbon atom (without any
H/Cl/Br/I atom attached to it). They have been used in a multitude of commercial,
industrial, and consumer products but some have been discovered to be very toxic.
They have been synthesized using either ECF or telomerization and the impurities
characteristic of each of these processes can be used for source forensics. A useful
conceptual prototype is a fluorinated, linear carbon-chain backbone with a polar
functional group attached to one end referred to as the head; the opposite end is referred
to as the tail. The tail is hydrophobic and oleophobic and the head is hydrophilic,
rendering the molecule an amphiphilic surfactant with a propensity to collect at media
interfaces such as between air and water. The tail can undergo hydrophobic interactions
with materials such as organic carbon particles in soil, and if the head is in a charged
state, it can undergo electrostatic interactions with materials such as charged mineral
surfaces or ions.
PFAS are commonly observed in environmental media including soil, surface water and
groundwater. PFAS may enter the environment through releases from industrial and
manufacturing facilities, the direct use of PFAS products such as AFFF at airports and
military bases, landfill leachate, WWTP effluent, land application of WWTP biosolids
and sludges for soil improvement, or irrigation using WWTP effluent. Multiple PFAS
sources sometimes exist at a single site, and many releases include a mixture of different
PFAS as well as co-contaminants such as hydrocarbons and CVOCs.
The movement of PFAS that enter the subsurface environment is retarded by their
affinity for interfaces between media such as air and water, hydrophobic interactions, and
electrophilic interactions. These retarding properties vary for individual PFAS and can
contribute to a chromatogram-like redistribution of originally homogeneous PFAS
mixtures. Redistribution can also be affected by the presence of co-contaminants and the
transformation of classes of PFAS (termed precursors) into more stable classes of PFAS.
The complexity of PFAS retardation and transformation in the subsurface can make
development of a CSM challenging. Conducting subsurface sampling is atypically
difficult because PFAS may be incorporated into the composition and coatings of
sampling equipment and supplies. Also, analytical methods are limited to either
measuring only several dozen individual PFAS or relying on some strong assumptions to
estimate a total amount of PFAS.
Some PFAS do not respond well to traditional remedial techniques due to properties such
as hydrophobicity, oleophobicity, and exceptional chemical stability. Numerous
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innovative remedial technologies are in development, but many have only been
demonstrated at the laboratory bench scale or do not seem economically viable for scale-
up to practical implementation. Treatment train approaches have been trialed to overcome
some of the limitations of specific remedial methods but have been similarly criticized.
Currently GAC and IXR are the primary means of PFAS removal, although RO and NF
may become more commonly used processes due to their greater removal efficiency if
they can be made more economical. Both GAC and IXR have several advantages and
disadvantages. Regeneration of GAC is relatively economical and therefore makes this
technology more attractive, but IXR regeneration is somewhat costly due to the need for
concentrate disposal, if regeneration is even possible. Until very recently a common
problem with these materials was preferential adsorption of some PFAS relative to
others, but this is reportedly less of a problem with newly developed GAC and IXR
intended for the treatment of PFAS. Real-world issues such as groundwater co-
contaminants, organic material, and geochemistry can affect the efficacy of GAC and
IXR. The presence of co-contaminants and organic material can adversely affect the
performance of GAC although this may be less of an issue with IXR, and both GAC and
IXR performance can be affected by solution conditions such as pH and alkalinity.
REFERENCES
3M Corporation. 1999. The Science of Organic Fluor ochemistry. US EPA. OPPT-2002-
0043-0006. Retrieved from
http://www.fluoridealert.org/pesticides/pfos.fr.final.docket.0006.pdf.
Adamson, D. T., A. Nickerson, P. R. Kulkarni, C. P. Higgins, J. Popovic, J. Field, A.
Rodowa, C. Newell, P. DeBlanc, J. J. Kornuc. 2020. Mass-Based, Field-Scale
Demonstration of PFAS Retention within AFFF-Associated Source Areas.
Environmental Science & Technology 54 (24): 15768-15777.
https://doi.ore/10.1021/acs.est.0c04472.
Adamson, D.T., P. R. Kulkarni, A. Nickerson, C. P. Higgins, J. Field, T. Schwichtenberg,
C. Newell, J. J. Kornuc. 2022. Characterization of Relevant Site-Specific PFAS
Fate and Transport Processes at Multiple AFFF Sites. Environmental Advances
Volume 7, 100167.
American Society for Testing Materials (ASTM) D7968. 2017. Standard Test Methodfor
Determination of Polyfluorinated Compounds in Soil by Liquid Chromatography
Tandem Mass Spectrometry (LC/MS/MS). West Conshohocken, PA: ASTM
International https://www.astm.ore/Standards/P?968.htm
American Society for Testing Materials (ASTM) D7979. 2019. Standard Test Methodfor
Determination of Per- andPolyfluoroalkyl in Water, Sludge, Influent, Effluent,
and Wastewater by Liquid Chromatography Tandem Mass Spectrometry
Page 52
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
(LC/MS/MS). West Conshohocken, PA: ASTM International
http s: //www, astm. ore/ Stan dards/D7979. htm.
Anderson, R. H., D. T. Adamson, and H. F. Stroo. 2019. Partitioning of Poly- and
Perfluoroalkyl Substances from Soil to Groundwater within Aqueous Film-
Forming Foam Source Zones. Journal of Contaminant Hydrology 220: 59-65.
Annunziato, K. M., J. Doherty, J. Lee, J. M. Clark, W. L. Liang, C. W. Clark, M.
Nguyen, M. A. Roy, and A. R. Timme-Laragy. 2020. Chemical Characterization
of a Legacy Aqueous Film-Forming Foam Sample and Developmental Toxicity in
Zebrafish (Danio Rerio). Environmental Health Perspectives 128, no. 9.
Appleman, T. D., C. P. Higgins, O. Quinones, B. J. Vanderford, C. Kolstad, J. C. Zeigler-
Holady, and E. R. V. Dickenson. 2014. Treatment of Poly- and Perfluoroalkyl
Substances in Us Full-Scale Water Treatment Systems. Water Research 51: 246-
55.
Appleman, T. D., E. R. V. Dickenson, C. Bellona, and C. P. Higgins. 2013.
Nanofiltration and Granular Activated Carbon Treatment of Perfluoroalkyl
Acids. Journal of Hazardous Materials 260: 740-46.
Arvaniti, O. S., Y. Hwang, H. R. Andersen, A. S. Stasinakis, N. S. Thomaidis, and M.
Aloupi. 2015. Reductive Degradation of Perfluorinated Compounds in Water
UsingMg-Aminoclay CoatedNanoscale Zero Valent Iron. Chemical Engineering
Journal 262: 133-39.
ASTDR. 2021. Toxicological Profile for Perfluoroalkyls. US Department of Health and
Human Services, Agency for Toxic Substances and Disease Registry (ASTDR).
https://www.atsdr.cdc.eov/toxprofiles/tp200.pdf.
Ateia, M., A. Maroli, N. Tharayil, and T. Karanfil. 2019. The Overlooked Short- and
Ultrashort-Chain Poly- and Perfluorinated Substances: A Review. Chemosphere
220: 866-82.
Backe, W. J., T. C. Day, and J. A. Field. 2013. Zwitterionic, Cationic, and Anionic
Fluorinated Chemicals in Aqueous Film Forming Foam Formulations and
Groundwater from US Military Bases by Nonaqueous Large-Volume Injection
HPLC-MS/MS. Environmental Science & Technology 47, no. 10: 5226-34.
Baghirzade, B. S., Y. Zhang, J. F. Reuther, N. B. Saleh, A. K. Venkatesan, and O. G.
Apul. 2021. Thermal Regeneration of Spent Granular Activated Carbon Presents
an Opportunity to Break the Forever PFAS Cycle. Environmental Science &
Technology 55, no. 9: 5608-19.
Banks, R. E., B. E. Smart, J. C. Tatlow. 1994. Organofluorine Chemistry: Principles and
Commercial Applications. Spring Science + Business Media. New York:
Page 53
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Springer. ISBN No. 978-1-4899-1202-2.
Barzen-Hanson, K. A., S. C. Roberts, S. Choyke, K. Oetjen, A. McAlees, N. Riddell, R.
McCrindle, et al. 2017. Discovery of 40 Classes of Per- and Polyfluoroalkyl
Substances in Historical Aqueous Film-Forming Foams (AFFFs) and AFFF-
ImpactedGroundwater. Environmental Science & Technology 51, no. 4: 2047-
57.
Barzen-Hanson, K.A., S.E. Davis, M. Kleber, J.A. Field. 2017. Sorption of Fluor otelomer
Sulfonates, Fluorotelomer Sulfonamido Betaines, and a Fluorotelomer
Sulfonamido Amine in National Foam Aqueous Film-Forming Foam to Soil.
Environ. Sci. Technol. 51 (21): 12394-12404.
https://doi.org/10.1021/acs.est.7b03452
Bekele, D., Y. Liu, M. Donaghey, A. U. Arachchige, S. Chadalavada, R. Naidu. 2020.
Separation and Lithological Mapping of PFAS Mixtures in the Vadose Zone at a
Contaminated Site. Front. Water, 23.
Bentel, M. J., Y. C. Yu, L. H. Xu, Z. Li, B. M. Wong, Y. J. Men, and J. Y. Liu. 2019.
Defluorination of Per- and Polyfluoroalkyl Substances (PFASs) with Hydrated
Electrons: Structural Dependence and Implications to PFAS Remediation and
Management. Environmental Science & Technology 53 (7):3718-3728. doi:
10.1021/acs.est.8b06648.
Bolan, N., B. Sarkar, Y. B. Yan, Q. Li, H. Wijesekara, K. Kannan, D. C. W. Tsang, M.
Schauerte, J. Bosch, H. Noll, Y. S. Ok, K. Scheckel, J. Kumpiene, K. Gobindlal,
M. Kah, J. Sperry, M. B. Kirkham, H. L. Wang, Y. F. Tsang, D. Y. Hou, and J.
Rinklebe. 2021. Remediation of Poly- and Perfluoroalkyl Substances (PFAS)
Contaminated Soils - To Mobilize or to Immobilize or to Degrade? Journal of
Hazardous Materials 401. doi: ARTN 12389210.1016/j.jhazmat.2020.123892.
Boo, C., Y. K. Wang, I. Zucker, Y. Choo, C. O. Osuji, and M. Elimelech. 2018. High
Performance Nanofiltration Membrane for Effective Removal of Perfluoroalkyl
Substances at High Water Recovery. Environmental Science & Technology 52
(13):7279-7288. doi: 10.1021/acs.est.8b01040.
Boonya-atichart, A., S. K. Boontanon, and N. Boontanon. 2018. Study of Hybrid
Membrane Filtration andPhotocatalysis for Removal of Perfluorooctanoic Acid
(PFOA) in Groundwater. Water Science and Technology: 561-569. doi:
10.2166/wst.2018.178.
Brooks, M. C., E. Yarney, J. Huang. 2021. Strategies for Managing Risk due to Back
Diffusion. Groundwater Monitoring and Remediation, Volume 41, Issuel: 76-98.
