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Bisphenol A in Solid Waste Materials, Leachate Water, and Air Particles from Norwegian WasteHandling Facilities: Presence and Partitioning Behavior


Bisphenol A in Solid Waste Materials, Leachate Water, and Air
Particles from Norwegian Waste-Handling Facilities: Presence and
Partitioning Behavior
Nicolas Morin,†,‡ Hans Peter H. Arp,*,† and Sarah E. Hale†

Department of Environmental Engineering, Norwegian Geotechnical Institute, P.O. Box 3930, Ullevål Stadion, N-0806 Oslo,

Department of Chemistry, Linnaeus väg 6, Umeå University, SE-901 87 Umeå, Sweden
S Supporting Information

ABSTRACT: The plastic additive bisphenol A (BPA) is
commonly found in landfill leachate at levels exceeding acute
toxicity benchmarks. To gain insight into the mechanisms
controlling BPA emissions from waste and waste-handling
facilities, a comprehensive field and laboratory campaign was

conducted to quantify BPA in solid waste materials (glass,
combustibles, vehicle fluff, waste electric and electronic
equipment (WEEE), plastics, fly ash, bottom ash, and
digestate), leachate water, and atmospheric dust from
Norwegian sorting, incineration, and landfill facilities. Solid
waste concentrations varied from below 0.002 mg/kg (fly ash)
to 188 ± 125 mg/kg (plastics). A novel passive sampling
method was developed to, for the first time, establish a set of
waste-water partition coefficients, KD,waste, for BPA, and to quantify differences between total and freely dissolved concentrations
in waste-facility leachate. Log-normalized KD,waste (L/kg) values were similar for all solid waste materials (from 2.4 to 3.1),
excluding glass and metals, indicating BPA is readily leachable. Leachate concentrations were similar for landfills and WEEE/
vehicle sorting facilities (from 0.7 to 200 μg/L) and dominated by the freely dissolved fraction, not bound to (plastic) colloids
(agreeing with measured KD,waste values). Dust concentrations ranged from 2.3 to 50.7 mg/kgdust. Incineration appears to be an
effective way to reduce BPA concentrations in solid waste, dust, and leachate.

Bisphenol A (BPA, 2,2-(4,4′-dihydroxydiphenyl)propane, CAS
Registry No. 80-05-07) is used in vast quantities,1,2 with an
estimated 4.6 million tons being produced globally in 2012.3 Its
primary use is as a monomer in the production of
polycarbonate and epoxy resins. Other uses are as a stabilizing
agent in plastics and as an additive in thermal paper or paper
coatings. BPA is a known endocrine disruptor. Predicted noeffect concentrations (PNEC) for chronic toxicity of 1.6 μg/L
and acute toxicity of 11 μg/L in fresh water have been
proposed in a European Union risk assessment,4 as well as a
soil chronic PNEC of 3700 μg/kg dry weight.5
With so much BPA being produced for use in consumer
products, it is not surprising that BPA is considered ubiquitous
in the environment.6 It is commonly included in environmental
monitoring studies from various countries (for example, The
Netherlands,7,8 China,9 Germany,10,11 Norway,12 Taiwan,13
Japan,14,15 and America16). Klecka et al.17 compiled BPA water
monitoring data from Europe and North America and reported
that median surface fresh water concentrations were notably
below the PNEC, at 0.08 μg/L (n = 1068) and 0.01 μg/L (n =
848) in North America and Europe, respectively. However,
© XXXX American Chemical Society

water levels can commonly be found above the acute PNEC in
landfill leachate. In Norway, a compilation of landfill leachate
data from 2002 to 201212 reported a median of 17 μg/L
(interquartile range, IQR, 1−62 μg/L, maximum 692 μg/L).
Outside of Norway, landfill leachate concentrations range from
0.1 to 17 200 μg/L in diverse Japanese studies18−23 and from
0.01 to 107 μg/L in four Swedish landfills,24 and exceptionally
high leachate concentrations of BPA (4200−25000 μg/L) were
reported in a German study.25 One study found that BPA in
landfills does not decompose under anaerobic conditions,26
implying that landfills can be a persistent source of BPA. In
response, researchers have been prompted to consider
remediation options to lower BPA levels in landfill leachate.27
As an alternative to landfilling, incineration has been found to
be an effective way to remove BPA from waste, as BPA is prone
to thermal degradation above 400 °C.28 A more detailed
Received: March 13, 2015
Revised: May 26, 2015
Accepted: June 9, 2015


DOI: 10.1021/acs.est.5b01307
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of 1, 10, 50, 100, and 1000 μg/L were prepared in 500 mL glass
flasks with glass stoppers by adding Milli-Q water and spiking
with a solution of BPA in ethyl acetate (such that the co-solvent
did not exceed 0.2% of the total volume). To calibrate for two
thicknesses, a 76 μm strip (0.2 g, CS Hyde USA) and a 55 μm
strip39 (0.2 g) were introduced to the flasks; they are referred to
as POM-76 and POM-55, respectively. The flasks were
equilibrated by shaking end-over-end at 13 rpm for a period
of 8 weeks in the dark at room temperature. The POM strips
were removed, and CPOM and Cwater were determined for BPA
as described below. A kinetic experiment was also carried out
for POM-76 at Cwater = 100 μg/L under the same conditions by
placing 10 POM-76 pieces (0.2 g each) in the flasks as above
and removing duplicate strips for BPA quantification at days 3,
7, 15, 21, and 28. BPA spiked into blank control flasks (no
POM added, 100 μg/L spike, triplicate) showed an average
mass loss of BPA of 17 ± 4% over 28 days, which was corrected
for in the kinetic uptake experiments.
Field Campaigns. Waste-handling facilities were chosen to
provide a broad range of waste-handling methods and types of
solid waste fractions. Twelve different facilities in southeastern
Norway were sampled during two or three sampling campaigns,
June−October 2013, October−December 2013, or March−
June 2014. The facilities included three specialized landfills
(accepting bottom ash, fly ash, and sewage-sludge digestate for
composting, though all containing municipal/industrial waste),
two combustible waste-sorting facilities (municipal/industrial
waste), and seven waste electric and electronic equipment
(WEEE)/vehicle shredding and sorting facilities. Due to
requests from some site owners to keep the data anonymous,
the locations are referred to as Landfill A−C, Incineration/
Sorting A,B, and WEEE/Vehicle A−E (only five WEEE/vehicle
locations are assigned, as two sets of two individual facilities
shared water drainage and therefore leachate drainage and air
emissions). Based on logistics or feasibility, solid waste, leachate
water, and air were sampled from these facilities. More details
related to the field sites and sampling campaign are presented
in the SI (Table S2 and Figure S1).
Sampling. Solid waste samples (4−12 kg) were collected by
hand (while wearing nitrile gloves) into 4 L polyurethane bags
from random locations within each facility. Samples were
collected such that they were visually homogeneous and
representative of a particular waste fraction (e.g., coarse/fine
ash, coarse/fine combustibles, cable plastics, etc.). Samples
were transported back to the laboratory and stored at 4 °C until
further processing. Descriptions of the waste fractions sampled
are presented in the SI (Table S3).
Grab (active) sampling was used to obtain total leachate
concentrations, and POM passive sampling was used to obtain
freely dissolved leachate concentrations. The grab samples were
obtained by submerging a pre-sterilized 1 L green-tinted glass
bottle in the leachate water (either an open stream or inside a
culvert or manhole) on the first day of the relevant field
campaigns. The bottles were wrapped in aluminum foil and
transported cool (4 °C) to the laboratory. The same day, 2 g of
sodium azide (Sigma-Aldrich, USA) was added to the water
samples to prevent microbial degradation of BPA; they were
also spiked with BPA-d6 (used as a recovery standard, TRC,
Canada) to check the degradation/extraction recovery and
frozen until analysis.
Leachate passive samples were obtained by deploying POM76 samplers housed in stainless steel frame into the leachate
water for the entirety of the sampling campaigns (ca. 2−3

