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Waste Management 107 (2020) 191–200
Contents lists available at ScienceDirect
Waste Management
journal homepage: www.elsevier.com/locate/wasman
Waste type, incineration, and aeration are associated with per- and
polyfluoroalkyl levels in landfill leachates
Helena M. Solo-Gabriele a,⇑, Athena S. Jones a, Andrew B. Lindstrom b, Johnsie R. Lang c
a
Department of Civil, Architectural, and Environmental Engineering, University of Miami, Coral Gables, FL 33146-0630, USA
U.S. Environmental Protection Agency, Research Triangle Park, NC 27709, USA
c
Oak Ridge Institute for Science and Education, 100 ORAU Way, Oak Ridge, TN 37830, USA
b
a r t i c l e
i n f o
Article history:
Received 4 November 2019
Revised 22 March 2020
Accepted 27 March 2020
Keywords:
PFAS
Leachate
Ash
Municipal solid waste
Construction and demolition
Gas condensate
a b s t r a c t
Per- and polyfluoroalkyl substances (PFAS) are found in many consumer products which will be ultimately disposed in landfills. Limiting environmental contamination and future exposures will require
managing leachates from different types of landfills, each with different PFAS levels depending upon
the source of the waste. The objective of this study was to evaluate the influence of waste type and
on-site treatment on PFAS levels in landfill leachates. Eleven PFAS species (7 carboxylic acids, 3 sulfonic
acids, and 5:3 fluorotelomer carboxylic acid) were evaluated in leachates from municipal solid waste
(MSW), construction and demolition (C&D), MSW ash (MSWA), and a mixture of MSWA and MSW with
landfill gas condensate (MSWA/MSW-GC). Leachates were also analyzed before and after on-site treatment at two of these facilities. Results indicate that MSWA leachate had significantly lower PFAS levels
relative to other leachate types. Lower total PFAS concentrations in MSWA leachates were correlated with
an increase in incineration temperature (R2 = 0.92, p = 0.008). The levels of PFAS in untreated C&D and
untreated MSW leachate were similar. The levels of targeted PFAS species in MSW leachate for one of
the facilities evaluated increased after on-site landfill treatment presumably due to the conversion of
PFAS precursors in the untreated leachate sample.
Ó 2020 Elsevier Ltd. All rights reserved.
1. Introduction
Landfill leachate presents a unique challenge for managing perand polyfluoroalkyl substances (PFAS) from products that have
reached the end of their service life. PFAS are used in many consumer products, including sealants (Favreau et al. 2017), sprays
for textiles (Ye et al. 2015), Teflon parts (U.S. EPA 2018), clothing,
carpet (Lang et al. 2016), ski waxes (Kotthoff et al. 2015), and in
non-stick surfaces such as cookware (U.S. EPA 2018). They are also
found in food packaging such as paper food wrappers and cups
(Wang et al. 2017, Schaider et al. 2017). Aqueous film forming
foam (AFFF) represents another source of PFAS release to the environment (Dauchy et al. 2017, Backe et al. 2013, Houtz et al. 2013).
Widespread uses and their resistance to degradation make management of PFAS in consumer products difficult at the end of their
service lives.
⇑ Corresponding author at: 1251 Memorial Drive, McArthur Engineering Building,
Room 252, Coral Gables, FL 33146 USA.
E-mail addresses: hmsolo@miami.edu (H.M. Solo-Gabriele), a.jones18@umiami.
edu (A.S. Jones), Lindstrom.Andrew@epa.gov (A.B. Lindstrom), Johnsie.lang@arcadis.
com (J.R. Lang).
https://doi.org/10.1016/j.wasman.2020.03.034
0956-053X/Ó 2020 Elsevier Ltd. All rights reserved.
The carbon and fluorine bonds in PFAS are the strongest bond
that is possible with carbon due to the highly electronegative nature of the fluorine (O’Hagan 2008). As a result of the strong bonds,
the C-F chain portion of the molecule is resistant to degradation,
including resistance to hydrolysis, photolysis, and biodegradation
(U.S. EPA, 2014, Schultz et al. 2003, OECD, 2002). While thousands
of PFAS exist in commerce, perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) are the most widely studied.
PFAS have been linked to human health effects. PFAS have been
detected in the blood serum of up to 98% of samples collected
between 1999 and 2004 in a population that is representative of
the U.S. (Calafat et al. 2007). In in-vivo studies with rodents, PFAS
have been linked to thyroid diseases and negative impacts on
immune system responses (Yang et al. 2002). They have been
linked to liver cancers (Lau et al. 2007) and renal cancers among
occupationally exposed individuals (Steenland and Woskie 2012).
In exposed communities, PFAS have also been linked with thyroid
disease (Li et al. 2017b), asthma, impaired lung function (Qin et al.
2017), and testicular tumors (Barry et al., 2013) and cancers of the
kidney and bladder (Li et al. 2017a).
As a result of the public health concerns associated with PFAS,
effective May 2016, a drinking water health advisory of 70 ng/L
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H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
was issued by the U.S Environmental Protection Agency (U.S. EPA)
for the sum of PFOA and PFOS (U.S. EPA 2016a, Hamid et al. 2018).
Some U.S. states have adopted stricter drinking water guidelines.
For example, Vermont has adopted a guideline of 20 ng/L for the
sum of PFOA and PFOS plus three additional species (PFNA, PFHxS,
and PFHpA, defined in Fig. 1). Similarly, New Jersey has adopted a
guideline of 14 ng/L for PFOA and 13 ng/L for PFOS (ASDWA, 2019)
and California has adopted guidelines of 5.1 ng/L for PFOA and
6.5 ng/L for PFOS (CWB, 2019).
Municipal solid waste (MSW) leachates have been documented
with PFOA concentrations on the order of 10000 s ng/L and PFOS on
the order of 1000 s ng/L in the U.S. (Huset et al. 2011, Lang et al.