Brusseau, M. L. 2018. Assessing the Potential Contributions of Additional Retention
Processes to PFAS Retardation in the Subsurface. The Science of the Total
Page 54
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Environment vol. 613-614: 176-185. doi:10.1016/j.scitotenv.2017.09.065
Brusseau, M. L., N. Yan, S. V. Glubt, Y. Wang, W. Chen, L. Ying, B. Dungan, K. C.
Carroll, F. O. Holguin. 2019. Comprehensive Retention Model for PFAS
Transport in Subsurface Systems. Water Research 148: 41-50.
Brusseau, M. L., S. V. Glubt. 2019. The Influence of Surfactant and Solution
Composition on Pfas Adsorption at Fluid-Fluid Interfaces. Water Research 161:
17-26.
Buck, R., J. Franklin, U. Berger, J. Conder, I. Cousins, P. Voogt, A. Jensen, K. Kannan,
S. Mabury, S. van Leeuwen. 2011. Perfluoroalkyl andPolyfluoroalkyl Substances
in the Environment: Terminology, Classification, and Origins. Integrated
Environmental Assessment and Management 7, no. 4: 513-541.
Butt, C. M., D. C. G. Muir, and S. A. Mabury. 2014. Biotransformation Pathways of
Fluorotelomer-BasedPolyfluoroalkyl Substances: A Review. Environmental
Toxicology and Chemistry 33, no. 2: 243-67.
Camdzic, D., R. A. Dickman, D. S. Aga. 2021. Total and class-specific analysis ofper-
andpolyfluoroalkyl substances in environmental samples using nuclear magnetic
resonance spectroscopy. Journal of Hazardous Materials Letters Volume 2,
November 2021, 100023.
Charbonnet, J.A., A. E. Rodowa, N. T. Joseph, J. L. Guelfo, J. A. Field, G. D. Jones, C.
P. Higgins, D. E. Helbling, E F. Houtz. 2021. Environmental Source Tracking of
Per- and Polyfluoroalkyl Substances within a Forensic Context: Current and
Future Techniques. Environmental Science & Technology 55 (11): 7237-7245.
doi: 10.1021/acs.est.0c08506.
Chemguard. 2005. Data Sheet #D10D03010, Revised 09/2005. Mansfield, TX, USA.
Chen, H., S. Chen, X. Quan, Y. Z. Zhao, and H. M. Zhao. 2009. Sorption of
Perfluorooctane Sulfonate (Pfos) on Oil and Oil-Derived Black Carbon:
Influence of Solution Ph and [Ca2+ J. Chemosphere 77 (10): 1406-11.
Chen, X., X. H. Xia, X. L. Wang, J. P. Qiao, H. T. Chen. 2011. A Comparative Study on
Sorption of Perfluorooctane Sulfonate (PFOS) by Chars, Ash and Carbon
Nanotubes. Chemosphere 83, no. 10: 1313-19.
Cheng, J., C. D. Vecitis, H. Park, B. T. Mader, and M. R. Hoffmann. 2008. Sonochemical
Degradation of Perfluorooctane Sulfonate (Pfos) and Perfluorooctanoate (PFOA)
in Landfill Groundwater: Environmental Matrix Effects. Environmental Science
& Technology 42, no. 21: 8057-63.
Chetverikov, S. P., D. A. Sharipov, T. Y. Korshunova, and O. N. Loginov. 2017.
Page 55
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Degradation of Perfluorooctanyl Sulfonate by Strain Pseudomonas
plecoglossicida 2.4-D. Applied Biochemistry and Microbiology 53 (5):533-538.
doi: 10.1134/S0003683817050027.
Chow, S. J., N. Ojeda, J. G. Jacangelo, and K. J. Schwab. 2021. Detection of Ultrashort-
Chain and Other Per- and Polyfluoroalkyl Substances (PFAS) in US Bottled
Water. Water Research 201. doi: ARTN 11729210.1016/j.watres.2021.117292.
Costanza, J., M. Arshadi, L. M. Abriola, K. D. Pennell. 2019. Accumulation ofPFOA and
PFOS at the Air-Water Interface. Environmental Science & Technology Letters 6
(8):487-491. https://doi.org/10.1021/acs.estlett.9b00355.
Crane, R. A., T. B. Scott. 2012. Nanoscale Zero-Valent Iron: Future Prospects for an
Emerging Water Treatment Technology. Journal of Hazardous Materials 211:
112-25.
Crimi, M., T. Holsen, C. Bellona, C. Divine, E. Dickenson. 2017. In Situ Treatment Train
for Remediation of Perfluoroalkyl Contaminated Groundwater: In Situ Chemical
Oxidation ofSorbed Contaminants (ISCO SC). Potsdam, New York: Clarkson
University.
Crone, B. C., T. F. Speth, D. G. Wahman, S. J. Smith, G. Abulikemu, E. J. Kleiner, J. G.
Pressman. 2019. Occurrence of Per- and Polyfluoroalkyl Substances (PFAS) in
Source Water and Their Treatment in Drinking Water. Critical Reviews in
Environmental Science and Technology 49 (24): 2359-2396. doi:
10.1080/10643389.2019.1614848.
Cui, J. K., P. P. Gao, Y. Deng. 2020. Destruction of Per- and Polyfluoroalkyl Substances
(PFAS) with Advanced Reduction Processes (ARPs): A Critical Review.
Environmental Science & Technology 54 (7): 3752-66.
D'Agostino, L. A., S. A. Mabury. 2017. Certain Perfluoroalkyl and Polyfluoroalkyl
Substances Associated with Aqueous Film Forming Foam Are Widespread in
Canadian Surface Waters. Environmental Science & Technology 51 (23): 13603-
13.
D'Agostino, L. A., S. A. Mabury. 2014. Identification of Novel Fluorinated Surfactants in
Aqueous Film Forming Foams and Commercial Surfactant Concentrates.
Environmental Science & Technology 48 (1): 121-29.
D'Hollander, W., P. de Voogt, W. De Coen, L. Bervoets. 2010. Perfluorinatedsubstances
in human food and other sources of human exposure. Rev Environ Contam
Toxicol. 2010;208:179-215. doi: 10.1007/978-1-4419-6880-7_4.
Dams, R., K. Hintzer. 2016. Chapter 1: Industrial Aspects of Fluorinated Oligomers and
Polymers, in: Fluorinated Polymers, Volume 2: Applications, doi:
Page 56
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
10.1039/9781782629368-00001. elSBN: 978-1-78262-936-8
Deng, S. B., Q. A. Yu, J. Huang, G. Yu. 2010. Removal of Perfluorooctane Sulfonate
from Wastewater by Anion Exchange Resins: Effects of Resin Properties and
Solution Chemistry. Water Research 44, no. 18: 5188-95.
Dixit, F., R. Dutta, B. Barbeau, P. Berube, and M. Mohseni. 2021. PFAS Removal by Ion
Exchange Resins: A Review. Chemosphere 272. doi: ARTN 129777
Domingo, J. L., and M. Nadal. 2019. Human Exposure to Per- andPolyfluoroalkyl
Substances (PFAS) through Drinking Water: A Review of the Recent Scientific
Literature. Environ Res 177: 108648.
Du, Z. W., S. B. Deng, Y. Bei, Q. Huang, B. Wang, J. Huang, G. Yu. 2014. Adsorption
Behavior and Mechanism of Perfluorinated Compounds on Various Adsorbents-a
Review. Journal of Hazardous Materials 274: 443-54.
Ebnesajjad, S. 2000. Fluoroplastics Volumes 1 and2. Plastic Design Library. ISBN:
9780815517276.
Efraty, A. 2012. Closed Circuit Desalination Series No-6: Conventional RO Compared
with the Conceptually Different New Closed Circuit Desalination Technology.
Desalination and Water Treatment 41 (1-3): 279-95.
Espana, V. A. A., M. Mallavarapu, Ravi Naidu, R. Naidu. 2015. Treatment Technologies
for Aqueous Perfluorooctane sulfonate (PFOS) and Perfluorooctanoate (PFOA):
A Critical Review with an Emphasis on Field Testing. Environmental Technology
and Innovation 4: 168-81.
Evich, M., M. Davis, J. Mccord, B. Acrey, J. Awkerman, D. Knappe, A. Lindstrom, T.
Speth, C. Tebes-Stevens, M. Strynar, Z. Wang, E. Weber, W. Henderson, J.
Washington. 2022. Per- and Polyfluoroalkyl Substances in the Environment.
Science, Vol 375, Issue 6580. DOI: 10.1126/science.abg9065
Fabregat-Palau, J., M. Vidal, A. Rigol. 2021. Modelling the Sorption Behaviour of
Perfluoroalkyl Carboxylates andPerfluoroalkane Sulfonates in Soils. Science of
The Total Environment, Volume 801: 149343.
Ferrey, M. L., J. T. Wilson, C. Adair, C. M. Su, D. D. Fine, X. Y. Liu, and J. W.
Washington. 2012. Behavior andFate of PFOA and PFOS in Sandy Aquifer
Sediment. Ground Water Monitoring and Remediation 32 (4): 63-71. doi:
10.1111/j. 1745-6592.2012.01395.x.
Field, J., T. Schwichtenberg, R. Deeb, E. Hawley, C. Sayler, D. Bogdan. C. Shaefer, B.
DiGuiseppi, A. Struse. 2021. Assessing the Potential for Bias in PFAS
Concentrations during Groundwater and Surface Water Sampling. Strategic
Page 57
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Environmental Research and Development Program (SRDEP), Project ER19-
1205.
Fitzgerald, N.J., A. Wargenau, C. Sorenson, J. Pedersen, N. Tufenkji, P.J. Novak, M.F.
Simcik. 2018. Partitioning and Accumulation of Perfluoroalkyl Substances in
Model Lipid Bilayers and Bacteria. Environmental Science & Technology 52
(18): 10433-40.
Forsthuber, M., A. M. Kaiser, S. Granitzer, I. Hassl, M. Hengstschlager, H. Stangl, C.
Gundacker. 2020. Albumin Is the Major Carrier Protein for PFOS, PFOA,
PFHXS, PFNA andPFDA in Human Plasma. Environment International 137.
Fridman, A., L.A. Kennedy. 2004. Plasma Physics and Engineering. Boca Raton,
Florida: Taylor and Francis.
Gagliano, E., M. Sgroi, P. P. Falciglia, F. G. A. Vagliasindi, P. Roccaro. 2020. Removal
of Poly- and Perfluoroalkyl Substances (PFAS) from Water by Adsorption: Role
of PFAS Chain Length, Effect of Organic Matter and Challenges in Adsorbent
Regeneration. Water Research 171.
Gauthier, J. R., S. A. Mabury. 2022. Noise-Reduced Quantitative Fluorine NMR
Spectroscopy Reveals the Presence of Additional Per- and Polyfluorinated Alkyl
Substances in Environmental and Biological Samples When Compared with
Routine Mass Spectrometry Methods. Anal Chem. 94 (7): 3278-3286. doi:
10.1021/acs.analchem.lc05107.