overview of monitoring levels of BPA is presented in the
Supporting Information (SI) and Table S1.
The focus of BPA emission and exposure research has been
on food and product packaging, with packaging materials made
from polycarbonate plastics and epoxy−resin-lined containers
identified as substantial sources of exposure, along with thermal
paper.29−33 When these materials are disposed of, they enter
the waste stream to form bulk waste fractions that are rich in
BPA, such as combustible waste (plastic sub-fractions)20 and
incineration residues.18,23 How these different types of waste
fractions contribute to BPA leachate concentrations at landfills
and other waste facilities remains unclear. Compared to food
packaging, comparatively limited research has been carried out
to investigate the mechanisms from which BPA can be released
from bulk waste fractions. Further, it remains unknown how
environmental concentrations around landfilling facilities
compare to other kinds of waste-handling facilities, such as
fragmenting, sorting, incineration, and recycling facilities. A
Japanese survey reported that BPA concentrations in leachate
appeared independent of waste composition at landfill sites.22
To gain new insight into the sources and mechanisms
regulating BPA concentrations at waste-handling facilities, we
conducted a comprehensive field and laboratory campaign
comprising 12 different facilities and eight types of waste
categories to study their presence and partitioning behavior. A
key novel aspect of the presented investigation is the
development and utilization of a passive sampling method to
specifically target the freely dissolved concentrations in water,
which allows for measuring the waste-water partitioning
behavior of these eight waste categories, as well as a comparison
of the total and freely dissolved concentration in landfill
leachate. Freely dissolved concentrations are more appropriate
to consider when describing partitioning behavior of contaminants, as they more closely regulate environmental fate and
bioavailability.34,35 The water-phase passive sampling material
used was polyoxymethylene (POM), which is slightly polar and
therefore appropriate for BPA.36,37 The hypotheses we set out
to test in this study were the following: (1) substantial amounts
of BPA in landfill leachate originate from plastic-containing
waste fractions; (2) BPA leachate concentrations are primarily
freely dissolved (and not bound to plastic particulates or
colloids); and (3) BPA concentrations in air and water from
waste-sorting and incineration facilities are lower than from

Bisphenol A. BPA is moderately hydrophobic (log Kow =
3.4), weakly acidic (pKa = 9.8 and 11.3), and soluble in water
(solubility of 300 mg/L), with recommended environmental
half-lives of 4 h in air, 4.5 days in water, and 300 days in soil.38
Calibration of Polyoxymethylene Samplers. A novel
method using POM passive samplers36 for quantifying the
freely dissolved fraction of BPA in leachate water and for
determining waste-water partition coefficients was developed in
this study. For this method, a reliable understanding of BPA
uptake kinetics into POM, as well as the POM−water partition
coefficient, KPOM, is needed:
KPOM(L/kg) = C POM /Cwater


where CPOM is the equilibrium concentration in the POM phase
(μg/kg) and Cwater is the equilibrium freely dissolved
concentration in the water phase (μg/L). To quantify KPOM
over a concentration range for landfill leachate, BPA solutions

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extraction recovery. The samples were extracted using a solid
phase extraction method with Strata-SDB L cartridges (500 mg,
6 mL, 100 μm, 260 Å, USA). The cartridges were first
conditioned with 5 mL of ethyl acetate and 5 mL of preextracted distilled water. Following this, between 1 and 50 mL
samples of leachate water were loaded onto the cartridges. The
cartridges were then rinsed with 10 mL of pre-extracted
distilled water and dried using a vacuum pump. BPA was eluted
from the cartridges using 10 mL of ethyl acetate. This extract
was evaporated to dryness and treated as above before GC-MS
POM passive samplers were placed in an extraction vial and
spiked with the BPA-d16 RS. They were then cold extracted for
7 days with ethyl acetate (20 mL) by shaking end-over-end at
13 rpm and then handled as the solid waste samples.
Quality Assurance and Control. All the solvents used
were GC grade or Emsure quality. All stock solutions were
prepared from pure BPA powder dissolved in GC-grade ethyl
acetate and were stored at −20 °C. All glassware was rinsed
with acetone, washed in a laboratory dishwasher, and pyrolyzed
at 450 °C (except the volumetric glassware used for standards,
which did not undergo the pyrolysation step) prior to use.
Before storage, water samples were spiked with sodium azide to
prevent microbial degradation and with BPA-d6 to check the
degradation/extraction recovery. In addition all samples were
also spiked with BPA-d16 prior to extraction. All raw data results
were corrected on the basis of the recovery percentages of all
these standards.
During a GC-MS sequence, the calibration standards (1−100
μg/L) were injected at least twice to take into account the
possibility of signal drift with time. The quantification was done
with the internal calibration technique (internal standard PCB77). All the samples were quantified within the range of the
calibration standard (1−100 μg/L).
Solvent blank samples were analyzed regularly to check for
potential contamination from the GC-MS. All analytes were
quantified using a quantification and a confirmation transition
from the MS. The quantification transition was chosen as the
most intense peak, and the ratio of the confirmation/
quantification transition was used when it was in the same
range (±20%) as that of the calibration standards.
GC-MS Analysis. The concentrations of BPA in the
derivatized sample extracts and standards were quantified
using a gas chromatograph 6850 coupled to a mass
spectrometer 5973 (Agilent, USA). The chromatographic
column was a SLB-5ms fused silica capillary column 30 m ×
0.25 mm × 0.25 μm (Supelco, USA). A five-point calibration
was made at 1, 5, 10, 50, and 100 μg/L BPA concentration
standards, in which BPA-d16, BPA-d6, and PCB 77 were added.
Further GC-MS method details are presented in the SI (section
Waste-water Partition Coefficients. A batch-shake
method was used to obtain waste-water partition coefficients,