2017), Canada (Benskin et al. 2012), and Europe (Fuertes et al.
2017, Busch et al. 2010, Perkola and Sainio 2013). A landfill known
to have received waste from PFOA and PFOS industrial processes
documented leachate levels as high as 82,000 ng/L and
31,000 ng/L, respectively (Oliaei et al. 2013). The highest levels
were measured in leachates collected from five landfills in China
with PFOA levels up to 214,000 ng/L (Yan et al. 2015).
The types of landfills used for disposal of waste vary in terms of
their composition and PFOA and PFOS levels. MSW landfills in the
U.S. evaluated by Lang et al. (2017) accepted household waste
including organics, cardboard, glass, paper and plastics. This can
be compared with an Australian study on the presence of PFAS
compounds in landfill leachate (Gallen et al. 2017). Average PFOA
and PFOS levels in the Lang et al. (2017) study were 800 ng/L
and 90 ng/L, respectively, whereas for the Gallen et al. (2017) study
the levels were 300 ng/L and 200 ng/L respectively. Gallen et al.
(2017) also evaluated a second class of landfills containing cardboard, glass, paper and plastics plus construction and demolition
(C&D) wastes (defined as concrete, soil, metals, timber, and plastics) which generated average PFOA and PFOS concentrations of
1400 ng/L and 1100 ng/L, respectively.
Landfill leachates are typically managed via transfer to a
wastewater treatment plant (WWTP). In WWTPs, some PFAS tend
to bioaccumulate in the sludge (typically PFAS with 8 carbon fluoroalkyl chains) (Venkatesan and Halden 2013). Exposures to
humans can occur when PFAS-containing sludge is land applied
which can then runoff into adjacent waterways, impact surficial
aquifers and also possibly impact crops and grazing animals for
Fig. 1. Defined acronyms and structural configuration of a PFAS species analyzed
during the current study.
sludges used on agricultural fields. Lang et al. (2017) and Busch
et al. (2010) found that PFAS concentrations were higher in leachate in comparison to lower levels found in domestic WWTP outflows. As a result, PFAS in leachates can impact receiving water
bodies including wastewater treatment plants and may limit
opportunities for downstream water reuse.
The objective of this study was to evaluate the influence of
waste type and on-site landfill treatment on PFAS levels in landfill
leachates. Samples were analyzed for 11 PFAS species (Fig. 1) in
leachate samples from landfills composed of different waste types
including MSW and C&D as well as MSW ash (MSWA) and a
mixture of MSWA and MSW containing gas condensate (MSWA/
MSW-GC), of which MSWA and GC have never been previously
evaluated for PFAS content. Gas condensate originates from
vapor-liquid separator (knock-out pot) that receives gas from the
landfill gas collection system. This condensate was discharged to
the leachate collection system. At facilities where leachates were
treated on-site at the landfill, samples were collected immediately
prior to and after on-site treatment for comparison.
2. Methods
2.1. Landfill sites
Samples were collected at five different landfill facilities within
Florida, USA (Table 1). Pre-treatment and the ultimate disposal of
leachate differed for each facility. Ultimate disposal at two landfill
facilities consisted of on-site aeration with disposal to a WWTP. For
two other landfill facilities, the leachate was discharged to a
WWTP without on-site treatment. At one facility, the leachate
was discharged to deep well injection without on-site treatment.
Some of the facilities had access to leachate flows from distinct
waste types by landfill cell. Leachate was obtained from cells containing predominantly MSW, predominantly C&D, predominantly
MSWA, and combinations thereof. The characteristics of the incineration facilities producing the ash varied. These variations
included differences in the boiler temperatures used to incinerate
the waste. Although the cells accepted both bottom and fly ash,
the treatment of the fly ash also differed between facilities prior
to its disposal within the landfill cell. A sample was also collected
of leachate (75% MSWA and 25% MSW) mixed with gas condensate.
The ratio of leachate to gas condensate was unknown. C&D landfills
are designed to accept wastes from construction and demolition
activities. Historically the majority of C&D landfills do not have
bottom liners designed to capture leachate. As of 2010, bottom liners are now required within the State of Florida. These C&D landfills, referred to as Class III in Florida, were included within the
C&D category. Class III landfills accept waste (yard trash, C&D
debris, processed tires, asbestos, carpet, cardboard, paper, glass,
plastic, and furniture other than appliances) that are not expected
to produce leachates that pose a threat to public health or the environment as per Florida statutes (FAC, 2016). MSWA landfills accept
ash from incineration for either volume reduction or wasteto-energy purposes. These landfills are also required to maintain
bottom liners. Although not all C&D (inclusive of Class III) landfills
have bottom liners, the landfills targeted as part of this study had
bottom liner systems.
Sample collection was initiated at the participating facilities
after two interviews: a telephone interview and an interview in
person with the facility managers. During the interviews questions
were asked about the type of waste disposed and the possibility of
collecting leachates that corresponded to specific waste types.
From these interviews, a sampling plan was devised to optimize
the isolation of leachates by their types (MSW, C&D, MSWA,
MSWA/MSW-GC) and categorized by leachate age. Leachate age
H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
193
Table 1
Description of: (a) waste in landfill cells from which leachate was collected, (b) age of cell, (c) pre-treatment of ash, and (d) on-site leachate treatment processes. Additional
landfill characteristics such as landfill acreage and yearly average leachate volumes are included in supplemental Table S1.
Facility
ID
Sample ID
Waste Type
Age of cell
(years)
Pre-treatment of Ash and On-site Leachate
Treatment
A
C&D(100%)
C&D(100%)
MSWA/MSW-GC
Untreated C&D (Class III) only.
Untreated C&D (Class III) only.
