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Bioaccumulation of Per- and polyfluoroalkyl substances (PFASs) in a
tropical estuarine food web
Daniele A. Miranda
a,b,
⁎, Jonathan P. Benskin
b
,RaedAwad
b,c
, Gilles Lepoint
d
, Juliana Leonel
e
, Vanessa Hatje
a
a
Centro Interdisciplinar de Energia e Ambiente (CIEnAm) and Inst. de Química, Universidade Federal da Bahia, 41170-115 Salvador, BA, Brazil
b
Department of Environmental Science, Stockholm University, Stockholm, Sweden
c
Swedish Environmental Research Institute (IVL), Stockholm, Sweden
d
Freshwater and Oceanic sciences Unit of reSearch (FOCUS –Oceanology), University of Liege, 4000 Liege, Belgium
e
Departamento de Oceanografia, Universidade Federal de Santa Catarina, 88040-900 Florianópolis, SC, Brazil
HIGHLIGHTS
•Perfluoroalkyl substances (PFAS) were
investigated in a tropical estuarine
food web.
•Leaves, bivalves, polychaeta, crusta-
ceans, and fishpresented PFAS residues.
•PFOS, PFTrDA, and FOSA were the com-
pounds most frequently detected in
Subaé biota.
•The PFOS precursors FOSA and EtFOSA
were also detected in Subaé estuarine
biota.
•PFOS, PFNA and a C
8
PFOS precursor
were observed to biomagnify in the
food chain.
GRAPHICAL ABSTRACT
abstractarticle info
Article history:
Received 8 June 2020
Received in revised form 30 August 2020
Accepted 31 August 2020
Available online 2 September 2020
Editor: Daniel Wunderlin
Keywords:
POPs
PFOS precursors
Tropical food web
Biomagnification
Todos os Santos Bay
The biomagnification of per- and polyfluoroalkyl substances (PFASs) was investigated in a tropical mangrove
food web froman estuary in Bahia, Brazil. Samplesof 44 organisms (21 taxa),along with biofilm,leaves, sediment
and suspended particulate matter were analyzed. Sum (∑) PFAS concentrations in biota samples were domi-
nated by perfluorooctane sulfonate (PFOS, 93% detection frequency in tissues; 0.05 to 1.97 ng g
−1
ww whole-
body (wb)), followed by perfluorotridecanoate (PFTrDA, 57%; 0.01 to 0.28 ng g
−1
ww wb). PFOS precursors
such as perfluorooctane sulfonamide (FOSA, 54%; 0.01 to 0.32 ng g
−1
ww wb) and N-ethyl perfluorooctane sul-
fonamide(EtFOSA; 30%; 0.01 to 0.21 ng g
−1
ww wb) were also detected. PFAS accumulation profiles revealeddif-
ferent routes of exposure amongbivalve, crustacean and fish groups.Statistics for left-censored datawere used in
order to minimize bias on trophic magnification factors (TMFs) calculations. TMFs >1 were observed for PFOS
(linear + branched isomers), EtFOSA (linear + branched isomers), and perfluorononanoate (PFNA), and in all
cases, dissimilar accumulation patterns were observed among different trophic positions. The apparent
biodilution of some long-chain PFCAs through the food chain (TMF < 1) may be due to exposure from multiple
PFAS sources. This is the first study investigating bioaccumulation of PFASs in a tropical food web and provides
new insight on the behavior of this ubiquitous class of contaminants.
© 2020 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://
creativecommons.org/licenses/by/4.0/).
Science of the Total Environment 754 (2021) 142146
⁎Corresponding author at: Universidade Federal da Bahia, CIEnAm I, Campus Ondina, CEP: 41170-115 Salvador, Bahia, Brazil.
E-mail addresses: daniele.miranda@aces.su.se,anielemirandaba@hotmail.com (D.A. Miranda).
https://doi.org/10.1016/j.scitotenv.2020.142146
0048-9697/© 2020 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).
Contents lists available at ScienceDirect
Science of the Total Environment
journal homepage: www.elsevier.com/locate/scitotenv
1. Introduction
Per- and polyfluoroalkyl substances (PFASs) represent a diverse
class of fluorinated anthropogenic chemicals with unique physical-
chemical properties including combined hydrophobicity/lipophilicity
(Buck et al., 2011), which make them desirable for a wide range of in-
dustrial processes and products. Notable applications of PFASs include
firefighting foams (Dauchy et al., 2019), fluoropolymer manufacturing
(Song et al., 2018), textiles (Wang et al., 2017), food packaging
(Schaider et al., 2018;Schultes et al., 2019a), cosmetics (Schultes
et al., 2018), and pesticides (Nascimento et al., 2018).