months, for logistic reasons and to ensure equilibrium
partitioning), then wrapped in aluminum foil, placed in glass
jars, and transported at 4 °C back to the laboratory where they
were stored at −20 °C until CPOM analysis. Further details of
both grab and passive sampling of leachate are provided in the
SI (section S2).
At selected waste-handling facilities, ambient air particulate
matter (PM) samples were obtained at the site of most activity,
either next to a shredder, waste sorter, loading dock or in a
central location, using a high-volume (HighVol) air sampler
(Digitel, Switzerland), which was equipped with a PM10 cutoff,
150 mm diameter glass fiber filter (GFF, Sigma-Aldrich, USA),
to quantify the air-particle-associated BPA concentration. The
HighVol was deployed for 1−5 days. Due to intense particle
loadings in some areas, the cutoff may have been compromised,
and particles larger than PM10 may have entered the HighVol.
Air passive samplers consisting of XAD-2 resin beads
contained in a stainless tube, as designed by Wania et al.,40
were deployed in central locations, but not directly next to
particle shredders or areas were dust was visible. These
samplers were deployed for the entirety of a given sampling
campaign (2−3 months). Both GFF and XAD-2 tubes were
wrapped in aluminum foil after deployment and transported at
4 °C back to the laboratory where they were stored at −20 °C
until analysis. Further details are provided in the SI (section
Sample Preparation. In the laboratory, solid waste
samples were further homogenized in the polyurethane bags
by shaking or manual mixing, before 20−400 g was randomly
sampled from within the bag for grinding. All samples were
ground until they could pass through a 2 or 4 mm sieve
(depending on the material, as indicated in Table S2). Crushing
was carried out using either a BB100 Retsch jaw crusher (VWR,
Norway) (typically for glass and coarse ashes), a kitchen handblender (Braun or Phillips), a hand-powered malt mill
(Bryggeland, Norway) (typically for fluff and plastic), a mortar
and pestle, or simply by sieving through the appropriate mesh.
Hard plastics and metal materials (>4 mm) were the most
difficult to crush and sieve, and thus for the four samples of this
consistency (two WEEE samples from the site “WEEE/Vehicle
B”, one vehicle fluff and one vehicle plastic sample from the site
“WEEE/Vehicle E”), the original mass fraction of these
materials may be slightly misrepresented in the mixed, crushed
sample that was used for analysis.
Quantification of BPA. Solid waste samples, including the
GFF filters and the XAD-2 resins, were extracted using a
Soxhlet method (Behr Labor-Technik, Germany) with 100 mL
of ethyl acetate (GC-MS grade, 99.8% purity, Merck, KGaA,
Germany) for 12 h at 105 °C. The samples were spiked with a
recovery standard (RS; BPA-d16, 99.9% purity, Supelco, USA)
prior to extraction to check extraction recovery. Following
extraction, between 10 μL and 10 mL of the solvent was
evaporated to dryness using a vacuum centrifuge (Vacuubrand
2C, Vakuum Service AS, Germany). The residue was then
dissolved in 950 μL of ethyl acetate that contained an internal
standard (IS) at 50 μg/L to check for matrix effects (PCB-77,
99.97% purity, Fluka, Switzerland) and 50 μL of a derivatization
reagent (MTBSTFA, at 60 °C for 30 min, > 97% purity, SigmaAldrich, USA) in order to allow for GC-MS quantification. All
solid concentration data are presented on a dry weight basis
Frozen grab leachate water samples were thawed in the dark
and then spiked with BPA-d16 to act as a RS to check for the

KD,waste(L/kg) = Cwaste/Cwater


where Cwaste is the concentration in the waste at equilibrium
(μg/kg). KD,waste was determined here using an approach that
was adapted from a standard method for metals in waste
materials (EN 12457), by using POM to quantify Cwater and by
increasing the duration of shake time from 1 day to 28 days to
ensure equilibrium. Between 0.5 and 2 g of ground solid waste
material (≤4 mm), along with 0.1 g of pre-cleaned POM, were

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inferred that the deployment time used here (2−3 months)
would be more than sufficient to reach equilibrium if BPA
leachate concentrations were time independent. In areas of
fluctuating concentrations, the polar organic chemical integrative sampler (POCIS), which has been used to measure the
concentration of BPA in spiked tap water, surface water, wastewater, and estuary water41−45 would be more appropriate as it
does not need to reach equilibrium in order to determine water
concentrations. However, in areas of stable concentration,
POM is advantageous compared to POCIS, as extrapolation of
Cwater is not dependent on (multi-phase) uptake rates, and
sample handling time is reduced.46
Solid and Leachable Waste Concentrations. The total
and leachable BPA concentrations quantified for the different
solid waste categories and sub-fractions are shown in Table S4,
with waste category results plotted in Figure 2. Note that the
waste category results presented in Figure 2 are weighted
averages based on the annual mass produced in Norway of the
various waste sub-fractions belonging to a waste category (as
presented in Tables S3 and S4). As an example, the waste
category WEEE contains the sub-fractions “BFR plastic”, “cable
plastic”, “other plastics”, and “metals”, which are reported to be
generated at 2, 15, 48, and 80 megatons/year in Norway,
respectively (Table S4).38 These WEEE sub-fractions were
measured in our study to have BPA concentrations of 84400,
29100, 200500, and 1170 μg/kg, respectively. Thus, the
weighted average of BPA in WEEE was 71100 μg/kg =
{(2×84400 μg/kgBFR plastic + 15×29100 μg/kgcable plastic +
48×200500 μg/kgother plastics + 80×1170 μg/kgmetals)/(2 + 15
+ 48 + 80)}.
From Figure 2, the waste category with the largest
concentrations of BPA was plastics (weighted average 188000
± 125000 μg/kg), followed by the plastic-rich waste fractions
WEEE (71200 ± 46700 μg/kg) and vehicle fluff (6490 ± 3350
μg/kg). The lowest concentrations were found in fly ash
(μg/kg); note that fly ash and bottom ash refer to ashes
collected in the chimney filters and ovens within incinerators,
respectively. The BPA concentration in the sampled combustibles (1248 ± 349 μg/kg) was larger than in ash, indicating
that BPA is substantially though not completely destroyed
through incineration, as observed in controlled studies.28
Incineration reduces waste mass by approximately a factor of
3;47 thus if BPA was stable, its concentration should have
increased by a factor of 3 and not decreased as observed.
Digestate (i.e., sewage sludge that was digested to make
methane) contained similar quantities of BPA as combustibles.
As Figure 2 indicates, the total concentrations quantified in
the vehicle fluff, WEEE, and plastic wastes are higher than the
proposed soil PNEC of 3700 μg/kg.4 The presence of these
wastes in landfill soils could therefore pose a risk to soil
dwelling organisms.
The leachable concentrations organized from greatest to
smallest in Figure 2 follows the same order as the total
concentrations (i.e., plastics > WEEE > vehicles > combustibles
≈ digestate > bottom ash > glass > fly ash). Leachable
concentrations at L/S 10 ranged from < LOQ for fly ashes to
1970 μg/kg for plastics (Figure 2 and Table S4), roughly
corresponding to 1% of the total BPA leaching into the water
phase for most waste samples (0.6−1.6%), except for glass
samples (30.6%) and fly ash (The higher leaching percentage of BPA from the glass samples
can be attributed to concentrations close to the limit of