Glubt, S.V., M. L. Brusseau. 2021. Contribution of Nonaqueous-Phase Liquids to the
Retention and Transport of Per and Polyfluoroalkyl Substances (PFAS) in Porous
Media. Environmental Science & Technology 55 (6): 3706-3715. doi:
10.1021/acs.est.0c07355.
Gliige, J., M. Scheringer, I. T. Cousins, J. C. DeWitt, G. Goldenman, D. Herzke, R.
Lohmann, C. A. Ng, X. Trieri, Z. Wang. 2020. An Overview of the Uses ofPer-
andPolyfluoroalkyl Substances (PFAS). Environmental Science-Processes &
Impacts 22(12): 2345-73.
Gong, Y. Y., L. Wang, J. C. Liu, J. C. Tang, D. Y. Zhao. 2016. Removal of Aqueous
Perfluorooctanoic Acid (PFOA) using Starch-Stabilized Magnetite Nanoparticles.
Science of the Total Environment 562:191-200. doi:
10.1016/j.scitotenv.2016.03.100.Guelfo, J. L., C. P. Higgins. 2013. Subsurface
Transport Potential of Perfluoroalkyl Acids at Aqueous Film-Forming Foam
(AFFF)-ImpactedSites. Environmental Science & Technology 47 (9):4164-4171.
doi: 10.1021/es3048043.
Hale, S. E., H. P. H. Arp, G. A. Slinde, E. J. Wade, K. Bjorseth, G. D. Breedveld, B. F.
Straith, K. G. Moe, M. Jartun, A. Fteisseter. 2017. Sorbent Amendment as a
Page 58
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Remediation Strategy to Reduce Pfas Mobility and Leaching in a Contaminated
Sandy Soil from a Norwegian Firefighting Training Facility. Chemosphere 171:
9-18.
Hansen, M. C., M. H. Borresen, M. Schlabach, G. Cornelissen. 2010. Sorption of
Perfluorinated Compounds from Contaminated Water to Activated Carbon.
Journal of Soils and Sediments 10 (2): 179-85.
Harding-Marjanovic, K. C., E. F. Houtz, S. Yi, J. A. Field, D. L. Sedlak, L. Alvarez-
Cohen. 2015. Aerobic Biotransformation of Fluorotelomer Thioether Amido
Sulfonate (Lodyne) in AFFF-AmendedMicrocosms. Environmental Science &
Technology 49 (13): 7666-74.
Hatton, J., C. Holton, B. DiGuiseppi. 2018. Occurrence and Behavior ofPer-and
Polyfluoroalkyl Substances from Aqueous Film-Forming Foam in Groundwater
Systems. Remediation 28(2): 89-99.
Higgins, C. P., J. A. Field, C. S. Criddle, R. G. Luthy. 2005. Quantitative Determination
of Perfluorochemicals in Sediments and Domestic Sludge. Environmental Science
& Technology 39 (11): 3946-3956. doi: 10.1021/es048245p.
Higgins, C.P., R.G. Luthy. 2006. Sorption of Perfluorinated Surfactants on Sediments.
Environ. Sci. Technol. 40: 7251-7256.
Higgins, C.P., R.G. Luthy. 2007. Modeling Sorption of Anionic Surfactants onto
Sediment Materials: an A-Priori Approach for Perfluoroalkyl Surfactants and
Linear Alkylbenzene Sulfonates. Environ. Sci. Technol., 41 (9): 3254-3261.
10.1021/es062449j
Hoque, A., M. Khaliquzzaman, M. Hossain, A. Khan, D. Bangladesh. 2002.
Determination of Fluoride in Water Residues by Proton Induced Gamma
Emission Measurements. Research Report 35: 176-84.
Hori, H., A. Yamamoto, E. Hayakawa, S. Taniyasu, N. Yamashita, S. Kutsuna. 2005.
Efficient Decomposition of Environmentally Persistent Perfluorocarboxylic Acids
by Use of Per sulfate as a Photochemical Oxidant. Environmental Science &
Technology 39 (7): 2383-88.
Hori, H., E. Hayakawa, H. Einaga, S. Kutsuna, K. Koike, T. Ibusuki, H. Kiatagawa, R.
Arakawa. 2004. Decomposition of Environmentally Persistent Perfluorooctanoic
Acid in Water by Photochemical Approaches. Environmental Science &
Technology 38 (22): 6118-6124. doi: 10.1021/es049719n.
Hori, H., Y. Nagaoka, A. Yamamoto, T. Sano, N. Yamashita, S. Taniyasu, S. Kutsuna, I.
Osaka, R. Arakawa. 2006. Efficient Decomposition of Environmentally Persistent
Perfluorooctanesulfonate and Related Fluorochemicals Using Zerovalent Iron in
Page 59
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Subcritical Water. Environmental Science & Technology 40, (3): 1049-54.
Hori, H., Y. Nagaoka, M. Murayama, S. Kutsuna. 2008a. Efficient Decomposition of
Perfluorocarboxylic Acids and Alternative Fluorochemical Surfactants in Hot
Water. Environmental Science & Technology 42 (19): 7438-43.
Hori, H., Y. Nagaoka, T. Sano, S. Kutsuna. 2008b. Iron-Induced Decomposition of
Perfluorohexanesulfonate in Sub- and Supercritical Water. Chemosphere 70 (5):
800-06.
Horst, J., J. McDonough, I. Ross, E. Houtz. 2020. Understanding and Managing the
Potential By-Products ofPFAS Destruction. Ground Water Monitoring and
Remediation 40 (2): 17-27. doi: 10.1111/gwmr.l2372.
Horst, J., J. McDonough, I. Ross, M. Dickson, J. Miles, J. Hurst, P. Storch. 2018. Water
Treatment Technologies for PFAS: The Next Generation. Ground Water
Monitoring and Remediation 38 (2): 13-23.
Houtz, E. F., C. P. Higgins, J. A. Field, D. L. Sedlak. 2013. Persistence ofPerfluoroalkyl
Acid Precursors in AFFF-Impacted Groundwater and Soil. Environmental
Science & Technology 47 (15): 8187-95.
Houtz, E. F., D. L. Sedlak. 2012. Oxidative Conversion as a Means of Detecting
Precursors to Perfluoroalkyl Acids in Urban Runoff. Environmental Science &
Technology 46 (17): 9342-49.
Houtz, E. F., R. Sutton, J. S. Park, M. Sedlak. 2016. Poly- and Perfluoroalkyl Substances
in Wastewater: Significance of Unknown Precursors, Manufacturing Shifts, and
Likely AFFFImpacts. Water Research 95: 142-49.
Hu, X. C., D. Q. Andrews, A. B. Lindstrom, T. A. Bruton, Laurel A. Schaider, Philippe
Grandjean, Rainer Lohmann, et al. 2016. Detection of Poly- and Perfluoroalkyl
Substances (PFASs) in U.S. Drinking Water Linked to Industrial Sites, Military
Fire Training Areas, and Wastewater Treatment Plants. Environmental Science
& Technology Letters 3 (10): 344-50.
Huang, K. C., Z. Q. Zhao, G. E. Hoag, A. Dahmani, P. A. Block. 2005. Degradation of
Volatile Organic Compounds with Thermally ActivatedPersulfate Oxidation.
Chemosphere 61 (4): 551-60.
Huang, S., M. Sima, Y. Long, C. Messenger, P. Jaffe. 2022. Anaerobic Degradation of
Perfluorooctanoic acid (PFOA) in Biosolids by Acidimicrobium sp. Strain A6.
Journal of Hazardous Materials 424.
Huang, S., P. Jaffe. 2019. Defluorination of Perfluorooctanoic Acid (PFOA) and
Perfluorooctane Sulfonate (PFOS) by Acidimicrobium sp. Strain A6. Environ Sci
Page 60
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Technol. 53(19): 11410-11419.
Hurley, M. D., M. P. S. Andersen, T. J. Wallington, D. A. Ellis, J. W. Martin, S. A.
Mabury. 2004. Atmospheric Chemistry ofPerfluorinated Carboxylic Acids:
Reaction with OH Radicals and Atmospheric Lifetimes. Journal of Physical
Chemistry A 108 (4): 615-20.
ITRC (Interstate Technology & Regulatory Council). 2022. PFAS Technical and
Regulatory Guidance Document and Fact Sheets PFAS-1. Washington, D.C.:
Interstate Technology & Regulatory Council, PFAS Team. https://pfas-
1 .itrcweb.org/.
Inyang, M., E. R. V. Dickenson. 2017. The Use of Carbon Adsorbents for the Removal of
Perfluoroalkyl Acids from Potable Reuse Systems. Chemosphere 184: 168-175.
doi: 10.1016/j.chemosphere.2017.05.161.
Jafvert, C.T. 1990. Sorption of Organic Acid Compounds to Sediments: Initial Model
Development. Environ. Toxicol. Chem., 9 (10): 1259-1268.
Jin, L., P. Zhang, T. Shao, S. Zhao. 2014. Ferric Ion MediatedPhotodecomposition of
Aqueous Perfluorooctane Sulfonate (PFOS) under UVIrradiation and Its
Mechanism. J Hazard Mater 271: 9-15.
Johnson, R. L., A. J. Anschutz, J. M. Smolen, M. F. Simcik, R. L. Penn. 2007. The
Adsorption of Perfluorooctane Sulfonate onto Sand, Clay, and Iron Oxide
Surfaces. Journal of Chemical and Engineering Data 52 (4): 1165-70.
Kempisty, D., Y. Xing, L. Racz. 2018. Perfluoroalkyl Substances in the Environment:
Theory, Practice, and Innovation. CRC Press. ISBN: 1498764185.
Kibbey, T.C.G., R. Jabrzemskib, D. M. O'Carroll. 2020. Supervised Machine Learning
for Source Allocation of Per- and Polyfluoroalkyl Substances (PFAS) in
Environmental Samples. Chemosphere Volume 252, 126593.
Kinniburgh, D. G., W. H. van Riemsdijk, L. K. Koopal, M. Borkovec, M. F. Benedetti,
M.J. Avena. 1999. Ion Binding to Natural Organic Matter: Competition,
Heterogeneity, Stoichiometry and Thermodynamic Consistency. Colloids Surf. A
Physicochem. Eng. Asp., 151 (1-2): 147-166, 10.1016/S0927-7757(98)00637-2
Kissa, E. 2001. Fluorinated Surfactants and Repellents, Second Edition Revised and
Expanded. Edited by A. T. Hubbard, Surfactant Science Series. New York, New
York: Marcel Dekker, Inc. ISBN No. 0-8247-0472-X.
Knight, E.R., L.J. Janik, D.A. Navarro, R.S. Kookana, M.J. McLaughlin. 2019.
Predicting Partitioning of Radiolabeled 14C-PFOA in a Range of Soils Using
Diffuse Reflectance Infrared Spectroscopy. Sci. Total Environ., 686: 505-513,
Page 61
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
10.1016/j.scitotenv.2019.05.339.
Kucharzyk, K. H., R. Darlington, M. Benotti, R. Deeb, E. Hawley. 2017. Novel
Treatment Technologies for PFAS Compounds: A Critical Review. Journal of
Environmental Management 204: 757-64.