Gas condensate mixed with leachate from several cells
composed of approx. 75% MSWA & 25% MSW.
MSW ash from cell containing 98% ash and 2% MSW.
26
25
20
Ash: The incinerator that produces the ash runs at a
boiler temperature of 760 to 870 °C. Fly ash is treated
with spray dryer absorbers, activated carbon, selective
non-catalytic reduction, and bag house filtration.
Overall the landfill contains 75% MSW & 25% C&D.
Landfill was separated into old (27-year old) versus
new (6-year old) cells. The leachate from the first
sampling point was a combination from old and new
cells (averaged). Leachate from the second sampling
point came from the old cell only.
Waste at this landfill facility consists of MSWA mixed
with MSW at an approximate proportion of 65:35.The
first sample corresponded to leachate entering the onsite treatment system and the second sample
corresponded to leachate after on-site treatment.
27
27
Ash: Not applicable. No ash disposal at this facility.
34
34
Ash: The incinerator that produces ash runs at a boiler
temperature of 815 to 870 °C. Fly ash is treated with
spray lime absorption, activated carbon, and bag
Ash monofill. Samples came from two different
manholes at the site.
21
18
MSWA(98%)
B
C
D
MSW(75%)/ C&D(25%)
MSW(75%)/ C&D(25%)
MSWA(65%)/ MSW(35%)_U
MSWA(65%)/ MSW(35%)_T
MSWA(100%)
MSWA(100%)
9
Leachate: There is no on-site treatment of leachate.
The leachate is disposed on-site via deep well
injection.
Leachate: No on-site treatment. Leachate sent to
regional wastewater treatment plant (WWTP).
house filtration. Leachate: Leachate was treated onsite using a continuous flow biological aeration
process designed to lower ammonia. The effluent from
the on-site treatment process was sent to a regional
WWTP.
Ash: The incinerator that produced ash runs at a boiler
temperature of 930 to 980 °C. Fly ash was treated with
spray dryer absorption, activated carbon, nitrogen
oxide reduction, and bag house filtration. Leachate: No
on-site treatment. Leachate sent to a regional WWTP.
E
MSW(100%)_U
MSW(100%)_T
The majority of the waste was MSW. The first sample
corresponded to leachate entering the on-site
treatment system and the second sample
corresponded to leachate after on-site treatment.
is defined as the number of years that the contributing landfill area
started to receive waste. If the leachate is a combination of leachates from multiple cells, the age is defined by the date the oldest
cell started to receive waste.
A total of 12 samples were collected across five facilities. They
consisted of one MSWA/MSW-GC sample from predominantly an
ash cell (75% MSWA/25% MSW), two samples from C&D landfills,
four samples from predominantly MSW (2 with 100% MSW and
2 with a mix of 75% MSW/25% C&D) and five samples from predominantly ash landfills (2 with 100% MSWA, 1 with 98%
MSWA/2% MSW, and 2 with 65% MSWA/35% MSW) (Table 1).
2.2. Sample collection methods
Leachate was collected in two half-liter high density polyethylene (HDPE) bottles per sampling location. One collection bottle
was used for subsequent PFAS analysis and the other was used to
measure pH and chemical oxygen demand (COD). The purpose of
these analyses was to define bulk physical–chemical characteristics as the leachates are produced by different waste types.
Samples were poured directly into the collection bottles if spigots were available. A new primary collection bottle, also made of
HDPE, was used when samples were to be collected from manholes
or pump stations. The primary collection bottle was attached to a
stainless-steel hose clamp, which in turn was attached to a zinccoated chain. The primary collection bottle was then lowered into
the manhole/well using the chain and bottle attachment. This
allowed for the collection of leachate samples in wells up to
10 m deep and containing leachate that was only a few centimeters
deep at the bottom. The lower end of the chain was detachable
allowing for replacement of the primary sample collection bottle
39
39
Ash: Not applicable. Leachate: Leachate is treated onsite for ammonia removal using sequential batch
reactors and biological aeration. The effluent from the
on-site treatment process is sent to a regional WWTP.
and lowest chain portion between sampling stations to avoid
cross-contamination.
One trip blank was processed per facility visited. The trip blank
consisted of an HDPE bottle that contained deionized water and
was closed throughout sample collection, storage, and shipment.
In addition, for each leachate sample a sample blank was also collected by opening a bottle containing deionized water during the
time of sampling and then closing it after the sample was collected.
Upon collection, samples were placed in a cooler with ice.
2.3. Laboratory analysis
After collecting samples at each facility, sample bottles were
immediately transported to the University of Miami (UM) laboratory (Coral Gables, FL). An aliquot was removed for the basic physical–chemical parameters of pH and COD at UM. The remaining
sample (earmarked for PFAS analysis) was frozen. The aliquot
was analyzed for pH (Orion Star A211) and for COD using predispensed ampules (Bioscience Inc.) to which 1 mL of 1:10 diluted
sample was added and analyzed spectrophotometrically (Milton
Roy, Spec 20 with calibration standards from 0 to 4,500 mg/L of
COD) as per standard methods (APHA, 2018).
The frozen samples were batched into two sets for PFAS analysis at the U.S. EPA Research Triangle Park (RTP) laboratory (Raleigh,
NC), with one set shipped for analysis during January 2018 and the
second set shipped for analysis during July 2018. The samples were
collected over the course of six months (from December 2018 to
May 2019). The maximum time the samples were stored frozen
was 7 months. Samples at EPA-RTP were placed in a 5°C freezer
upon receipt. Samples were thawed in the refrigerator overnight
prior to analysis of PFAS concentrations. Pre-processing of the
samples for each batch was slightly different. All samples were
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H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
analyzed in the second batch with the exception of the MSWA/
MSW-GC sample.