shaken for 28 days at room temperature with pre-extracted
distilled water at a liquid-to-solid weight ratio (L/S) of 10.
Afterward the POM strips were removed, CPOM was quantified,
and Cwater was determined on the basis of KPOM. In addition, the
concentration of BPA leached from the waste fraction, Cleachable
(μg/kg d.w.), was calculated using the relationship Cleachable =

BPA POM−Water Partitioning. The POM−water sorption isotherms for the 76 and 55 μm materials were similar and
linear in the concentration range from 1 to 1000 μg/L (R2 =
0.99, Figure 1A). The log KPOM values of the 76 and 55 μm

Figure 1. (A) POM sorption isotherm over the spiked concentration
range of 1−1000 μg/L (n = 9; RSD = 5%; error bars shown but are
smaller than the markers) for the 76 and 55 μm thick materials at 28
days of shaking. (B) Changes in log CPOM/Cw as a function of shaking
time (in days), along with a fitted first-order kinetic model.

materials were 2.45 ± 0.12 and 2.58 ± 0.11, respectively, which
agree well with a single-point measurement of the 76 μm
material by Endo et al.37 of 2.63 (at 33 μg/L). The kinetic
uptake (n = 2, 76 μm) could be described as first-order (R2 =
0.95), with >80% sorption equilibrium being achieved within 7
days of shaking (Figure 1). A follow-up experiment with more
sampling events in the first 7 days would be recommended to
better characterize the uptake kinetics.
For determining waste-water partitioning coefficients, the 28
day shaking test is therefore conservative regarding POM
uptake, though it is still recommended to account for potential
slow desorption kinetics from waste fractions. This kinetic
system is not representative of that encountered when fielddeploying POM into flowing leachate water, but it can be

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Figure 2. Waste category total BPA concentrations, Cwaste (left n-value), and leachable BPA concentrations, Cleachable (right n-value) at a liquid/solid
(L/S) ratio of 10 of the types of solid waste fractions considered in this study. Data presented are weighted averages based on the annual mass
produced in Norway of the various waste sub-fractions belonging to a waste category. The waste categories are organized from the smallest to largest
BPA concentrations. Also presented is the predicted no-effect concentration (PNEC) of BPA in soil of 3700 μg/kg d.w.2

latter observation could be due to BPA typically not being
associated within these materials, though trace levels could be
on their surface as coatings or labels (similar to our results for
glass). Regarding incinerator ashes, a Japanese survey of
leachate from ash-landfills concluded that bottom ash was a
minor source of BPA in leachate, but “solidified fly ash” and
incombustibles were major sources of BPA.23 Our study agrees
with the former but not with the latter conclusion; we speculate
that this disagreement may be related to the solidification
process of the fly ash in this Japanese survey.23
Waste-water Partitioning. The solid waste-water partitioning coefficients (KD,waste) for all waste categories and subfractions are shown in Table S4. All waste sub-fractions have
similar log KD,waste values (ranging from 2.1 ± 0.6 to 3.1 ± 1.0)
except for glass waste (log KD = 1.5 ± 0.3), where BPA
concentrations were close to the LOQ, and WEEE metals
(log KD = 1.8 ± 0.4). The mass fraction of total organic carbon
in the waste, f TOC, was used to normalize KD,waste, as this
parameter is generally correlated with KD values of organic
compounds, according to

quantification, and also to BPA residue being mostly on the
glass surface (e.g., from epoxy−resin coatings or labels).
Additionally, the average pH for the leachable BPA samples
for the glass (pH 9.9) were slightly higher than the pKa1 (pKa1
= 9.8) for BPA, which means that about 50% of BPA was
present in its single negatively charged form. For bottom ashes,
the BPA was present at about 50% in its single- and 50% in its
double-negatively charged form (pH 10.8 and pKa2 = 11.3). For
fly ashes, the pH was 12.2, which is higher than pKa2, meaning
that if any BPA was present it would be in its double-negatively
charged form.
A recent survey of contaminants in Norwegian vehicle fluff48
measured an average BPA concentration of 5000 μg/kg (±5000
μg/kg; n = 10), which is comparable to the result determined
here (6492 ± 3350 μg/kg; n = 12). Another Norwegian survey
reported median BPA concentrations of 536 ± 446 μg/kg (n =
32) in sludge samples from 8 different water treatment plants10
(SI, section S1), which is consistent with the digestate samples
(888 ± 401 μg/kg; n = 8). Several studies have quantified
concentrations of BPA in plastic waste and products.
Yamamoto and Yasuhara49 measured the total BPA concentration in 17 plastic waste samples. Ten of their samples had
concentrations below the limit of detection, while the other
seven had concentrations between 71000 and 1280000 μg/kg
(average 605000 μg/kg). Xu et al.50 reported total BPA
concentrations between 1600 and 12100 μg/kg for five
different plastic wastes, and Biles et al.51 quantified BPA
concentrations in PC bottles to be between 7000 and 58000
μg/kg. Thus, concentrations of BPA in plastics are highly
variable and quite dependent on the type of material (this is
further discussed in the SI, section S1); however, the results
measured here are well within the range reported in these other
Regarding leachable concentrations, a previous Japanese
study reported that less than 3% of the total BPA was leached
from nine plastic samples,49 in agreement with our average of
1%. Xu et al.50 reported a similar level of BPA leaching from
polycarbonate and polyethylene (<1.5%) but reported much
higher (20−55%) BPA leaching from high-density polyethylene, polyvinyl chloride, and polystyrene at pH 6.2. This