Kwadijk, C. J., I. Velzeboer, A. A. Koelmans. 2013. Sorption of Perfluorooctane
Sulfonate to Carbon Nanotubes in Aquatic Sediments. Chemosphere 90 (5): 1631-
6.
Kwon, B. G., H. J. Lim, S. H. Na, B. I. Choi, D. S. Shin, S. Y. Chung. 2014.
Biodegradation of Perfluorooctane sulfonate (PFOS) as an Emerging
Contaminant. Chemosphere 109: 221-225. doi:
10.1016/j.chemosphere.2014.01.072.
Kwon, Y. N., K. Shih, C. Y. Tang, J. O. Leckie. 2012. Adsorption ofPerfluorinated
Compounds on Thin-Film Composite Polyamide Membranes. Journal of Applied
Polymer Science 124(2): 1042-1049. doi: 10.1002/app.35182.
Lee, H., J. D'eon, and S. A. Mabury. 2010a. Biodegradation ofPolyfluoroalkyl
Phosphates as a Source of Perfluorinated Acids to the Environment.
Environmental Science & Technology 44 (9): 3305-10.
Lee, T., A. Rahardianto, Y. Cohen. 2019. Multi-Cycle Operation of Semi-Batch Reverse
Osmosis (SBRO) Desalination. Journal of Membrane Science 588. doi: ARTN
11709010.1016/j .memsci.2019.05.015.
Lee, Y. C., S. L. Lo, J. Kuo, Y. L. Lin. 2012. Persulfate Oxidation of Perfluorooctanoic
Acid under the Temperatures of20-40 Degrees C. Chemical Engineering Journal
198: 27-32.
Lee, Y. C., S. L. Lo, P. T. Chiueh, Y. H. Liou, M. L. Chen. 2010b. Microwave-
Hydrothermal Decomposition of Perfluorooctanoic Acid in Water by Iron-
Activated Persulfate Oxidation. Water Research 44 (3): 886-92.
Lenka, S. P., M. Kah, L. P. Padhye. 2021. A Review of the Occurrence, Transformation,
and Removal of Poly- and Perfluoroalkyl Substances (PFAS) in Wastewater
Treatment Plants. Water Research 199. doi: ARTN
11718710.1016/j .watres.2021.117187.
Li Y., DP. Oliver, R.S. Kookana. 2018. A Critical Analysis of Published Data to Discern
the Role of Soil and Sediment Properties in Determining Sorption of Per and
Polyfluoroalkyl Substances (PFAS). Sci Total Environ. 628-629: 110-120. doi:
10.1016/j.scitotenv.2018.01.167.
Lin, A. Y. C., S. C. Panchangam, C. Y. Chang, P. K. A. Hong, H. F. Hsueh. 2012a.
Page 62
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Removal of Perfluorooctanoic Acid and Perfluorooctane Sulfonate Via Ozonation
under Alkaline Condition. Journal of Hazardous Materials 243: 272-77.
Lin, H., J. F. Niu, S. T. Liang, C. Wang, Y. J. Wang, F. Y. Jin, Q. Luo, Q. G. Huang.
2018. Development of Macroporous Magneli Phase Ti407 Ceramic Materials: As
an Efficient Anode for Mineralization of Poly- and Perfluoroalkyl Substances.
Chemical Engineering Journal 354:1058-1067. doi: 10.1016/j.cej.2018.07.210.
Lin, H., J. F. Niu, S. Y. Ding, L. L. Zhang. 2012b. Electrochemical Degradation of
Perfluorooctanoic Acid (PFOA) by Ti/Sn02-Sb, Ti/Sn02-Sb/Pb02 and Ti/Sn02-
Sb/Mn02 Anodes. Water Research 46 (7): 2281-89.
Lin, J., S. Lo, C. Hu, Y. Lee, J. Kuo. 2015. EnhancedSonochemicalDegradation of
Perfluorooctanoic Acid by Sulfate Ions. Ultrasonics Sonochemistry 22: 542-547.
Lindstrom, A. B., M. J. Strynar, E. L. Libelo. 2011. Polyfluorinated Compounds: Past,
Present, and Future. Environmental Science & Technology 45 (19): 7954-61.
Liou, J. S., B. Szostek, C. M. DeRito, E. L. Madsen. 2010. Investigating the
Biodegradability of Perfluorooctanoic Acid. Chemosphere 80 (2): 176-83.
Liu, J., L. Weinholtz, L. A. Zheng, M. Peiravi, Y. Wu, D. Chen. 2017. Removal of PFOA
in Groundwater by Fe-0 andMn02 Nanoparticles under Visible Light. Journal of
Environmental Science and Health Part a-Toxic/Hazardous Substances &
Environmental Engineering 52 (11): 1048-1054. doi:
10.1080/10934529.2017.1338889.
Liu, J., S. Mejia Avendano. 2013 .Microbial Degradation of Polyfluoroalkyl Chemicals
in the Environment: A Review. Environ Int 61: 98-114.
Liu, L. F., Y. L. Liu, C. L. Li, R. Ji, X. F. Tian. 2018a. Improved Sorption of
Perfluorooctanoic Acid on Carbon Nanotubes Hybridized by Metal Oxide
Nanoparticles. Environmental Science and Pollution Research 25 (16): 15507-
15517. doi: 10.1007/sl 1356-018-1728-5.
Liu, Y., X. M. Hu, Y. Zhao, J. Wang, M. X. Lu, F. H. Peng, J. Bao. 2018b. Removal of
Perfluorooctanoic Acid in Simulated and Natural Waters with Different Electrode
Materials by Electrocoagulation. Chemosphere 201:303-309. doi:
10.1016/j. chemosphere.2018.02.129.
Llorca, M., G. Schirinzi, M. Martinez, D. Barcelo, M. Farre. 2018. Adsorption of
Perfluoroalkyl Substances on Microplastics under Environmental Conditions.
Environmental Pollution 235: 680-691. doi: 10.1016/j.envpol.2017.12.075.
Locke, B.R., S.M. Thagard. 2017. Electrical Discharge Plasma for Water Treatment. In:
Advanced Oxidation Processes for Water Treatment: Fundamentals and
Page 63
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Applications (in print). London: IWA Publishing.
Lu, D. N., S. Sha, J. Y. Luo, Z. R. Huang, X. Z. Jackie. 2020. Treatment Train
Approaches for the Remediation of Per- and Polyfluoroalkyl Substances (PFAS):
A Critical Review. Journal of Hazardous Materials 386. doi: ARTN
12196310.1016/j.jhazmat.2019.121963.
Lu, X. Y., S. B. Deng, B. Wang, J. Huang, Y. J. Wang, G. Yu. 2016. Adsorption
Behavior and Mechanism of Perfluorooctane Sulfonate on Nanosized Inorganic
Oxides. Journal of Colloid and Interface Science 474: 199-205. doi:
10.1016/j .j cis.2016.04.032.
Lyu Y., M.L. Brusseau, W. Chen, N. Yan, X. Fu, X. Lin. 2018. Adsorption ofPFOA at
the Air-Water Interface during Transport in Unsaturated Porous Media. Environ
Sci Technol. 52 (14): 7745-7753. doi:10.1021/acs.est.8b02348.
Mahinroosta, R., L. Senevirathna. 2020. A Review of the Emerging Treatment
Technologies for PFAS Contaminated Soils. Journal of Environmental
Management 255. doi: ARTN 10989610.1016/j.jenvman.2019.109896.
Martin, D., G. Munoz, S. Mejia-Avendano, S. V. Duy, Y. Yao, K. Volchek, C. E. Brown,
J. X. Liu, S. Sauve. 2019. Zwitterionic, Cationic, and Anionic Perfluoroalkyl and
Polyfluoroalkyl Substances Integrated into Total Oxidizable Precursor Assay of
Contaminated Groundwater. Talanta 195: 533-42.
McDonough, C. A., J. L. Guelfo, C. P. Higgins. 2019. Measuring TotalPFASs in Water:
The Tradeoff between Selectivity and Inclusivity. Curr Opin Environ Sci Health 7:
13-18.
McGuire, M. E., C. Schaefer, T. Richards, W. J. Backe, J. A. Field, E. Houtz, D. L.
Sedlak, J. L. Guelfo, A. Wunsch, C. P. Higgins. 2014. Evidence of Remediation-
Induced Alteration of Subsurface Poly- and Perfluoroalkyl Substance Distribution
at a Former Firefighter Training Area. Environmental Science & Technology 48
(12): 6644-52.
Meegoda, J. N., J. A. Kewalramani, B. Li, R. W. Marsh. 2020. A Review of the
Applications, Environmental Release, and Remediation Technologies of Per- and
Polyfluoroalkyl Substances. International Journal of Environmental Research and
Public Health 17(21).
Merino, N., Y. Qu, R. A. Deeb, E. L. Hawley, M. R. Hoffmann, S. Mahendra. 2016.
Degradation and Removal Methods for Perfluoroalkyl and Polyfluoroalkyl
Substances in Water. Environmental Engineering Science 33 (9): 615-49.
Michigan Department of Environmental Quality (MDEQ). 2018. General PFAS
Sampling Guidance. Lansing, Michigan: MDEQ.
Page 64
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Michigan Department of Environmental Quality (MDEQ). 2020. PFAS Sampling Quick
Reference Field Guide. Lansing, Michigan: MDEQ.
Milinovic J., S. Lacorte, M. Vidal, A. Rigol. 2015. Sorption Behaviour ofPerfluoroalkyl
Substances in Soils. Sci Total Environ. 511: 63-71. doi:
10.1016/j.scitotenv.2014.12.017.
Miyake, Yuichi, Nobuyoshi Yamashita, Pawel Rostkowski, Man Ka So, Sachi Taniyasu,
Paul K. S. Lam, K. Kannan. 2007. Determination of Trace Levels of Total
Fluorine in Water Using Combustion Ion Chromatography for Fluorine: A Mass
Balance Approach to Determine Individual Perfluorinated Chemicals in Water.
Journal of Chromatography A 1143 (1): 98-104.
Moody, C. A., J. A. Field. 2000. Perfluorinated Surfactants and the Environmental
Implications of Their Use in Fire-Fighting Foams. Environmental Science &
Technology 34 (18): 3864-70.
Moriwaki, H., Y. Takagi, M. Tanaka, K. Tsuruho, K. Okitsu, Y. Maeda. 2005.
Sonochemical Decomposition ofPerfluorooctane Sulfonate and
Perfluorooctanoic Acid. Environmental Science & Technology 39 (9): 3388-92.
Munoz, G., J. Liu, S. V. Duy, S. Sauve. 2019. Analysis ofF-53B, Gen-X, ADONA, and
Emerging Fluoroalkylether Substances in Environmental and Biomonitoring
Samples: A Review. Trends in Environmental Analytical Chemistry, Volume 23.
National Ground Water Association (NGWA). 2017. Groundwater and PFAS: State of
Knowledge and Practice. Westerville, Ohio: NGWA.
Newell, C. J., D. T. Adamson, P. R. Kulkarni, B. N. Nzeribe, H. Stroo. 2020. Comparing
PFAS to Other Groundwater Contaminants: Implications for Remediation.