The pre-processing of the samples after shipment included the
addition of internal standards that were isotopically labeled
(Wellington Laboratories, MPFAC-MXA and MFTA-MXA), followed
by a filtration step (Whatman GF/A glass fiber), and then followed
by a solid phase extraction (SPE) process using Oasis WAX cartridges (Huset et al. 2011, Backe and Field 2012). Transfer of the
samples from the sampling bottles used in the field to the filtration
flask followed EPA standard protocols requiring volume measurement and a methanol rinse (U.S. EPA 2016b). For the first batch
only, 200 mL of sample were processed using Envi-carb cartridges
(Sigma Aldrich). Eluates (5 mL of ammonium hydroxide and
methanol solution at a ratio of 1:1000) from the Oasis
WAX/Envi-Carb cartridge (batch 1) and Oasis WAX (batch 2) were
concentrated to 2 mL by evaporation using nitrogen gas. Sample
aliquots of 100 mL were prepared for analyses with the addition
of 300 mL of 2.5 mM ammonium acetate. For the first batch a calibration curve was prepared using the purchased standards
(Wellington Laboratories, PFAC-MXA: fluorinated acid/sulfonate
mix, FTA-MXA: native telomer mix, 5:3 fluorotelomer carboxylic
acid (FTCA)) with an analytical range of 300–1200 ng/L.
The second batch of samples were diluted with an equal volume
of PFAS-free deionized water after the filtration step but prior to
passing 200 mL of the diluted sample through SPE. The purpose
of the dilution was to bring the leachate concentrations to levels
that would fall within the linear portion of the calibration curves.
For the second analysis date, the calibration curve prepared at
EPA-RTP consisted of a wider range of concentrations (10–
2000 ng/L for FTA-MXA, 50 to 5000 ng/L 5:3 FTCA: fluorotelomer
carboxylic acid, PFAC-MXA 10 to 2000 ng/L). The solid phase
extraction for this batch was pH-adjusted with 2.5 mL of nitric acid
on the WAX cartridge to optimize the recovery of short chain PFAS.
Thus the pre-processing of samples analyzed in the first batch
failed to recover the short chain PFAS species as determined from
low R2 values from the calibration curves resulting from low recovery for the higher concentration standards. The short chain PFAS
species (PFBA and PFPeA) were captured in the second batch. Since
short chain PFAS are less hydrophobic than long chain PFAS, their
retention at higher pHs was lower.
Sample extracts were analyzed by isotopic dilution on a Time of
Flight-Liquid Chromatograph/Mass Spectrometer TOF-LC/MS
(Agilent, 1100 Series). Compound identification was established
using the exact mass of the target PFAS species and the expected
retention time. Extracts from the second batch of samples were
analyzed in duplicate (extracts were analyzed twice). The column
consisted of a Poroshell 120 EC-C8 (2.1 50 mm, 2.7 mm). The flow
rate was 300 mL/min with a gradient consisting of an aqueous
phase (A: 95% deionized water and 5% MeOH in 0.4 mM ammonium formate) and an organic phase (B: 95% methanol and 5% of
deionized water in 0.4 mM ammonium formate). The initial gradient (75% A, 25% B) was ramped to 80% B over 5 min and held for
5 min. This was followed by a second ramp to 100% B for 2 min
and held for 3 min. For both analysis batches, analytical blanks
were also added to the process (300 mL of 2.5 mM ammonium
acetate + 100 mL of MeOH) as a check for contamination during
analysis.
Quality control samples showed that all trip blanks, field blanks,
and analytical blanks were below the limits of detection, except for
PFHxS, which was detected in all trip and field blanks at a factor of
10 below the limit of quantification. For this study, the limit of
quantification (LOQ) was defined as 300 ng/L. Species that were
detected but below the LOQ represent best estimates based upon
the measured signal. The limit of detection (LOD) was defined as
any transition ion peak that was three or more times the noise
level at that compound’s specific retention time. PFHxS was not
detected in the analytical blanks. The presence of PFHxS in the trip
blanks and field blanks, but not the analytical blanks indicates the
source of contamination was related to sample handling or the
extraction procedure. All calibration curves (ranges listed in
methods section) were characterized by correlation coefficients
(R2 values) of 0.99 with the exception of the calibration curve for
PFDA for which the R2 value was 0.98 and for 5:3 FTCA for which
the R2 value was 0.91. Duplicate analyses of the standards were
characterized by coefficients of variation of 2.4% on average. Percent recoveries were between 95% and 100% for all mass labeled
PFAS species measured except for PFDA for which recovery was
89% and 5:3 FTCA for which recovery was 86%.
2.4. Statistical analysis
Statistical differences in the mean values for categories compared were evaluated by t-tests assuming the categories had
unequal variances and applying alpha equal to 0.05. Correlations
between PFAS and environmental factors were computed using
the coefficient of determination, R2, and were considered strong
for R2 greater than 0.7 and weak if less than 0.3. Correlations were
considered significant for p values less than 0.05. Correlations were
tested for significance using t-tests with the null hypothesis that
the independent variable had no correlation with the dependent
variable.
3. Results and discussion
3.1. Leachate characterization
Overall landfill age was not significantly correlated with total
measured PFAS (R2 = 0.16, p = 0.19), whereas the correlations of
physical–chemical parameters, pH and COD, varied with total measured PFAS. Overall a significant and positive correlation was
observed between pH and total PFAS (R2 = 0.55, p = 0.005,
Fig. S1). The pH of the leachates varied from 6.2 to 8.1, with MSWA
leachate at the lowest pH and MSW leachate at the highest pH
(Table 2). The observed trend may therefore be due to the fact that
ash landfills had lower pH as opposed to the impacts of pH directly.
Also notable is the significant and positive correlation between pH
and landfill age (R2 = 0.57, p = 0.004, Fig. S1). These results were
consistent with the low pH for landfills undergoing the younger
acidic phase whereas the higher pH values were consistent with
landfills undergoing the methanogenic phase (Kjeldsen et al.