KTOC,waste(L/kg TOC) = KD,waste/fTOC


For non-ash wastes, log KTOC,waste (presented in Table S4)
were between 2.5 ± 0.3 (cable plastic) and 3.8 ± 0.5 (fine
vehicle fluff). These results are in agreement with calculated
values from the literature for soils (log KTOC = 2.5) or
sediments (log KTOC = 3.2),52 though are lower than the
log KTOC reported by Heemken et al.11 for suspended particle
matter in surface water (log KTOC = 4.5) and the log KDOC
quantified by Kalmikova et al.24 in landfill leachates (log KDOC
= 4.5). For bottom ash, the log KTOC was measured at 4.9 ±
0.5, likely due the presence of strong sorbing black carbon
phases.53 A correlation between the log KTOC,waste and the
percent TOC in the solid waste samples is shown in Figure S2
(R2 = 0.75), from which it is evident that the smaller the TOC,
the higher the sorption strength of that TOC, with plastic-rich
waste having the lowest values and digestate and bottom ash
the highest. This can be related to the expectation that surface
area, porosity, and accessibility of sorption sites in ash and
digestate are larger than those of synthetic plastic phases.

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Figure 3. Comparison of BPA concentrations measured in leachate water, Cleachate, with grab sampling (left n-value) and POM passive sampling
(right n-value). Results are presented according to waste-handling facility type, and standard deviations are calculated from the average of data
collected from all sampling campaigns.

Figure 4. (A) Air-particle (PM10) associated bisphenol A concentrations (pg/m3) near hot spots of various waste facilities (near shredders, sorting
activities or in the center of the landfill). Note that concentrations for the Incinerator/Sorting A,B and WEEE/Vehicle C facilities are indoors, and
for Landfills A−C and WEEE/Vehicle A facilities they are outdoors. (B) The concentration of BPA in PM10 dust, CPM10 (μg/kg), produced from
specific types of waste as they are being processed at a facility (left n-value, referring to the number of PM10 samples) in relation to the
corresponding concentrations in the waste itself, Cwaste (μg/kg) (right n-value, referring to the number of waste samples).


DOI: 10.1021/acs.est.5b01307
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Environmental Science & Technology
Waste Facility Leachate Water Concentrations. Figure
3 and Table S5 show the average BPA concentrations
quantified in the leachate water, Cleachate, using grab and passive
sampling methods, which comprise sampling time intervals of
2−3 min and 2−3 months, respectively. Figure 3 shows that
BPA Cleachate using both methods are similar (p < 0.05). This
provides some indication that BPA concentrations were stable
over time. Corresponding to this, no substantial seasonal
variation was observed in the BPA concentration from the
different sampling campaigns (see Table S6). Only minor
deviations in seasonal trends of BPA concentrations were also
seen in Japanese studies,21,22 though a slight decrease on the
scale of years was observed in one of these studies.20
As shown in Figure 3, the BPA leachate water concentrations
quantified in this study ranged from approximately 0.7 to 200.0
μg/L for the landfills and from 5.0 to 100.0 μg/L for the
WEEE/vehicle facilities, which is comparable to levels for
landfills measured in other countries as mentioned in the
Introduction (e.g., average 66.5 μg/L in Norwegian landfills,
0.1−17200 μg/L in Japanese landfills, and 0.01−107 μg/L in
Swedish landfills). With the exception of the bottom-ash-rich
Landfill A, all Cleachate values exceed the chronic PNEC, with
Landfill C and WEEE/Vehicle C exceeding the acute PNEC by
more than an order of magnitude. This study is the first to
show that leachate concentrations from WEEE/vehicle facilities
can be similar to landfills (p < 0.05). The Cleachate for the
bottom-ash-rich Landfill A being lower than for other landfills
agreed with a Japanese study which reported ash-rich landfills
had untreated Cleachate levels of <0.05−34 μg/L, which was less
than for landfills containing mixtures of ash and other waste,
2.6−4960 μg/L.16 A different Japanese study of incinerator ash
landfills showed generally smaller concentrations of BPA
(median 1.7 μg/L)21 than other solid waste landfills (with
medians of 350,17 269,18 and 70.9 and 91.420 μg/L).
Grab samples quantify both the freely dissolved and
dissolved organic carbon (DOC)-bound BPA, whereas passive
sampling quantifies exclusively the freely dissolved compound.
The good agreement in Figure 3 between grab and passive
measurements suggests leachate water BPA is predominately in
the freely dissolved phase. The DOC concentration, CDOC, in
the tested leachate water was measured between 5 and 23 mg/
L, except for Landfill C (365 mg/L) (Table S5). Assuming the
DOC has the same KTOC values as waste of approximately 1000
L/kg, multiplying KD × CDOC would imply only 0.5−2.3% of
BPA would be DOC bound, except Landfill C at 36%, therefore
implying the majority of BPA should be freely dissolved. Thus,
there is good agreement between the comparison between grab
and passive sampling results with model expectations based on
the reported KTOC,waste range of BPA. More discussion on the
role of DOC in leachate can be found in the SI (Figure S3).
Air and Dust Concentrations. The concentrations of
PM10-bound BPA in the air, Cair,PM10 (pg/m3), quantified near
waste shredders or the site of most activity, are presented in
Figure 4A and Table S7. BPA was not detected in the XAD-2
resin, thus the presence of volatile BPA could not be quantified
(nor could we provide a limit of detection for this method).
However, due to the low vapor pressure of BPA (5.4 × 10−6
Pa), it is generally assumed to be predominately particle
For Landfills A−C and WEEE/Vehicle A, the Cair,PM10 was
near 1000 pg/m3. At these facilities PM10 was collected
outdoors, near the center of the landfills or next to an outdoor
metal shredder (WEEE/Vehicle A). At the other facilities,