Remediation Journal 30 (3): 7-26.
Nguyen, T.M.H., J. Braunig, K. Thompson, J. Thompson, S. Kabiri, D. A. Navarro, R. S.
Kookana, C. Grimison, C. M. Barnes, C. P. Higgins, M. J. McLaughlin, J. F.
Mueller. 2020. Influences of Chemical Properties, Soil Properties and Solution
pH on Soil-Water Partitioning Coefficients of Per- and Polyfluoroalkyl
Substances (PFAS). Environmental Science & Technology 54(24): 15883-15892.
Nickerson, A., A. C. Maizel, P. R. Kulkarni, D. T. Adamson, J. J. Kornuc, C. P. Higgins.
2020. Enhanced Extraction of AFFF-Associated PFASs from Source Zone Soils.
Environmental Science & Technology 54 (8): 4952-62.
Nickerson, A., A. E. Rodowa, D. T. Adamson, J. A. Field, P. R. Kulkarni, J. J. Kornuc,
C. P. Higgins. 2021. Spatial Trends of Anionic, Zwitterionic, and Cationic PFAS
at an AFFF-ImpactedSite. Environmental Science & Technology 55 (1): 313-23.
Page 65
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Niu, J. F., H. Lin, J. L. Xu, H. Wu, Y. Y. Li. 2012. Electrochemical Mineralization of
Perfluorocarboxylic Acids (PFCAs) by Ce-Doped Modified Porous
Nanocrystalline Pb02 Film Electrode. Environmental Science & Technology 46
(18): 10191-98.
Nzeribe, B. N., M. Crimi, S. M. Thagard, T. M. Holsen. 2019. Physico-Chemical
Processes for the Treatment of Per- and Polyfluoroalkyl Substances (PFAS): A
Review. Critical Reviews in Environmental Science and Technology 49 (10): 866-
915.
O'Hagan, D. 2008. Understanding Organofluorine Chemistry. An Introduction to the C—
FBond. Chem. Soc. Rev. 2008, 37, 308-319.
Ochoa-Herrera, V., J. A. Field, A. Luna-Velasco, R. Sierra-Alvarez. 2016. Microbial
Toxicity and Biodegradability ofPerfluorooctane Sulfonate (PFOS) and Shorter
Chain Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS). Environmental
Science-Processes & Impacts 18 (9): 1236-46.
Ochoa-Herrera, V., R. Sierra-Alvarez. 2008. Removal of Perfluorinated Surfactants by
Sorption onto Granular Activated Carbon, Zeolite and Sludge. Chemosphere 72
(10): 1588-93.
Organisation for Economic Co-operation and Development (ODEC). 2018. Toward a
New Comprehensive Global Database of Per- and Polyfluoroalkyl Substances
(PFASs): Summary report on updating the OECD 2007 list of per- and
Polyfluoroalkyl Substances (PFASs). In OECD Series on Risk Management, No.
39 , Environment, Health and Safety, Environment Directorate, OECD. Report
ENV/JM/MONO(2018)7.
Organisation for Economic Co-operation and Development (OECD). 2002. Hazard
Assessment of Perfluorooctane Sulfonate (PFOS) and its Salts. Paris, France:
OECD. Report ENV/JM/RD(2002)17/FINAL.
Organisation for Economic Co-operation and Development (OECD). 2007. Lists of
PFOS, PFAS, PFOA, PFCA, Related Compounds and Chemicals that may
Degrade to PFCA. Paris, France: OECD Report ENV/JM/MONO(2006)15.
Organization for Economic Co-operation and Development (ODEC). 2020. PFASs and
Alternatives in Food Packaging (Paper and Paperboard) Report on the
Commercial Availability and Current Uses. In: OECD Series on Risk
Management, No. 58, Environment, Health and Safety, Environment Directorate,
OECD.
Organization for Economic Co-operation and Development (ODEC). 2021. Reconciling
Terminology of the Universe of Per- and Poly-fluoroalkyl Substances:
Recommendations and Practical Guidance. In: OECD Series on Risk
Page 66
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Management, No. 61.
Pan, C. G., Y. S. Liu, G. G. Ying. 2016. Perfluoroalkyl Substances (PFASs) in
Wastewater Treatment Plants and Drinking Water Treatment Plants: Removal
Efficiency and Exposure Risk. Water Research 106: 562-70.
Park, M., S. Wu, I. J. Lopez, J. Y. Chang, T. Karanfil, S. A. Snyder. 2019. Adsorption of
Perfluoroalkyl Substances (PFAS) in Groundwater by Granular Activated
Carbons: Roles of Hydrophobicity of PFAS and Carbon Characteristics. Water
Research 170: 115364.
Park, S., L. S. Lee, V. F. Medina, A. Zull, and S. Waisner. 2016. Heat-Activated
Per sulfate Oxidation ofPFOA, 6:2 Fluorotelomer Sulfonate, and PFOS under
Conditions Suitable for in-Situ Groundwater Remediation. 2016. Chemosphere
145: 376-83.
Patterson, C., J. Burkhardt, D. Schupp, E. R. Krishnan, S. Dyment, S. Merritt, L. Zintek,
D. Kleinmaier. 2019. Effectiveness of Point-of-Use/Point-of-Entry Systems to
Remove Per- and Polyfluoroalkyl Substances from Drinking Water. AWWA
Water Sci 1 (2): 1-12.
Paul, A. G., K. C. Jones, A. J. Sweetman. 2009. A First Global Production, Emission,
and Environmental Inventory for Perfluorooctane Sulfonate. Environmental
Science & Technology 43 (2): 386-92.
Pereira, H. C., M. Ullberg, D. B. Kleja, J. P. Gustafsson, L. Ahrens. 2018. Sorption of
Perfluoroalkyl Substances (PFASs) to an Organic Soil Horizon - Effect of Cation
Composition andpH. Chemosphere 207: 183-191.
Phenrat, T., N. Saleh, K. Sirk, R. D. Tilton, G. V. Lowry. 2007. Aggregation and
Sedimentation of Aqueous Nanoscale Zerovalent Iron Dispersions. Environmental
Science & Technology 41 (1): 284-90.
Pignatello, J. J, B. Xing. 1995. Mechanisms of Slow Sorption of Organic Chemicals to
Natural Particles. Environ. Sci. Technol. 30 (1): 1-11.
Place, B. J., J. A. Field. 2012. Identification of Novel Fluorochemicals in Aqueous Film-
Forming Foams Used by the Us Military. Environmental Science & Technology
46 (19): 10859-59.
Prevedouros, K., I. T. Cousins, R. C. Buck, S. H. Korzeniowski. 2006. Sources, Fate and
Transport of Perfluorocarboxylates. Environmental Science & Technology 40
(1): 32-44.
Proffitt, B. 2020. Do's and Dont 's of PFAS Sampling: Guide to PFAS-Free Sampling.
Page 67
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Tempe, Arizona: SGS North America Inc.
Psillakis, E., J. Cheng, M. R. Hoffmann, A. J. Colussi. 2009. Enrichment Factors of
Perfluoroalkyl Oxoanions at the Air/Water Interface. The Journal of Physical
Chemistry A 113 (31):8826-8829. https://doi.org/10.1021/jp902795m.
Qu, Y., C. J. Zhang, F. Li, J. Chen, Q. Zhou. 2010. Photo-Reductive Defluorination of
Perfluorooctanoic Acid in Water. Water Research 44 (9): 2939-47.
Renner, R. 2006. The Long and the Short ofPerfluorinatedReplacements. Environmental
Science & Technology 40 (1): 12-13. doi: 10.1021/es062612a.
Ritter, E. E., M. E. Dickinson, J. P. Harron, D. M. Lunderberg, P. A. DeYoung, A. E.
Robel, J. A. Field, G. F. Peaslee. 2017. PIGE as a Screening Tool for Per- and
Polyfluorinated Substances in Papers and Textiles. Nuclear Instruments &
Methods in Physics Research Section B-Beam Interactions with Materials and
Atoms 407: 47-54.
Ritter, S. K. 2010. Fluorochemicals Go Short. Chemical & Engineering News 88 (5): 12-
17.
Robel, A. E., K. Marshall, M. Dickinson, D. Lunderberg, C. Butt, G. Peaslee, H. M.
Stapleton, J. A. Field. 2017. Closing the Mass Balance on Fluorine on Papers and
Textiles. Environmental Science & Technology 51 (16): 9022-32.
Ross, I., J. McDonough, J. Miles, P. Storch, P. T. Kochunarayanan, E. Kalve, J. Hurst, S.
S. Dasgupta, J. Burdick. 2018. A Review of Emerging Technologies for
Remediation ofPFAS. Remediation Journal 28 (2): 101-26.
Rovero, M., D. Cutt, R. Griffiths, U. Filipowicz, K. Mishkin, B. White, S. Goodrow, R.
T. Wilkin. 2021. Limitations of Current Approaches for Predicting Groundwater
Vulnerability from PFAS Contamination in the Vadose Zone. Groundwater
Monitoring & Remediation 41 (4): 62-75.
Ruyle, B. J., C. P. Thackray, J. P. McCord, M. J. Strynar, K. A. Mauge-Lewis, S. E.
Fenton, E. M. Sunderland. 2021. Reconstructing the Composition of Per- and
Polyfluoroalkyl Substances in Contemporary Aqueous Film-Forming Foams.
Environmental Science & Technology Letters 8 (1): 59-65.
https://doi.org/10.1021/acs.estlett.0c00798.
Sahu, S. P., M. Qanbarzadeh, M. Ateia, H. Torkzadeh, A. S. Maroli, E. L. Cates. 2018.
Rapid Degradation and Mineralization of Perfluorooctanoic Acid by a New
Petitjeanite Bi30(0H)(P04)(2) Microparticle Ultraviolet Photocatalyst.
Environmental Science & Technology Letters 5 (8): 533-538. doi:
10.1021/acs.estlett.8b00395.
Page 68
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Santos, A., S. Rodriguez, F. Pardo, A. Romero. 2016. Use of Fenton Reagent Combined
with Humic Acids for the Removal ofPFOA from Contaminated Water. Science
of the Total Environment 563: 657-63.
Savu, P. M. 2000. Fluorinated Higher Car boxy lie Acids. Kirk-Othmer Encyclopedia of
Chemical Technology. doi:10.1002/0471238961.0612211519012221.a01. ISBN
978-0-471-23896-6.
Schaefer, C. E., C. Andaya, A. Urtiaga, E. R. McKenzie, C. P. Higgins. 2015.
Electrochemical Treatment ofPerfluorooctanoicAcid (PFOA) and
Perfluorooctane Sulfonic Acid (PFOS) in Groundwater Impacted by Aqueous
Film Forming Foams (AFFF). Journal of Hazardous Materials 295: 170-75.
Scheringer, M., X. Trier, I. T. Cousins, P. de Voogt, T. Fletcher, Z. Y. Wang, T. F.
Webster. 2014. Helsingor Statement on Poly- and Per fluorinated Alkyl
Substances (PFASs). Chemosphere 114: 337-39.