2002). Thus landfill age, in addition to waste type, may also be confounding the positive correlation between leachate pH and total
PFAS. In the case of ash landfills, the low pH may be due to either
the low buffering capacity of the ash resulting in leachate pH more
consistent with the natural lower pH of rainwater or the age of the
ash monofill (18–21 years) for which the alkalinity may have leached from the ash. Specifically, experiments should be conducted
subjecting a waste type (e.g., MSWA, C&D, and MSW) to solutes
of different pH to determine whether PFAS sorption increases at
lower pH.
The correlation between COD and total PFAS concentrations
was weak and not significant (R2 = 0.005, p = 0.83, Fig. S1). The
COD of the samples ranged from 700 mg/L for the treated
MSWA/MSW leachate, to 14,000 mg/L for the MSWA/MSW-GC leachate (Table 2). The association between landfill age and COD was
also weak and insignificant (R2 = 0.27, p = 0.08). The COD values
tended to be low in comparison to landfills undergoing acidic
phase decomposition. These values were more consistent with
the typical values observed during methanogenic phases (3,000
COD mg/L on average) (Kjeldsen et al. 2002).
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H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
Table 2
Landfill cell composition, age, leachate pH, leachate COD and individual PFAS concentrations for the five facilities visited.
Facility ID Waste Type
COD
PFAS (ng/L)à
Waste
Age
Proportions (years) pH (mg/L)
PFBA PFPeA PFHxA PFHpA PFOA PFNA PFDA PFBS
A
C&D
100%
26
A
C&D
100%
25
B
MSW and C&D
75:25
17
B
MSW and C&D
75:25
27
E
MSW untreated
100%
39
100%
39
c
E
MSW treated
C
MSWA/MSW untreated 65:35
34
C
MSWA/MSW treatedc
34
65:35
d
A
A
MSWA/MSW-GC
MSWA/MSW
75:25
98:2
21
12
D
Ash
100
18
D
Ash
100
18
7.6 2700
1170
1150
7.6 2000
1250
1200
7.7 3800
1460
688
7.7 3800
ND
2200
8.1 4600
1410
1660b
8.0 4100
2710
2560
7.5 1800
1380
1450
8.1 700
1290
1380
7.3 14,000 NDe
6.9 8800
1040
917
6.2 4200
421
512
6.4 4300
450
470
1620
1790
1720
322
NDa
ND
ND
ND
ND
ND
2950
3140
990
1150
1050
1040
NDe
1360
1230
652
567
437
477
2190
2250
2200
2130
3560
1830
4270
4240
3570
3590
4290
4295
1691
1720
1610
1630
1140
1770
1680
742
726
589
637
1160
1120
1260
1160
1060
1090
1310
1320
1180
1182
17710
1760
695
722
819
791
299
546
485
328
292
256
255
1720
1740
1750
1680
2200
2290
2860
2860
2620
2643
2990
2960
1177
1166
1610
1596
609
1010
964
360
387
259
269
59
56
58
66
104
116
144
116
119
125
146
154
108
101
106
99
159
160
136
ND
ND
ND
ND
40
40
51
51
121
104
121
167
169
189
256
318
ND
ND
ND
ND
81
105
99
ND
ND
ND
ND
781
828
529
560
3150
3220
ND
ND
3420
3351
2670
2630
331
363
388
386
3800f
5510
4900
508
547
534
552
PFHxS PFOS 5:3 FTCA Total
4130
4230
4630
4530
2250
2330
3560
3580
651
635
643
612
994
992
1400
1390
313
606
540
182
184
179
176
875
874
965
1000
557
600
770
736
875
870
1230
1180
330
319
296
305
720
342
347
166
158
120
124
1930
1900
1650
1760
2540
2540
2990
3050
1590
1600
314
306
748
736
ND
ND
2710
1000
954
ND
ND
ND
ND
15,670
15,960
16,060
14,450
17,010
14,800
16,030
18,270
15,610
15,840
19,970
19,920
8450
8730
8570
8600
9830
13,450
12,260
3360
3370
2820
2960
à
Results correspond to the second batch of analyses for which extracts were analyzed in duplicate (same sample extract which was analyzed twice). The only exception
was the MSWA/MSW-GC sample, which was analyzed with the first batch of samples and only one analysis was available.
a
Not Detected.
b
The internal standard for this sample was in error and so the value listed corresponds to the value without the correction for the internal standard.
c
Leachates that were treated on-site are shown in italics.
d
The proportion of GC in the MSWA/MSW-GC sample is unknown. A thick layer of viscous fluid, presumably GC, was observed at the top of this sample.
e
In the first batch of analyses, the extraction was not optimized to measure the low carbon PFAS (PFBA and PFPeA) and so these measured as non-detects for the gas
condensate.
f
The PFBS concentration for the gas condensate sample was above the limit of the calibration curve so the value listed is an estimate.
3.2. Total PFAS by leachate type and MSWA trends with incineration
temperature
Among the factors evaluated, waste type appears to have the
largest impact on leachate total PFAS levels (sum of the 11 PFAS
measured in the current study) (Fig. 2). The largest concentration
differences were observed between waste types with MSWA samples characterized by the lowest total PFAS levels (as low
as < 3,400 ng/L) and MSW and C&D leachates characterized by
the highest (greater than 18,000 ng/L). Within waste type, incineration temperature was inversely correlated with total PFAS levels.