samples were collected indoors, either near the waste
defragmentation and sorting area (WEEE/Vehicle C, Incineration/Sorting B), or near the waste loading dock (Incineration/
Sorting A), and Cair,PM10 were ca. 10 times higher, at around
10000 pg/m3, compared to the indoor samples, with less air
circulation and exposure to sunlight.
Previous studies that have quantified air BPA concentrations
are scarce. Matsumoto et al.54 measured a BPA concentration
of 514 pg/m3 in air-particulate matter following a 6 month
exposure to outdoor air at Osaka, Japan. Much higher
concentrations were reported for an urban outdoor area in
India where peak concentrations were 17400 pg/m3 (lowest
concentration 200 pg/m3, average 4550 pg/m3).55 Air
associated BPA concentrations of 1110 pg/m3 were recorded
inside of an WEEE workshop in China. 56 Thus, the
concentrations at these Norwegian waste facilities seem similar
to reported urban and WEEE workshop environments in Asia.
A comparison of BPA concentration in the PM10 dust itself,
CPM10 (μg/kg) produced from the sorting/shredding processes,
along with the concentration in the corresponding solid waste,
Cwaste (μg/kg), is presented in Figure 4B and Table S7. These
values, which range from 2343 μg/kg (bottom ash sorting) to
50651 μg/kg (WEEE fragmenting), are elevated compared to
CPM10 levels reported in a survey of American households, with
a median of 821 μg/kg and maximum of 17600 μg/kg. The
ratio (CPM10/Cwaste) was calculated for each facility, and results
are given in Table S7. The ratios for WEEE and vehicle facilities
were near 1, indicating the dust is representative of BPA
concentration in the wastes being shredded. For ash and
combustible sorting, ratios were >1, indicating that finer, airborne waste fractions contain higher concentrations of BPA
than the total waste fraction (perhaps from air-suspended paper
fibers or other BPA sources present at these facilities).
Environmental Implications. The results of this study
support the first hypothesis stated at the end of the
Introduction, that substantial amounts of BPA in landfill
leachate originate from plastic-containing waste fractions
(Figure 2). Note, however, that other sources of BPA could
exist in landfill leachate, such as thermal-paper coatings. The
study also supports the second hypothesis, that BPA leachate
concentrations are primarily freely dissolved and not bound to
(plastic) colloids (p < 0.05). However, the study did not
support the third hypothesis, that BPA concentrations from
waste-sorting and incineration facilities are lower than those
from waste landfills. Leachate concentrations from WEEE/
vehicle facilities were similar in range to those from landfills
(Figure 3). Both WEEE/vehicle and incineration/sorting
facilities can exhibit higher atmospheric BPA concentrations
than landfills, though this is partially explained by lower air
circulation and exposure to sunlight in some of the WEEE/
vehicle and incineration/sorting facilities, which were enclosed
to varying extents.
Incineration lowers the total amount of BPA in waste, based
on the relatively low concentrations of BPA in ash compared to
the source waste reported here and in controlled experiments
elsewhere.28 Further, incineration provides a waste residue with
a comparatively high log KTOC sorption coefficient for BPA
(and potentially other organic contaminants), indicating
reduced concentrations and emissions from bottom ash
landfills. Though recycling is generally a favored option over
incineration for waste handling (as specified in the European
Union’s Waste Framework Directive (1975/442/EEC)), careful
selection of highly BPA-contaminated waste for incineration in

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Environmental Science & Technology

(10) Bolz, U.; Hagenmaier, H.; Korner, W. Phenolic xenoestrogens in
surface water, sediments, and sewage sludge from Baden-Wurttemberg, south-west Germany. Environ. Pollut. 2001, 115 (2), 291−301.
(11) Heemken, O. P.; Reincke, H.; Stachel, B.; Theobald, N. The
occurrence of xenoestrogens in the Elbe river and the North Sea.
Chemosphere 2001, 45 (3), 245−259.
(12) Arp, H. P. H., Compilation of Norwegian Screening Data for
Selected Contaminants (2002−2012), Miljødirektorat rapport TA
2982, 2012; 610 pp (http://www.miljodirektoratet.no/old/klif/
(13) Chen, T.-C.; Shue, M.-F.; Yeh, Y.-L.; Kao, T.-J. Bisphenol A
occurred in Kao-Pin River and its tributaries in Taiwan. Environ. Monit.
Assess. 2010, 161 (1−4), 135−145.
(14) Funakoshi, G.; Kasuya, S. Influence of an estuary dam on the
dynamics of bisphenol A and alkylphenols. Chemosphere 2009, 75 (4),
(15) Hashimoto, S.; Horiuchi, A.; Yoshimoto, T.; Nakao, M.; Omura,
H.; Kato, Y.; Tanaka, H.; Kannan, K.; Giesy, J. P. Horizontal and
vertical distribution of estrogenic activities in sediments and waters
from Tokyo Bay, Japan. Arch. Environ. Contam. Toxicol. 2005, 48 (2),
(16) Boyd, G. R.; Reemtsma, H.; Grimm, D. A.; Mitra, S.
Pharmaceuticals and personal care products (PPCPs) in surface and
treated waters of Louisiana, USA and Ontario, Canada. Sci. Total
Environ. 2003, 311 (1−3), 135−149.
(17) Klecka, G. M.; Staples, C. A.; Clark, K. E.; van der Hoeven, N.;
Thomas, D. E.; Hentges, S. G. Exposure Analysis of Bisphenol A in
Surface Water Systems in North America and Europe. Environ. Sci.
Technol. 2009, 43 (16), 6145−6150.
(18) Sakamoto, H.; Fukui, H.; Souta, I.; Kaneko, K. Studies on
Bisphenol A and its Origins in Leachates from Solid Waste Landfills.
Jpn. Soc. Waste Manage. 2004, 15, 511−520.
(19) Yasuhara, A.; Shiraishi, H.; Nishikawa, M.; Yamamoto, T.;
Uehiro, T.; Nakasugi, O.; Okumura, T.; Kenmotsu, K.; Fukui, H.;
Nagase, M.; Ono, Y.; Kawagoshi, Y.; Baba, K.; Noma, Y.
Determination of organic components in leachates from hazardous
waste disposal sites in Japan by gas chromatography mass
spectrometry. J. Chromatogr. A 1997, 774 (1−2), 321−332.
(20) Yamamoto, T.; Yasuhara, A.; Shiraishi, H.; Nakasugi, O.
Bisphenol A in hazardous waste landfill leachates. Chemosphere 2001,
42 (4), 415−418.
(21) Taro, U.; Kenichiro, M. Factors affecting the concentration of
bisphenol A in leachates from solid waste disposal sites and its fate in
treatment processes. J. Mater. Cycles Waste Manage. 2003, 5, 77−82.
(22) Asakura, H.; Matsuto, T.; Tanaka, N. Behavior of endocrinedisrupting chemicals in leachate from MSW landfill sites in Japan.
Waste Manage. (Oxford) 2004, 24 (6), 613−622.
(23) Kurata, Y.; Ono, Y.; Ono, Y. Occurrence of phenols in leachates
from municipal solid waste landfill sites in Japan. J. Mater. Cycles Waste
Manage. 2008, 10 (2), 144−152.
(24) Kalmykova, Y.; Bjorklund, K.; Stromvall, A.-M.; Blom, L.
Partitioning of polycyclic aromatic hydrocarbons, alkylphenols,
bisphenol A and phthalates in landfill leachates and stormwater.
Water Res. 2013, 47 (3), 1317−1328.
(25) Schwarzbauer, J.; Heim, S.; Brinker, S.; Littke, R. Occurrence
and alteration of organic contaminants in seepage and leakage water
from a waste deposit landfill. Water Res. 2002, 36 (9), 2275−2287.
(26) Limam, I.; Mezni, M.; Guenne, A.; Madigou, C.; Driss, M. R.;
Bouchez, T.; Mazéas, L. Evaluation of biodegradability of phenol and
bisphenol A during mesophilic and thermophilic municipal solid waste
anaerobic digestion using 13C-labeled contaminants. Chemosphere
2013, 90 (2), 512−520.
(27) He, P.-J.; Zheng, Z.; Zhang, H.; Shao, L.-M.; Tang, Q.-Y. PAEs
and BPA removal in landfill leachate with Fenton process and its
relationship with leachate DOM composition. Sci. Total Environ. 2009,
407 (17), 4928−4933.
(28) Šala, M.; Kitahara, Y.; Takahashi, S.; Fujii, T. Effect of
atmosphere and catalyst on reducing bisphenol A (BPA) emission