Schloffer, F., O. Scherer. 1934. DRP 677091. IG Farbenindustrie.
Schroder, H. F. 2003. Determination of Fluorinated Surfactants and Their Metabolites in
Sewage Sludge Samples by Liquid Chromatography with Mass Spectrometry and
Tandem Mass Spectrometry after Pressurized Liquid Extraction and Separation
on Fluorine-Modified Reversed-Phase Sorbents. Journal of Chromatography A
1020 (1): 131-51.
Schroder, H. F., H. J. Jose, W. Gebhardt, R. F. P. M. Moreira, J. Pinnekamp. 2010.
Biological Wastewater Treatment Followed by Physicochemical Treatment for the
Removal of Fluorinated Surfactants. Water Science and Technology 61 (12):
3208-15.
Schroder, H. F., R. J. Meesters. 2005. Stability of Fluorinated Surfactants in Advanced
Oxidation Processes—a Follow up of Degradation Products Using Flow
Injection-Mass Spectrometry, Liquid Chromatography-Mass Spectrometry and
Liquid Chromatography-Multiple Stage Mass Spectrometry. J Chromatogr A
1082 (1): 110-9.
Schultz, M. M., D. F. Barofsky, J. A. Field. 2004. Quantitative Determination of
Fluorotelomer Sulfonates in Groundwater by LC MS/MS. Environmental Science
& Technology 38 (6): 1828-35.
Schulz, K., M. R. Silva, R. Klaper. 2020. Distribution and Effects of Branched Versus
Linear Isomers ofPFOA, PFOS, and PFHXS: A Review of Recent Literature.
Science of the Total Environment 733.
Science History Institute. 2017. Historical Biography: Roy J. Plunkett. Last Modified
December 14 2017, accessed 2021. https://www.sciencehistorY.ore/historical-
Page 69
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
profile/roy-i -plunkett.
Sepulvado, J. G., A. C. Blaine, L. S. Hundal, C. P. Higgins. 2011. Occurrence and Fate
of Perfluorochemicals in Soil Following the Land Application of Municipal
Biosolids. Environmental Science & Technology 45 (19): 8106-12.
Shahsavari, E., D. Rouch, L. S. Khudur, D. Thomas, A. Aburto-Medina, A. S. Ball.
2021. Challenges and Current Status of the Biological Treatment ofPFAS-
ContaminatedSoils. Frontiers in bioengineering and biotechnology 8: 602040-40.
Sharifan, H., M. Bagheri, D. Wang, J. G. Burken, C. P. Higgins, Y. N. Liang, J. X. Liu,
C. E. Schaefer, J. Blotevogel. 2021. Fate and Transport of Per- and
Polyfluoroalkyl Substances (PFASs) in the Vadose Zone. Science of the Total
Environment 111.
Silva, J. A. K., J. Simunek, J. E. McCray. 2020. A Modified HYDRUS Model for
Simulating PFAS Transport in the Vadose Zone. Water (12) 2758.
doi:10.3390/wl2102758.
Sima, M. W., P. R. Jaffe. 2021. A Critical Review of Modeling Poly- and Perfluoroalkyl
Substances (PFAS) in the Soil-Water Environment. Sci Total Environ. 25:
757:143793. doi: 10.1016/j.scitotenv.2020.143793.
Siriwardena, D. P., M. Crimi, T. M. Holsen, C. Bellona, C. Divine E. Dickenson. 2019.
Influence of Groundwater Conditions and Co-Contaminants on Sorption of
Perfluoroalkyl Compounds on Granular Activated Carbon. Remediation 29 (3):
5-15.
Stonebridge, J., R. Baldwin, N. R. Thomson, C. Ptacek. 2020. Fluoride-Selective
Electrode as a Tool to Evaluate the Degradation of PFAS in Groundwater: A
Bench-Scale Investigation. Ground Water Monitoring and Remediation 40 (2):
73-80.
Stratton, G. R., F. Dai, C. L. Bellona, T. M. Holsen, E. R. V. Dickenson, S. M. Thagard.
2017. Plasma-Based Water Treatment: Efficient Transformation of Perfluoroalkyl
Substances in Prepared Solutions and Contaminated Groundwater.
Environmental Science & Technology 51 (3): 1643-48.
Sudicky, E. A., R. W. Gillham, E. O. Frind. 1985. Experimental Investigation of Solute
Transport in Stratified Porous Media: 1. The Nonreactive Case. Water Resources
Research 21 (7): 1035-1041.
Suthersan, S., J. Quinnan, J. Horst, I. Ross, E. Kalve, C. Bell, T. Pancras. 2016. Making
Strides in the Management of "Emerging Contaminants Groundwater
Monitoring & Remediation 36 (1): 15-25.
Page 70
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Tang, C. Y., Q. S. Fu, A. P. Robertson, C. S. Criddle, J. O. Leckie. 2006. Use of Reverse
Osmosis Membranes to Remove Perfluorooctane Sulfonate (PFOS) from
Semiconductor Wastewater. Environ Sci Technol 40 (23): 7343-9.
Tang, C. Y., Q. S. Fu, C. S. Criddle, J. O. Leckie. 2007. Effect of Flux (Transmembrane
Pressure) and Membrane Properties on Fouling and Rejection of Reverse
Osmosis and Nanofiltration Membranes Treating Perfluorooctane Sulfonate
Containing Wastewater. Environmental Science & Technology 41 (6): 2008-14.
Tang, C. Y., Q. S. Fu, D. W. Gao, C. S. Criddle, J. O. Leckie. 2010. Effect of Solution
Chemistry on the Adsorption of Perfluorooctane Sulfonate onto Mineral Surfaces.
Water Research 44 (8): 2654-62.
Tokranov, A. K., N. Nishizawa, C. A. Amadei, J. E. Zenobio, H. M. Pickard, J. G. Allen,
C. D. Vecitis, E. M. Sunderland. 2019. How Do We Measure Poly- and
Perfluoroalkyl Substances (PFASs) at the Surface of Consumer Products?
Environmental Science & Technology Letters 6 (1): 38-43.
Tow, E.W., M.S. Ersan, S. Kum, T. Le, T. F. Speth, C. Owen, C. Bellona, M. N.
Nadagouda, A. M. Mikelonis, P. Westerhoff, C. Mysore, V. S. Frenkel, V.
deSilva, W. S. Walker, A. K. Safulko A.K., D. A. Ladner. 2021 .Managingand
Treating Per- and Polyfluoroalkyl Substances (PFAS) in Membrane
Concentrates. AWWA Water Science, 2021;el233,
https://doi.org/10.1002/aws2.1233.
Transportation Research Board (TRB). 2017. Use and Potential Impacts of AFFF
Containing PFAS at Airports. Retrieved from
https://www.trb.org/Main/Blurbs/175866.aspx
Trojanowicz, M., A. Bojanowska-Czajka, I. Bartosiewicz, K. Kulisa. 2018. Advanced
Oxidation/Reduction Processes Treatment for Aqueous Perfluorooctanoate
(PFOA) and Perfluorooctane sulfonate (PFOS) - a Review of Recent Advances.
Chemical Engineering Journal 336: 170-99.
Tsitonaki, A., B. Petri, M. Crimi, H. Mosbaek, R. L. Siegrist, P. L. Bjerg. 2010. In Situ
Chemical Oxidation of Contaminated Soil and Groundwater Using Per sulfate: A
Review. Critical Reviews in Environmental Science and Technology 40 (1): 55-
91.
US Department of Defense (DOD). 2014. Chemical & Material Emerging Risk Alert;
Aqueous Film-Forming Foam. U.S. DOD Materials of Evolving Regulatory
Interest Team. Accessed September 1, 2021
https://static.ewe.ore/reports/2019/pfas-dod-timeline/201 I U U kri Mm t.pdf
US EPA. 1990. Groundwater Issue: Basic Concepts of Contaminant Sorption at
Hazardous Waste Sites. EPA/540/4-90/053.
Page 71
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
US EPA. 1996. Soil Screening Guidance: Technical Background Document. Washington,
DC: U.S. Environmental Protection Agency. EPA/540/R95/128.
US EPA. 2003. Perjluorooctanoic Acid (PFOA), Fluorinated Telomers; Request for
Comment, Solicitation of Interested Parties for Enforceable Consent Agreement
Development, andNotice of Public Meeting. 68 Federal Register 18626.
Washington D.C.: USEPA.
US EPA. 2016a. Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA).
Washington, DC: U.S. Environmental Protection Agency. EPA 822-R-16-005
US EPA. 2016b. Drinking Water Health Advisory for Perfluorooctane Sulfonate (PFOS).
Washington, DC: U.S. Environmental Protection Agency. EPA 822-R-16-004.
US EPA. 2017. EPAs Non-CBI Summary Tables for 2015 Company Progress Reports
(Final Progress Reports). PFOA Stewardship Program
https://www.epa.gov/sites/production/files/2017-
02/documents/2 :ba stewardship summary table O.pdf
US EPA. 2018. Risk Management for Per- and Polyfluoroalkyl Substances (PFASs)
under TSCA. Washington, DC: U.S. Environmental Protection Agency.
US EPA. 2019a. Method 533: Determination of Per- and Polyfluoroalkyl Substances in
Drinking Water by Isotope Dilution Anion Exchange Solid Phase Extraction and
Liquid Chromatography/Tandem Mass Spectrometry. Cincinnati, Ohio: U.S.
Environmental Protection Agency. EPA/815-B-19-20.
US EPA. 2019b. Perfluoroalkyl and Polyfluoroalkyl Substances (PFAS) Methods and
Guidance for Sampling and Analyzing Water and Other Environmental Media.
Washington, DC: U.S. Environmental Protection Agency. EPA/600/F-17/022e
(Updated February 2019).
US EPA. 2019c. Sampling for Per- and Polyfluoroalkyl Substances (PFAS) in
Groundwater, Standard Operating Procedures. 14 pp.
US EPA. 2020a. Regional screening levels (RSLs). Environmental Protection Agency.
https://www.epa.gov/risk/regional-screeninglevels-rsls (accessed October 1,
2020).
US EPA. 2020b. Method 537.1. Determination of Selected Per- and polyfluorinated Alkyl
Substances in Drinking Water by Solid Phase Extraction, and Liquid
Chromatography/TandemMass Spectrometry (LC/MS/MS). Washington, DC: U.
S. Environmental Protection Agency, National Center for Environmental
Assessment. EPA/600/R-18/352.
US EPA. 2021a. Human Health Toxicity Values for Perfluorobutane Sulfonic Acid and
Page 72
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Related Compound Potassium Perfluorobutane Sulfonate. 2021. Washington, DC:
U.S. Environmental Protection Agency. EPA/600/R-20/345F.
US EPA. 2021b. SW-846, Test Method 3512: Solvent Dilution of Non-Potable Waters.
Washington, DC: U.S. Environmental Protection Agency.
US EPA. 2021c. SW 846, Validated Test Method 8327: Per-and Polyfluoroalkyl
Substances (PFAS) Using External Standard Calibration and Monitoring (MRM)
Liquid Chromatography/Tandem Mass Spectrometry (LC/MS/MS). Washington,
DC: U.S. Environmental Protection Agency.