To begin with, the ash leachate from facility D had the lowest levels
of total PFAS (<3400 ng/L) relative to other landfills that also contained predominantly ash (p < 0.001). This landfill is a pure ash
monofill with no integration of other waste types. Additionally,
the incinerator temperature (930 to 980 °C) that produced the
ash for this monofill was the highest among all the landfills that
accepted ash. The MSWA landfills that received ash incinerated
at intermediate temperatures (facility C, 815 to 870 °C) had intermediate levels of total PFAS, at 8,400 to 8,700 ng/L. The MSWA
landfill that received ash incinerated at the lowest temperatures
(facility A, 760 to 870 °C) had the highest total PFAS levels among
the MSWA leachates, at 12,300 to 13,500 ng/L. The differences in
total PFAS levels between facility C and facility A was dominated
by a relatively large difference in the concentration of PFBS at facility A (~5,000 ng/L) versus facility C (~350 ng/L). One hypothesis is
that all PFAS species are destroyed by incineration in the same proportion. If this were to be the case, given that not all of the PFAS
species were decreased in a similar proportion between facility A
and C, the differences in total PFAS concentrations may be coincidental. Alternatively, there is a possibility that incineration of PFAS
may result in the production of shorter carbon chain compounds
such as PFBS which would explain the large proportion of this species for the MSWA samples for facility A. Regardless of possible
conversions, the correlation between total PFAS and incineration
temperature for the ash leachates was significant (R2 = 0.92,
p = 0.008), with lower total PFAS concentration associated with
an increase in incineration temperature (Fig. 3). More work is
needed to confirm the trend of decreasing total PFAS concentrations with incineration temperature.
The trend with incineration temperature is consistent with laboratory studies that have shown that PFAS are transformed within
the 500 to 1000 °C range (Krusic et al. 2005, Yamada et al. 2005,
Taylor et al. 2014, Merino et al. 2016). For example, Ellis et al.
(2001) found that fluoropolymers at 500 °C decompose and rearrange to form halogenated organic acids and produce polyfluoro(C3-C14) carboxylic acids. Garcia et al. (2007) found that at
850 °C, C2F6 and CF4 are formed. Feng et al. (2015) described a thermolysis mechanism for a perfluorosulfonic acid membrane that
involved cleavage of both the polymer backbone and its side chains
to produce perfluorocarboxylic acids. As such, the results observed
in Fig. 3 are consistent with the transformation of PFAS to other
species or to the partial destruction of PFAS during the waste incineration process. Further evidence of transformation is provided by
evaluating the ratios of PFBA/PFOA and PFPeA/PFOA (Fig. S2). These
ratios were greater than one, on average, for ash leachate samples
(1.25 and 1.20, respectively) and less than one, on average, for
MSW and C&D leachate samples (0.87 and 0.68, respectively) in
the current study. When evaluating the ash samples in more detail,
the low and medium temperature ash samples (facilities A and C)
showed ratios of PFBA/PFOA of 0.99 to 1.2 and ratios of
PFPeA/PFOA of 0.91 to 1.32, on average by facility. For high
196
H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
Fig. 2. Overall PFAS results for leachates collected from five facilities. All results provided in duplicate on the same sample extract and correspond to the second batch (July
2018) with the exception of the sample containing gas condensate which was analyzed once and corresponded to the first batch (January 2018). Brackets of 2 samples
correspond to duplicates on the extracts of the same leachate sample. The ‘‘U” and ‘‘T” set of samples correspond to untreated (U) leachates and the corresponding treated (T)
effluents. The temperatures indicate the average operating temperature of the facility where the ash was generated.
mation of PFAS species, direct measurement of the exhaust gases
from the waste-to-energy incinerators is warranted to confirm that
PFAS are not being transferred to the gaseous phase and lost to the
atmosphere. All possible releases of PFAS from incineration
facilities should be evaluated in addition to gaseous emissions,
including fly ash, bottom ash, and scrubbing solutions. This should
be a priority for future studies.
3.3. Observations from the MSWA/MSW-GC sample
Fig. 3. Total measured PFAS in ash leachates versus incineration temperatures
(R2 = 0.92, p = 0.008). Horizontal error bars correspond to the range of incineration
temperatures. Vertical error bars correspond to the standard deviation of duplicate
analyses.
temperature ash (facility D) the ratio of PFBA/PFOA was computed
as 1.49 and ratio of PFPeA/PFOA was computed as 1.68. It is possible that the higher incineration temperature resulted in more PFAS
transformation towards shorter C-F chain species relative to the
lower incineration temperature causing this shift in the proportions, with higher proportions at higher temperatures. Given the
evidence from laboratory-based studies concerning the transfor-
The MSWA/MSW-GC sample from facility A originated from a
leachate stream that was receiving predominantly MSWA (75%).
This sample was the only one from the set that was analyzed during the first analysis round (January 2018) that did not capture the
lower carbon chain alkylated PFAS (PFBA and PFPeA), suggesting
that the total PFAS levels could have been higher than those shown
in Fig. 2. Overall, the levels for the MSWA/MSW-GC sample are
consistent with the levels observed in the samples from facility A
(MSWA, 98%) except for the levels of shorter chain PFAS. The total
PFAS levels for the MSWA/MSW-GC sample are consistent with the
total PFAS levels for the MSWA samples at the corresponding landfill (right hand side of Fig. 2). With respect to PFAS species, the
sample with the lowest total levels of measured PFAS (facility D
leachate, high temperature ash monofill) had the lowest levels of
all 11 individual PFAS species. Individual PFAS species for leachate
from the 98% MSWA and the 75:25 MSWA/MSW-GC from facility A
197
H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
were also low for all but the PFBS species. The PFBS levels were
elevated for these three samples (Fig. S3). Overall the MSWA/
MSW-GC sample was very similar to the characteristics of the corresponding leachate sample (MSWA at facility A). This comparison
excludes shorter chain PFCAs which were not captured and analysed in the MSWA/MSW-GC sample. Additional MSWA/MSW-GC
samples should be collected to confirm trends.