state-of-the-art, low-emission facilities rather than for recycling
would be a way to reduce the amount of BPA in recycled


S Supporting Information

More background information on BPA, extended materials and
methods including the field campaign, and raw data. The
Supporting Information is available free of charge on the ACS
Publications website at DOI: 10.1021/acs.est.5b01307.


Corresponding Author

*Phone: +47 950 20 667; e-mail: hpa@ngi.no.

The authors declare no competing financial interest.

Funding for this research was provided by the Research Council
of Norway (WASTEFFECT, grant number 221440/E40,
http://www.ngi.no/no/Prosjektnett/WASTEFFECT/). Field
sampling, planning, and logistics were greatly aided by Geir
Allum Sørensen (NG, Mepex), Magnus Sparrevik (NGI,
Forsvarsbygg), Gudny Okkenhaug (NGI), Frank Wania
(University of Toronto), and Knut Breivik (NIVA) along
with site owners. Discussions and planning with the WASTEFFECT steering committee (G. A. Sørensen, Rita Vigdis
Hansen (Miljødirektoratet), Line Diana Blytt (Aquateam,
Avfallsforsk), and Laila Borgen Skaiaa (Renas, Avfallsforsk))
helped guide the direction of this research.


(1) Huang, Y. Q.; Wong, C. K. C.; Zheng, J. S.; Bouwman, H.; Barra,
R.; Wahlström, B.; Neretin, L.; Wong, M. H. Bisphenol A (BPA) in
China: A review of sources, environmental levels, and potential human
health impacts. Environ. Int. 2012, 42 (0), 91−99.
(2) Pivnenko, K.; Eriksson, E.; Astrup, T. F. Waste paper for
recycling: Overview and identification of potentially critical substances.
Waste Manage. (Oxford) 2015, DOI: 10.1016/j.wasman.2015.02.028.
(3) Merchant Research & Consulting Ltd. Website: http://mcgroup.
co.uk/news/20131108/bpa-production-grew-372000-tonnes.html (accessed Feb 9, 2015).
(4) European Commission, Updated European Risk Assessment
Report, 4,4′-Isopropylidenediphenol (Bisphenol-a), 2008.
(5) Bakke, T.; Kallqvist, T.; Ruus, A.; Breedveld, G. D.; Hylland, K.
Development of sediment quality criteria in Norway. J. Soils Sediments
2010, 10 (2), 172−178.
(6) Flint, S.; Markle, T.; Thompson, S.; Wallace, E. Bisphenol A
exposure, effects, and policy: A wildlife perspective. J. Environ. Manage.
2012, 104, 19−34.
(7) Belfroid, A.; van Velzen, M.; van der Horst, B.; Vethaak, D.
Occurrence of bisphenol A in surface water and uptake in fish:
evaluation of field measurements. Chemosphere 2002, 49 (1), 97−103.
(8) Vethaak, A. D.; Lahr, J.; Schrap, S. M.; Belfroid, A. C.; Rijs, G. B.
J.; Gerritsen, A.; de Boer, J.; Bulder, A. S.; Grinwis, G. C. M.; Kuiper,
R. V.; Legler, J.; Murk, T. A. J.; Peijnenburg, W.; Verhaar, H. J. M.; de
Voogt, P. An integrated assessment of estrogenic contamination and
biological effects in the aquatic environment of The Netherlands.
Chemosphere 2005, 59 (4), 511−524.
(9) Jin, X. L.; Jiang, G. B.; Huang, G. L.; Liu, J. F.; Zhou, Q. F.
Determination of 4-tert-octylphenol, 4-nonylphenol and bisphenol A
in surface waters from the Haihe River in Tianjin by gas
chromatography-mass spectrometry with selected ion monitoring.
Chemosphere 2004, 56 (11), 1113−1119.