US EPA. 2021d. Draft Method 1633: Analysis of Per- and Polyfluoroalkyl Substances
(PFAS) in Aqueous, Solid, Biosolids, and Tissue Samples by LC-MS/MS. EPA
82 l-D-21-001.
US EPA. 2022a. Technical Fact Sheet: Drinking Water Health Advisories for Four PFAS
(PFOA, PFOS, GenX chemicals, and PFBS). US EPA Office of Water.
https://www.epa.eov/sYStem/files/dociiments/2022-06/technical-factsheet-foiir-
US EPA. 2022b. Draft Method 1621: Screening Methodfor the Determination of
Adsorbable Organic Fluorine (AOF) in Aqueous Matrices by Combustion Ion
Chromatography (CIC). EPA 821-D-22-002.
https://www.epa.gov/svstern/files/docurrients/2Q22~Q4/draft~rriethod~1621~for~
screening~aof~in~aqueous~matrices~bv~ci ;
Van der Bruggen, B., L. Lejon, C. Vandecasteele. 2003. Reuse, Treatment, and
Discharge of the Concentrate of Pressure-Driven Membrane Processes.
Environmental Science & Technology 37 (17): 3733-3738. doi:
10.1021/es0201754.
Vecitis, C. D., H. Park, J. Cheng, B. T. Mader, M. R. Hoffmann. 2009. Treatment
Technologies for Aqueous Perfluorooctane sulfonate (PFOS) and
Perfluorooctanoate (PFOA). Frontiers of Environmental Science & Engineering
in China 3, (2): 129-51.
Vu, C. T., T. T. Wu. 2020. Recent Progress in Adsorptive Removal of Per- andPoly-
Fluoroalkyl Substances (PFAS) from Water/Wastewater. Critical Reviews in
Environmental Science and Technology.
Wagner, A., B. Raue, H. J. Brauch, E. Worch, F. T. Lange. 2013. Determination of
Adsorbable Organic Fluorine from Aqueous Environmental Samples by
Adsorption to Polystyrene-Divinylbenzene Based Activated Carbon and
Combustion Ion Chromatography. Journal of Chromatography A 1295: 82-89.
Wang, B., A. Agrawal, M.A. Mills. 2019. Per- and Polyfluoroalkyl Substances inAFFF
Page 73
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Impacted Soil and Groundwater and Their Treatment Technologies. In:
Perfluoroalkyl Substances in the Environment: Theory, Practice, and Innovation.
Kempisty, D. (Ed.), Xing, Y. (Ed.), Racz, L. (Ed.) Boca Raton, Florida: CRC
Press.
Wang, D., X. Lyu, F. Xiao, C. Shen, J. Chen, C. M. Park, Y. Sun, M. Flury. 2021.
Critical Review on the Fate and Transport of Per- and Polyfluoroalkyl
Substances (PFAS) in Subsurface Environments. Earth and Space Science Open
Archive (ESSOAr). (Draft.)
Wang, F., K. M. Shih. 2011. Adsorption of Perfluorooctane sulfonate (PFOS) and
Perfluorooctanoate (PFOA) on Alumina: Influence of Solution pH and Cations.
Water Research 45 (9): 2925-30.
Wang, S. W., J. Huang, Y. Yang, Y. M. Hui, Y. X. Ge, T. Larssen, G. Yu, S. Deng, B.
Wang, C. Harman. 2013. First Report of a Chinese PFOS Alternative Overlooked
for 30 Years: Its Toxicity, Persistence, and Presence in the Environment.
Environmental Science & Technology 47 (18): 10163-70.
Wang, Y. J., H. Lin, F. Y. Jin, J. F. Niu, J. B. Zhao, Y. Bi, Y. Li. 2016.
Electrocoagulation Mechanism of Perfluorooctanoate (PFOA) on a Zinc Anode:
Influence of Cathodes and Anions. Science of the Total Environment 557: 542-
550. doi: 10.1016/j.scitotenv.2016.03.114.
Wang, Z. Y., J. C. DeWitt, C. P. Higgins, I. T. Cousins. 2017. A Never-Ending Story of
Per- and Polyfluoroalkyl Substances (PFASs)? Environmental Science &
Technology 51 (5): 2508-18.
Wanninayake, D. M. 2021. Comparison of Currently Available PFAS Remediation
Technologies in Water: A Review. J Environ Manage. 283:111977. doi:
10.1016/j.jenvman.2021.111977.
Weber, A. K., L. B. Barber, D. R. LeBlanc, E. M. Sunderland, C. D. Vecitis. 2017.
Geochemical and Hydrologic Factors Controlling Subsurface Transport ofPoly-
andPerfluoroalkyl Substances, Cape Cod, Massachusetts. Environmental Science
& Technology 51 (8): 4269-79.
Weiner, B., L. W. Y. Yeung, E. B. Marchington, L. A. D'Agostino, S. A. Mabury. 2013.
Organic Fluorine Content in Aqueous Film Forming Foams (AFFF) and
Biodegradation of the Foam Component 6: 2 Fluorotelomermercaptoalkylamido
Sulfonate (6: 2 FTAS). Environmental Chemistry 10 (6): 486-93.
Wetzel, W. 2005. Entdeckungsgeschichte der Polyfluorethylene. N.T.M. 13: 79-91.
https://doi.org/10.1007/s00048-005-0210-x.
Wood, R. J., J. Lee, M. J. Bussemaker. 2017. A Parametric Review of Sonochemistry:
Page 74
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Control and Augmentation of Sonochemical Activity in Aqueous Solutions.
Ultrasonics Sonochemistry 38: 351-370. doi: 10.1016/j.ultsonch.2017.03.030.
Xiao, F., B. Jin, S. A. Golovko, M. Y. Golovko, B. Xing. 2019. Sorption andDesorption
Mechanisms of Cationic and Zwitterionic Per- and Polyfluoroalkyl Substances in
Natural Soils: Thermodynamics and Hysteresis. Environ. Sci. Technol. 53:
11818-11827.
Xiao, F., P. C. Sasi, B. Yao, A. Kubatova, S. A. Golovko, M. Y. Golovko, D. Soli. 2020.
Thermal Stability and Decomposition of Perfluoroalkyl Substances on Spent
Granular Activated Carbon. Environ. Sci. Technol. Lett. 7: 343-350.
Xiao, F., M. F. Simcik, T. R. Halbach, J. S. Gulliver. 2015. Perfluorooctane Sulfonate
(PFOS) and Perfluorooctanoate (PFOA) in Soils and Groundwater of a US
Metropolitan Area: Migration and Implications for Human Exposure. Water
Research 72: 64-74.
Xu, B. T., J. L. Zhou, A. Altaee, M. B. Ahmed, M. A. Johir, J. W. Ren, X. W. Li. 2020a.
Improved Photocatalysis of Perfluorooctanoic Acid in Water and Wastewater by
Ga203/UVSystem Assisted by Peroxymonosulfate. Chemosphere 239. doi:
ARTN 12472210.1016/j. chemosphere.2019.124722.
Xu, B. T., M. B. Ahmed, J. L. Zhou, A. Altaee. 2020b. Visible and IIVPhotocatalysis of
Aqueous Perfluorooctanoic Acid by Ti02 and Per oxymonosulf ate: Process
Kinetics and Mechanistic Insights. Chemosphere 243. doi: ARTN
12536610.1016/j.chemosphere.2019.125366.
Yang, B., Y. N. Han, G. Yu, Q. F. Zhuo, S. B. Deng, J. H. Wu, P. X. Zhang. 2016.
Efficient Removal of Perfluoroalkyl Acids (PFAA) from Aqueous Solution by
Electrocoagulation using Iron Electrode. Chemical Engineering Journal 303:
384-390. doi: 10.1016/j cej.2016.06.011.
Yi, L. B., L. Y. Chai, Y. Xie, Q. J. Peng, Q. Z. Peng. 2016. Isolation, Identification, and
Degradation Performance of a PFOA-Degrading Strain. Genetics and Molecular
Research 15 (2). doi: ARTN 1502804310.4238/gmr. 15028043.
Yin, P. H., Z. H. Hu, X. Song, J. G. Liu, N. Lin. 2016. Activated Per sulfate Oxidation of
Perfluorooctanoic Acid (PFOA) in Groundwater under Acidic Conditions.
International Journal of Environmental Research and Public Health 13 (6).
You, C., C. Jia, G. Pan. 2010. Effect of Salinity and Sediment Characteristics on the
Sorption and Desorption of Perfluorooctane Sulfonate at Sediment-Water
Interface. Environ Pollut 158 (5): 1343-7.
You, X., S. Liu, C. Dai, Y. Guo, G. Zhong, Y. Duan. 2020. Contaminant Occurrence and
Migration Between High- and Low-Permeability Zones in Groundwater Systems:
Page 75
-------
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
A Review. Sci Total Environ. 743:140703. doi: 10.1016/j.scitotenv.2020.140703.
Young, C. V., V. I. Furdui, J. Franklin, R. M. Koerner, D. C. G. Muir, S.A. Mabury.
2007. Perfluorinated Acids in Arctic Snow: New Evidence for Atmospheric
Formation. Environ. Sci. Technol. 41: 3455-346.
Yu, Q., R. Q. Zhang, S. B. Deng, J. Huang, G. Yu. 2009. Sorption ofPerfluorooctane
Sulfonate and Perfluorooctanoate on Activated Carbons and Resin: Kinetic and
Isotherm Study. Water Research 43 (4): 1150-58.
Zhang, D. Q., Q. C. He, M. Wang, W. L. Zhang, Y. N. Liang. 2021. Sorption of
Perfluoroalkylated Substances (PFASs) onto Granular Activated Carbon and
Biochar. Environmental Technology 42 (12): 1798-1809. doi:
10.1080/09593330.2019.1680744.
Zhang, D. Q., W. L. Zhang, Y. N. Liang. 2019. Adsorption of Perfluoroalkyl and
Polyfluoroalkyl Substances (PFASs) from Aqueous Solution - a Review. Science
of the Total Environment 694.
Zhang, W. P., S. M. Pang, Z. Q. Lin, S. Mishra, P. Bhatt, S. H. Chen. 2021.
Biotransformation of Perfluoroalkyl Acid Precursors from Various
Environmental Systems: Advances and Perspectives. Environmental Pollution
272.
Zhang, Z., D. Sarkar, J. K. Biswas, R. Datta. 2022. Biodegradation of Per- and
Polyfluoroalkyl Substances (PFAS): A Review. Bioresource Technology, Volume
344, Part B, 126223.
Zhao, C. W., C. Y. Tang, P. Li, P. Adrian, G. S. Hu. 2016. Perfluorooctane Sulfonate
Removal by Nanofiltration Membrane-The Effect and Interaction of Magnesium
Ion/Humic Acid. Journal of Membrane Science 503:31-41. doi:
10.1016/j.memsci.2015.12.049.
Zhao, H. Y., J. X. Gao, G. H. Zhao, J. Q. Fan, Y. B. Wang, Y. J. Wang. 2013.