Since the early 2000s, there has been considerable concern over the
widespread occurrence of PFASs, in particular the long-chain
perfluoroalkyl acids (PFAAs), which display high persistence and have
been found in the blood of humans (Giesy and Kannan, 2001;Lau
et al., 2007;Wang et al., 2018) and biota (Dorneles et al., 2008;Giesy
and Kannan, 2001;Hong et al., 2015;Houde et al., 2006b, 2011;
Leonel et al., 2008;Quinete et al., 2009;Routti et al., 2017;
Sanganyado et al., 2018). In 2009, one of the most widespread PFAAs,
perfluorooctane sulfonate (PFOS), along with its salts and synthetic pre-
cursor perfluorooctane sulfonyl fluoride (POSF), were added to Annex B
of the Stockholm Convention on Persistent Organic Pollutants
(UNSCPOPs, 2009). Despite these regulatory initiatives, production of
PFOS and other long-chain PFAAs continues in some parts of the
world under regulatory exemptions.
In Brazil, manufacture and use exemptions have been obtained to
continueproducing and using Sulfluramid, which contains the active in-
gredient N-ethyl perfluorooctane sulfonamide (EtFOSA), a PFOS-
precursor (Avendaño and Liu, 2015;Zabaleta et al., 2018). Sulfluramid
is a formicide extensively introduced as a substitute for the organochlo-
rine pesticide Mirex (Nagamoto et al., 2004), which is mostly used in
eucalyptus and pine plantations to combat leaf-cutting ants (Atta sp.
and Acromyrmex spp.). Brazil is among the largest consumers of
Sulfluramid in the world (MMA, 2015a, 2015b). In the past five years,
around 280 t year
−1
of Sulfluramid (0.3% EtFOSA) was imported, but
there are considerable uncertainties with this estimate since the quan-
tities of Sulfluramid precursor (i.e. POSF) imported into Brazil are not
available (MDIC, 2019). Based on increasing production of eucalyptus
and pine-derived cellulose (Rossato et al., 2018), it is expected that
the demand for Sulfluramid will continue in the next years. A recent
study suggested an elevated contributions from a PFOS precursor (e.g.
Sulfluramid) in Subaé river waters in the Northeast of Brazil (Gilljam
et al., 2016), indicated by the presence of PFOS precursors (PreFOS)
compounds in the region (i.e. perfluorooctane sulfonamide - FOSA/
PFOS ratio above 5:1).
While there are few data available on PFAS-based products beyond
Sulfluramid in Brazil (MDIC, 2019), a recent ecotoxicological study sug-
gested that PFAS-based aqueous film-forming foams (AFFFs) continue
to be used (da Silva et al., 2019) while the National Implementation
Plan for the Stockholm Convention in Brazil (MMA, 2015a, 2015b)high-
lights several PFAS-based consumer products which are permitted. Also,
PFASs contamination in tropical areas has beenassociated with the AFFF
use (Munoz et al., 2017b). Together with the use of Sulfluramid in for-
estry, PFAS-containing AFFF and consumer products (e.g., food packag-
ing, cosmetics, etc.) may be important PFAS sources to the local
environment.
The process of biomagnification is an important mechanism for mi-
cronutrient transfer through food chains (Barclay et al., 1994;Gribble
et al., 2016), but it is also a mechanism that promotes accumulation of
anthropogenic toxic compounds (Borga et al., 2012;Mackay and
Boethling, 2000). Unlike the well-known lipophilic bioaccumulation
mechanism of most of the POPs, the PFASs enter the food chain due to
their proteinophilic characteristics (Goeritz et al., 2013). Despite studies
reporting the occurrence of PFASs in organisms globally (Becker et al.,
2010;Hong et al., 2015;Houde et al., 2006b, 2011;Martin et al., 2003;
Munoz et al., 2017a;Sanganyado et al., 2018), few data are available
for South American ecosystems (Dorneles et al., 2008;Leonel et al.,
2008;Olivero-Verbel et al., 2006;Quinete et al., 2009), and there is a
lack of PFAS bioaccumulation data for tropical ecosystems. Coastal trop-
ical environments concentrate the greatest extent of biodiversity and
productivity (Brown, 2014). Additionally, numerous freshwater and
marine species use these ecosystems (e.g., mangroves and estuaries)
as a nursery, shelter and feeding area, highlighting the importance of
those environments as providers of a myriad of ecosystem services.