DOI: 10.1021/acs.est.5b01307
Environ. Sci. Technol. XXXX, XXX, XXX−XXX


Environmental Science & Technology
during thermal degradation of polycarbonate. Chemosphere 2010, 78
(1), 42−45.
(29) Sajiki, J.; Yonekubo, J. Leaching of bisphenol A (BPA) to
seawater from polycarbonate plastic and its degradation by reactive
oxygen species. Chemosphere 2003, 51 (1), 55−62.
(30) Sajiki, J.; Yonekubo, J. Leaching of bisphenol A (BPA) from
polycarbonate plastic to water containing amino acids and its
degradation by radical oxygen species. Chemosphere 2004, 55 (6),
(31) Sajiki, J.; Miyamoto, F.; Fukata, H.; Mori, C.; Yonekubo, J.;
Hayakawa, K. Bisphenol A (BPA) and its source in foods in Japanese
markets. Food Additives Contam. 2007, 24 (1), 103−112.
(32) Cooper, J. E.; Kendig, E. L.; Belcher, S. M. Assessment of
bisphenol A released from reusable plastic, aluminium and stainless
steel water bottles. Chemosphere 2011, 85 (6), 943−947.
(33) Fürhacker, M.; Scharf, S.; Weber, H. Bisphenol A: emissions
from point sources. Chemosphere 2000, 41 (5), 751−756.
(34) Reichenberg, F.; Mayer, P. Two complementary sides of
bioavailability: Accessibility and chemical activity of organic contaminants in sediments and soils. Environ. Toxicol. Chem. 2006, 25 (5),
(35) Mayer, P.; Tolls, J.; Hermens, J. L. M.; Mackay, D. Peer
Reviewed: Equilibrium Sampling Devices. Environ. Sci. Technol. 2003,
37 (9), 184A−191A.
(36) Jonker, M. T. O.; Koelmans, A. A. Polyoxymethylene solid
phase extraction as a partitioning method for hydrophobic organic
chemicals in sediment and soot. Environ. Sci. Technol. 2001, 35 (18),
(37) Endo, S.; Hale, S. E.; Goss, K.-U.; Arp, H. P. H. Equilibrium
Partition Coefficients of Diverse Polar and Nonpolar Organic
Compounds to Polyoxymethylene (POM) Passive Sampling Devices.
Environ. Sci. Technol. 2011, 45 (23), 10124−10132.
(38) Cousins, I. T.; Staples, C. A.; Klecka, G. M.; Mackay, D. A
multimedia assessment of the environmental fate of bisphenol A.
Human Ecol. Risk Assess. 2002, 8 (5), 1107−1135.
(39) Arp, H. P. H.; Hale, S. E.; Elmquist Kruså, M.; Cornelissen, G.;
Grabanski, C. B.; Miller, D. J.; Hawthorne, S. B. Review of
polyoxymethylene passive sampling methods for quantifying freely
dissolved porewater concentrations of hydrophobic organic contaminants. Environ. Toxicol. Chem. 2015, 34 (4), 710−720.
(40) Wania, F.; Shen, L.; Lei, Y. D.; Teixeira, C.; Muir, D. C. G.
Development and calibration of a resin-based passive sampling system
for monitoring persistent organic pollutants in the atmosphere.
Environ. Sci. Technol. 2003, 37 (7), 1352−1359.
(41) Arditsoglou, A.; Voutsa, D. Passive sampling of selected
endocrine disrupting compounds using polar organic chemical
integrative samplers. Environ. Pollut. 2008, 156 (2), 316−324.
(42) Zhang, Z.; Hibberd, A.; Zhou, J. L. Analysis of emerging
contaminants in sewage effluent and river water: Comparison between
spot and passive sampling. Anal. Chim. Acta 2008, 607 (1), 37−44.
(43) Li, H.; Helm, P. A.; Metcalfe, C. D. Sampling in the Great Lakes
for Pharmaceuticals, Personal Care Products, And EndocrineDisrupting Substances Using the Passive Polar Organic Chemical
Integrative Sampler. Environ. Toxicol. Chem. 2010, 29 (4), 751−762.
(44) Morin, N.; Camilleri, J.; Cren-Olive, C.; Coquery, M.; Miege, C.
Determination of uptake kinetics and sampling rates for 56 organic
micropollutants using “pharmaceutical” POCIS. Talanta 2013, 109,
(45) Bayen, S.; Segovia, E.; Loh, L. L.; Burger, D. F.; Eikaas, H. S.;
Kelly, B. C. Application of Polar Organic Chemical Integrative Sampler
(POCIS) to monitor emerging contaminants in tropical waters. Sci.
Total Environ. 2014, 482, 15−22.
(46) Fauvelle, V.; Mazzella, N.; Belles, A.; Moreira, A.; Allan, I. J.;
Budzinski, H. Optimization of the polar organic chemical integrative
sampler for the sampling of acidic and polar herbicides. Anal. Bioanal.
Chem. 2014, 406 (13), 3191−3199.
(47) Sabbas, T.; Polettini, A.; Pomi, R.; Astrup, T.; Hjelmar, O.;
Mostbauer, P.; Cappai, G.; Magel, G.; Salhofer, S.; Speiser, C.; HeussAssbichler, S.; Klein, R.; Lechner, P. Management of municipal solid

waste incineration residues. Waste Manage. (Oxford) 2003, 23 (1),
(48) COWI, Utrangerte kjøretøy og miljøgifter i materialstrømmer
ved fragmenteringsverk, COWI report 137155-01, 2013 (in
(49) Yamamoto, T.; Yasuhara, A. Quantities of bisphenol A leached
from plastic waste samples. Chemosphere 1999, 38 (11), 2569−2576.
(50) Xu, S.-Y.; Zhang, H.; He, P.-J.; Shao, L.-M. Leaching behaviour
of bisphenol A from municipal solid waste under landfill environment.
Environ. Technol. 2011, 32 (11), 1269−1277.
(51) Biles, J. E.; McNeal, T. P.; Begley, T. H.; Hollifield, H. C.
Determination of bisphenol-A in reusable polycarbonate food-contact
plastics and migration to food-simulating liquids. J. Agric. Food. Chem.
1997, 45 (9), 3541−3544.
(52) Staples, C. A.; Dorn, P. B.; Klecka, G. M.; O’Block, S. T.; Harris,
L. R. A review of the environmental fate, effects, and exposures of
bisphenol A. Chemosphere 1998, 36 (10), 2149−2173.
(53) Cornelissen, G.; Gustafsson, O.; Bucheli, T. D.; Jonker, M. T.
O.; Koelmans, A. A.; Van Noort, P. C. M. Extensive sorption of
organic compounds to black carbon, coal, and kerogen in sediments
and soils: Mechanisms and consequences for distribution, bioaccumulation, and biodegradation. Environ. Sci. Technol. 2005, 39 (18), 6881−
(54) Matsumoto, H.; Adachi, S.; Suzuki, Y. Bisphenol A in ambient
air particulates responsible for the proliferation of MCF-7 human
breast cancer cells and its concentration changes over 6 months. Arch.
Environ. Contam. Toxicol. 2005, 48 (4), 459−466.
(55) Fu, P.; Kawamura, K. Ubiquity of bisphenol A in the
atmosphere. Environ. Pollut. 2010, 158 (10), 3138−3143.
(56) Bi, X.; Simoneit, B. R. T.; Wang, Z.; Wang, X.; Sheng, G.; Fu, J.
The major components of particles emitted during recycling of waste
printed circuit boards in a typical e-waste workshop of South China.
Atmos. Environ. 2010, 44 (35), 4440−4445.


DOI: 10.1021/acs.est.5b01307
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