Fabrication of Novel Sn02-Sb/Carbon Aerogel Electrode for Ultrasonic
Electrochemical Oxidation of Perfluorooctanoate with High Catalytic Efficiency.
Applied Catalysis B-Environmental 136: 278-86.
Zhi, Y., J. Liu. 2018. Sorption and Desorption of Anionic, Cationic and Zwitterionic
Polyfluoroalkyl Substances by Soil Organic Matter and Pyrogenic Carbonaceous
Materials. Chemical Engineering Journal 346: 682-691.
Zhou, Q., G. Pan, J. Zhang. 2013. Effective Sorption ofPerfluorooctane Sulfonate
(PFOS) on Hexadecyltrimethylammonium Bromide ImmobilizedMesoporous
Si02 Hollow Sphere. Chemosphere 90 (9): 2461-2466.
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EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Zhou, Q., S. Deng, Q. Zhang, Q. Fan, J. Huang, G. Yu. 2010. Sorption of
Perfluorooctane Sulfonate andPerfluorooctanoate on Activated Sludge.
Chemosphere 81: 453-458.
Zhou, Y. S., Z. Y. He, Y. Tao, Y. H. Xiao, T. T. Zhou, T. Jing, Y. K. Zhou, S. R. Mei.
2016. Preparation of a Functional Silica Membrane Coated on Fe304
Nanoparticle for Rapid and Selective Removal of Perfluorinated Compounds from
Surface Water Sample. Chemical Engineering Journal 303: 156-166. doi:
10.1016/j .cej .2016.05.137.
Zintek, L. B., D. Kleinmaier, D. J. Wesolowski, S. Bonina, C. Acheson. 2017. Region 5
CRL Methods for the Analysis ofPolyfluorinated Compounds (PFCs) Using a
Quick Sample Extraction/Preparation Followed by UPLC/MS/MS Analysis. U.S.
Environmental Protection Agency (U.S. EPA) Region 5 Chicago Regional
Laboratory (CRL).
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Table 1. Selected properties of example PFAS. Data downloaded on 07/07/2021 from the US EPA CompTox Chemicals Dashboard, available at: https://comptox. epa.sov/dashboard.
Property
Units
HFPO-DA
PFBA
PFBS
PFDA
PFHxS
PFNA
PFOA
PFOS
Boiling Point
°C
68.3 to 106
108 to 123
205 to 214
205 to 239
218 to 238
190 to 222
188 to 204
219 to 244
Density
g/cm3
1.69 to 1.70
1.55 to 1.68
1.81 to 1.85
1.76 to 1.82
1.84
1.75 to 1.80
1.70 to 1.75
1.84 to 1.85
Flash Point
°C
18.7 to 20.3
18.0 to 47.3
#N/A
79.6 to 82.7
#N/A
72.6 to 74.0
62.1 to 73.9
#N/A
Henry's Law
atm-
m3/mole
0.0201
0.0000501
2.95E-10
1.5E-10
1.94E-10
1.18E-09
1.92E-10
1.8E-11
Index of Refraction
unitless
1.27
1.29
1.32
1.29
1.31
1.29
1.29
1.3
LogKoa: Octanol-
Air
unitless
2.29
3.46
4.16
4.28
4.27
4.2
4.16
4.75
LogKow: Octanol-
Water
unitless
3.37 to 9.12
1.43 to 3.93
1.95 to 3.68
4.15 to 9.53
2.20 to 5.25
3.54 to 8.64
3.11 to 7.75
4.17 to 7.03
Melting Point
°C
-107 to -53.3
-17.9 to 13.5
20.4 to 106
5.98 to 90.0
26.7 to 190
4.71 to 68.5
-8.69 to 54.2
15.2 to 185
Molar Refractivity
cm3
33.4
23.2
32
52.7
41.8
47.8
42.9
51.5
Molar Volume
cm3
197
127
162
292
217
265
237
272
Polarizability
A3
13.2
9.19
12.7
20.9
16.6
19
17
20.4
Surface Tension
dyn/cm
14.7
15.4 to 18.7
23.4
16.4
21
16.6
16.8
19.6
Vapor Pressure
mm Hg
18.2 to 41.0
3.92 to 33.6
1.14e-8 to
0.208
1.46e-3 to
4.63e-2
8.19E-09
8.44e-3 to
0.171
0.111 to
0.345
0.00000248
mol/L
4.81e-4 to
1.53e-3 to
-0.523 to
2.62e-10 to
1.49e-6 to
4.06e-9 to
6.27e-8 to
6.25e-9 to
Water Solubility
7.00
1.37
7.25e-3
3.73
0.853
3.35
2.98
2.27
PFOA = Perfluorooctanoic acid
PFOS = Perfluorooctane sulfonic acid
PFBA = Perfluorobutanoic acid
PFBS = Perfluorobutate sulfonic acid
PFNA = Perfluorononanoic acid
PFDA = Perfluorodecanoic acid
PFHxS = Perfluorohexane sulfonic acid
HFPO-DA = Hexafluoropropylene oxide dimeric acid
#N/A = data not available from the source used
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Table 2. Four US EPA Methods used to analyze PFAS analytes in environmental media. US EPA SW-846
Method 8327 addresses 24 PFAS analytes in four aqueous matrices of reagent water, groundwater, surface water,
and wastewater effluent. US EPA Method 537.1 is a solid phase extraction (SPE) LC/MS-MS method for the
determination of 18 selected PFAS in drinking water matrices. US EPA Method 533 addresses "short chain"
PFAS (C4-C12) and can be used to test for 11 additional PFAS in drinking water matrices. US EPA Method 1633
addresses 40 PFAS analytes in wastewater, surface water, groundwater, soil, biosolids, sediment, landfill leachate,
and fish tissue.
Target Analyte Name
Abbreviation
CAS
Number
533
537.1
8327
1633
Perfluoroalkyl carboxylic acids
Perfluorobutanoic acid
PFBA
375-22-4
+
+
+
Perfluoropentanoic acid
PFPeA
2706-90-3
+
+
+
Perfluorohexanoic acid
PFHxA
307-24-4
+
+
+
+
Perfluoroheptanoic acid
PFHpA
375-85-9
+
+
+
+
Perfluorooctanoic acid
PFOA
335-67-1
+
+
+
+
Perfluorononanoic acid
PFNA
375-95-1
+
+
+
+
Perfluorodecanoic acid
PFDA
335-76-2
+
+
+
+
Perfluoroundecanoic acid
PFUnA
2058-94-8
+
+
+
+
Perfluorododecanoic acid
PFDoA
307-55-1
+
+
+
+
Perfluorotridecanoic acid
PFTrDA
72629-94-8
+
+
+
Perfluorotetradecanoic acid
PFTeDA
376-06-7
+
+
+
Perfluoroalkyl sulfonic acids
Perfluorobutanesulfonic acid
PFBS
375-73-5
+
+
+
+
Perfluoropentansulfonic acid
PFPeS
2706-91-4
+
+
+
Perfluorohexanesulfonic acid
PFHxS
355-46-4
+
+
+
+
Perfluoroheptanesulfonic acid
PFHpS
375-92-8
+
+
+
Perfluorooctanesulfonic acid
PFOS
1763-23-1
+
+
+
+
Perfluorononanesulfonic acid
PFNS
68259-12-1
+
+
Perfluorodecanesulfonic acid
PFDS
335-77-3
+
+
Perfluorododecanesulfonic acid
PFDoS
79780-39-5
+
Fluoroteloiner sulfonic acids
1 II. 1 II. 211. 2//-Perfluorohexane sulfonic acid
4:2FTS
757124-72-4
+
+
+
1 II. 1 II. 211. 2//-Perfluorooctane sulfonic acid
6:2FTS
27619-97-2
+
+
+
1II. 1II. 211. 2//-Perfluorodecane sulfonic acid
8:2FTS
39108-34-4
+
+
+
Perfluorooctane sulfonamides
Perfluorooctanesulfonamide
PFOSA
754-91-6
+
+
N-methyl perfluorooctanesulfonamide
NMeFOSA
31506-32-8
+
N-ethyl perfluorooctanesulfonamide
NEtFOSA
4151-50-2
+
Perfluorooctane sulfonainidoacetic acids
N-methyl perfluorooctanesulfonamidoacetic acid
NMeFOSAA
2355-31-9
+
+
+
N-ethyl perfluorooctanesulfonamidoacetic acid
NEtFOSAA
2991-50-6
+
+
+
Perfluorooctane sulfonamide ethanols
N-methyl perfluorooctanesulfonamidoethanol
NMeFOSE
24448-09-7
+
N-ethyl perfluorooctanesulfonamidoethanol
NEtFOSE
1691-99-2
+
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Per- and Polyfluoroether carboxylic acids
Hexafluoropropylene oxide dimer acid
HFPO-DA
13252-13-6
+
+
+
4,8-Dioxa-3//-perfluorononanoic acid
ADONA
919005-14-4
+
+
Perfluoro-3-methoxypropanoic acid
PFMPA
377-73-1
+
+
Perfluoro-4-methoxybutanoic acid
PFMBA
863090-89-5
+
+
Nonafluoro-3,6-dioxaheptanoic acid
NFDHA
151772-58-6
+
+
Ether sulfonic acids
9-Chlorohexadecafluoro-3 -oxanonane-1 -sulfonic acid
9C1-PF30NS
-
+
1 l-Chloroeicosafluoro-3-oxaundecane-l-sulfonic acid
llCl-PF30UdS
-
+
Perfluoro(2-ethoxyethane)sulfonic acid
PFEESA
-
+
Fhioroteloiner carboxylic acids
3-Perfluoropropyl propanoic acid
3:3FTCA
-
+
2//.2//.3//.3//-Pcrriuorooctanoic acid
5:3FTCA
-
+
3-Perfluoroheptyl propanoic acid
7:3FTCA
-
+
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3-Perfluoroheptylpropanoic acid
Fluorotelomer sulfonic acid
Figure 1. Examples of PFAS.
EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Perfluorooctanoic acid
Perfluorooctanesulfonic acid
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EPA 600/R-22/066 I August 2022 I www.epa.gov/research
Aeration
Finished AFFF
Runoff to Surface
Water or Sewer
. Reverts to Foam Solution jf
r V:
\v s*r2?vv2
Residual Soil Impact
infiltration
Vadose
Zone
Water Table
Groundwater
Flow i
Infiltration
&
Leaching
Retained by Sorption
(physical, hydrophobic,
electrostatic, interfacial)
Retardation due to
aquifer matrix
interactions
downgradient
migration
Dissolved PFAS
Figure 2. Conceptual site model for firefighter training source area. Source: Adapted from figure by J. Hale, Kleinfelder
and 1TRC 2022. Used with permission.
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EPA 600/R-22/066 I August 2022 I www.epa.gov/research
PRESORTED
STANDARD
POSTAGE & FEES
PAID EPA
PERMIT NO. G-35
Office of Research and Development
(8101R)Washington, DC 20460
Official Business
Penalty for Private Use
$300
50% postconsumer fiber content processed chlorine free
oEPA
United States
Environmental Protection
Agency
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