3.4. MSW versus C&D leachates
The total PFAS concentration for the non-ash cells varied from
14,000 to 20,000 ng/L. The total PFAS levels between C&D (mean
of 15,530 ng/L, standard deviation (SD) of 740 ng/L) and MSW
landfill (mean of 15,730 ng/L, SD of 170 mg/L) types were not statistically different (p = 0.65). The finding that total PFAS levels in
C&D and MSW leachates were similar in this study differs from
the results in the studies by Gallen et al. (2016, 2017) who found
that C&D leachates had higher levels of PFAS than the MSW leachates by about a factor of 3. The elevated levels of PFAS in C&D
is relevant given that many C&D landfills are not designed to the
same standards as MSW landfills. Liners at C&D landfills tend to
be less elaborate (e.g., single liners with single drainage systems)
with some of the older C&D landfill facilities without bottom liners.
Groundwater in the vicinity of both MSW and C&D landfills should
be monitored for PFAS to evaluate the impacts of the combined
effects of bottom systems and waste type.
3.5. On-site landfill leachate treatment
With respect to on-site leachate treatment via aeration, one onsite treatment system resulted in an increase in PFAS concentrations (MSW(100%) at facility E, p = 0.02) whereas the other facility
with leachate treatment (MSWA(65%)/MSW(35%) at facility C) did
not result in total PFAS levels that were statistically different
(p = 0.99) between before and after treatment. This difference
may be due to differences in leachate types. Facility E treated
100% MSW leachate and had an increase in PFAS levels. Facility C
treated predominantly MSWA, the chemistry of which could have
responded differently to aeration. The similar concentrations in the
untreated and treated leachate from facility C suggests that the ash
contained fewer precursors.
The mean concentrations of total PFAS for facility E were
15,730 ng/L and 19,940 ng/L, before and after treatment, respectively. These results of on-site leachate treatment are consistent
with studies at WWTPs (Arvaniti and Stasinakis 2015). For example, Bossi et al. (2008) found that PFAS increased in the effluent
water at a plant that received municipal wastewater. This increase
was attributed to the degradation of fluorinated precursors such as
8:2 FTOHs to form PFOA and 6:2 FTOH to form PFHxA (Xiao et al.
2012).
et al. 2017, Allred et al. 2015). Lang et al. (2017) found that, 5:3
FTCA represented a major component of PFAS in untreated landfill
leachate (400 to 1,500 ng/L). Among the different leachate types,
MSW in the current study had the highest levels of 5:3 FTCA (maximum of 3050 ng/L for facility B). Ash leachates had no measurable
levels of 5:3 FTCA suggesting the loss or conversion of this precursor during the incineration process. Treated leachates also had
lower 5:3 FTCA levels relative to untreated leachates (p < 0.001).
This difference is particularly evident for the MSW (100%) leachate
where concentrations of 5:3 FTCA decreased by a factor of 5 (from
1600 to 310 ng/L, Fig. 4, top panel) after treatment. There could be
several reasons for the lower values of 5:3 FTCA in leachates after
treatment. The lower values could be due to volatilization resulting
in a loss of 5:3 FTCA to the atmosphere, differential sorption resulting in greater adherence to the particulate phase thus removing
5:3 FTCA from solution, or the conversion of FTCA to other PFAS
species, in particular to possibly a species with the same perfluorinated five carbon chain backbone, PFPeA. For PFPeA (Fig. 4, bottom
panel), a marked increase in this species was observed for one of
the facilities (Facility E) between untreated and treated MSW leachate. The results observed at facility E are consistent with studies
that focused on transformation pathways in activated sludge
WWTP processes (Wang et al. 2012, Xiao et al. 2012) that showed
a conversion of PFAS from 5:3 FTCA to PFPeA during the treatment
process. Similarly, studies using landfill leachates have observed
the loss of 5:3 FTCA during aeration in constructed wetland systems (Yin et al. 2017). Given the evidence of this conversion, it
would be of interest to evaluate the influence of aeration conditions (temperature, time, air flow rates) on the transformation of
PFAS species. Future studies should include an evaluation of additional PFAS precursors and the possibility of their conversion to
PFAS species. Also research should focus on evaluating the role of
leachate type (100% MSW versus 65:35 MSWA/MSW) as the
increase in PFPeA was observed for only one of the two facilities.
When comparing levels of individual PFAS species to the
national average of MSW leachates identified by Lang et al.
5:3 FTCA
PFPeA
3.6. PFAS species
When evaluating the carboxylated PFAS species, the treated
MSW leachate from facility E had the highest carboxylated PFAS
concentrations of all 12 leachate samples examined in this study,
from the shortest chain carboxylated PFAS (PFBA, mean concentration of 2640 ng/L) to the longest chain (PFDA, mean concentration
of 290 ng/L) measured. The only exception was PFNA where the
highest concentration was observed in the leachate sample from
facility A (MSWA/MSW-GC) that contained the gas condensate.
Treated MSW leachate concentration (150 ng/L) was still elevated
in PFNA but slightly below the level observed for MSWA/MSW-GC
leachate from facility A (159 ng/L).
Although the number of samples is limited in the current study,
the results are consistent with the findings in other studies (Lang
Fig. 4. Levels of 5:3 FTCA (top panel) and PFPeA (bottom panel) in different types of
landfill leachate. A significant increase in PFPeA is observed between untreated and
treated MSW leachate suggesting a transformation of 5:3 FTCA to this species
during landfill leachate treatment.
198
H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
Fig. 5. Sum of PFAS species for all samples collected, organized by functional groups of carboxylated, sulfonic and FTCA and by number of carbons in the carbon–fluorine
chain. Error bars correspond to two times standard deviations of the sums of the duplicate analyses. The value in the stacked bar corresponds to the sum of the duplicates.
(2017), PFHxS was over an order of magnitude higher in C&D leachates (means of 4,380 ng/L compared to the national average of
350 ng/L). Even for landfills with C&D mixed with MSW, the PFHxS
concentrations were noticeably high (greater than 2,200 ng/L)
overall, suggesting that the source may be associated with C&D
types of waste (Fig. S3). The elevated levels of PFHxS in C&D leachates are consistent with the use of PFHxS as a surfactant coating
for carpets and other building materials (Jin et al. 2011). Such
materials are commonly found in C&D waste and can serve as a
possible source for the elevated PFHxS levels. An additional source
of PFHxS may have included AFFF. PFHxS has been found at firefighting facilities that use these materials during training activities
(Bräunig et al. 2019) and waste contaminated with AFFF, such as
C&D materials disposed from a facility that caught fire, could possibly impact the quality of landfill leachate.
For PFOA and PFOS, two of the most widely studied PFAS, levels
were 2 to 9 times higher in the current study in comparison to
other studies conducted at MSW landfills in the U.S. (Huset et al.
2011, Lang et al. 2017) and in European countries (Fuertes et al.
2017, Busch et al. 2010). However, the concentrations were lower
in comparison to landfill leachates measured in China (Yan et al.
2015). The treated MSW leachate samples were observed to have
the highest PFOA level (~3000 ng/L) and the highest PFOS level
(~1200 ng/L) (Fig. S4). These results are consistent with the predominance of PFOA and PFOS in treated wastewaters (Kwon
et al. 2017). Additionally, PFOA and PFOS were observed in all leachates even for the oldest (39 years) landfill consistent with prior
studies that show that PFOA and PFOS are stable in the environment (Arvaniti and Stasinakis 2015, Hamid et al. 2018), with landfills serving as concentrated sources when discharged to aqueous
systems.
Overall, considering all samples, the 8-carbon species (PFOA
and PFOS) were not the most abundant species (Fig. 5). The 6carbon carboxylated (PFHxA) and sulfonate (PFHxS) species were
the most abundant, with 4-carbon sulfonate (PFBS) a close second.
The U.S. EPA has not released health-based guidance for shortchained PFAS. It would be of interest to develop health-based toxicity values for shorter chain PFAS given their higher abundance in
the leachates evaluated.
4. Conclusion
Overall this study suggests that leachates from ash landfills had
lower levels of PFAS relative to leachates from MSW and C&D land-
fills. The degree to which total measured PFAS levels decreased
correlated with the incineration temperatures used to generate
the ash. However, not all PFAS species decreased with increasing
temperature and work is needed to evaluate the impact of incineration on the distribution of PFAS species in ash leachates. This is
the first paper to show that leachates from ash landfills have lower
PFAS levels relative to leachates from other waste types. Furthermore, the total PFAS levels were correlated with the incineration
temperature of the waste. More work is needed to confirm the correlation with incineration temperature. As part of future work the
fate of PFAS at full-scale incineration facilities should be evaluated
to determine whether the decreased PFAS in ash leachate is due to
destruction of PFAS, transformation to volatile species, or transformation to non-volatile species that were not measured.
Total PFAS levels in C&D and MSW leachates were observed to
be at similar concentrations, indicating that wastes in C&D landfills
could also serve as a source of PFAS release to the environment.
Additionally, C&D leachates exhibited unusually high levels of
PFHxS, compared to other landfill types and consistent with their
use as sealants and water repellants in building materials, emphasizing the need to evaluate all waste types, not only MSW. Given
the observed levels of PFAS in C&D leachates, the impact of C&D
landfills, in particular unlined landfills, to groundwater should be
evaluated in future studies.
This study is also the first to report on PFAS leachate concentrations at field-scale on-site landfill leachate treatment systems. As
observed in other studies at WWTPs, on-site leachate treatment
using aeration processes increased PFAS concentrations presumably due to the conversion of precursors. To further evaluate the
role of precursor conversion, future studies should consider implementing the total oxidizable precursor (TOP) assay which provides
information about all precursors in samples. Such analysis may be
useful in understanding why PFAS may have increased after aeration for facility E but not for facility C. Additionally, precursor data
might add value in terms of differentiating wastes within MSW
versus C&D. Additional work is needed to confirm trends, to evaluate alternative treatment systems, and to establish a mass balance analysis to determine the efficacy for removal of PFAS from
the environment through on-site leachate treatment. Of interest
would be to include non-targeted PFAS analysis to track the majority of PFAS species through the incineration and/or leachate pretreatment process. This type of analysis would be useful for identifying the major PFAS species that can be analyzed on a more routine basis.
H.M. Solo-Gabriele et al. / Waste Management 107 (2020) 191–200
Four of the facilities included in the current study discharge
their leachates (~3,000 – 20,000 ng/L) to WWTPs, two after onsite treatment and two without on-site treatment. One facility
included in the current study disposed its leachate to deep well
injection (~10,000–17,000 ng/L). The fate of PFAS handled through
deep well injection is not known, nor is the overall long-term
impact of this practice. More work is needed to track the fate of
PFAS in leachates currently discharged from landfills.
Declaration of Competing Interest
The authors declared that there is no conflict of interest.
Acknowledgements
Funding for this study was received through the Hinkley Center
for Solid and Hazardous Waste Management. We are thankful to
the landfill staff who facilitated landfill access and appreciate the
input and feedback received from the project Technical Awareness
Group, from the U.S. EPA, and anonymous reviewers. This article
was reviewed in accordance with the policy of the National Exposure Research Laboratory, U.S. Environmental Protection Agency,
and approved for publication. Approval does not signify that the
contents necessarily reflect the view and policies of the Agency,
nor does mention of trade names or commercial products constitute endorsement or recommendation for use.
Appendix A. Supplementary material
Supplementary data to this article can be found online at
https://doi.org/10.1016/j.wasman.2020.03.034.
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