Despite its ecological importance, the Subaé estuary, located in
northeastern Brazil (Fig. 1), has been subject to a high load of domestic
and industrial effluents for many years (Hatje et al., 2006;Hatje and
Barros, 2012;Krull et al., 2014;Tavares et al., 1999). Seafood harvested
in the Subaé estuary is still the main protein consumed by the local peo-
ple and may also represent a source of contaminant exposure for the lo-
cals (Souza et al., 2011, 2014). Moreover, the consumption of
contaminated fish and seafood may be a significant pathway of expo-
sure for humans in several countries, including Brazil (Pérez et al.,
2014).
The present study addresses the paucity of data on PFAS accumula-
tion in tropical environments by examining the transfer of these com-
pounds through an estuarine food web. Samples were obtained from
the Subaé estuary, Brazil, and included several speciesof bivalves, crus-
taceans, polychaeta, and fish, along with Suspended Particulate Matter
(SPM), sediment, leave of mangrove trees, and algal biofilm. To the
best of our knowledge, this is the first study investigating the bioaccu-
mulation of PFAS in a tropical food web.
2. Methodology
2.1. Standards and reagents
Authentic standards of 3 (N-alkyl substituted) perfluorooctane sulfon-
amides (N-methylperfluoro-1-octanesulfonamide (MeFOSA), EtFOSA,
and FOSA), 3 perfluoroctane sulfonamidoacetates (N-methylperfluoro-
1-octanesulfonamidoacetic acid (MeFOSAA), Perfluoro-1-octanesulfo
namidoacetic acid (FOSAA), and N-ethylperfluoro-1-octanesulfo
namidoacetic acid (EtFOSAA)), 9 perfluoroalkyl carboxylates (C
5
-C
14
PFCAs: Perfluoropentanoic acid (PFPeA), Perfluoroheptanoic acid
(PFHpA), Perfluorooctanoic acid (PFOA), Perfluorononanoic acid
(PFNA), Perfluorodecanoic acid (PFDA), Perfluoroundecanoic acid
(PFUnDA), Perfluorododecanoic acid (PFDoDA), Perfluorotridecanoic
acid (PFTrDA), Perfluorotetradecanoic acid (PFTeDA), and 4
perfluoroalkyl sulfonates (PFSAs; Perfluorobutanesulfonic (PFBS),
perfluorohexanesulfonic (PFHxS), PFOS, and Perfluorodecanesulfonic
(PFDS)), as well as the isotopically-labelled standards
13
C
4
-PFHpA,
13
C
4
-
PFOA,
13
C
5
-PFNA,
13
C
2
-PFDA,
13
C
2
-PFUnDA,
13
C
2
-PFDoDA,
18
O
2
-PFHxS,
13
C
4
-PFOS,
13
C
8
-FOSA, D
5
-EtFOSA, D
3
-MeFOSAA, D
5
-EtFOSAA, and the re-
covery standards used to monitor internal standards performance
13
C
8
-
PFOA and
13
C
8
-PFOS were purchased from Wellington Laboratories
(Guelph, ON, Canada). Perfluoropentadecanoic acid (PFPeDA) was in-
cluded in the target method as a qualitative analyte. A complete list of
chemicals (Table S1) and reagents used for sample preparation is pro-
vided in the supporting information (SI).
2.2. Study area
The Todos os Santos Bay (BTS) (12°35′30″-13°07′30″S and 038°29′
00″- 038°48′00″W) is characterized by a tropical humid climate, with
an annual mean temperature around 25 °C and precipitation of
2100 mm, with mesotidal semidiurnal tides that control the currents in-
side the bay (Cirano and Lessa, 2007). The bay has three major tribu-
taries: Paraguaçu River (56,300 km
2
), Jaguaripe River (2200 km
2
), and
Subaé River (600 km
2
). While Subaé is the smallest of the three rivers
and has a low mean monthly discharge (9 m
3
s
−1
), it is known to be a
contamination hotspot, due to historically high anthropogenic pressure,
especially regarding the introduction of trace elements to the estuarine
2D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
system (Hatje et al., 2006;Krull et al., 2014). The anthropogenic activi-
ties over the last decade around Subaé river represent a risk to the bio-
diversity and ecological services in the estuary (Almeida et al., 2018;
Gilljam et al., 2016;Hatje and Barros, 2012;Ribeiro et al., 2016). The lit-
erature regarding biocenosis in the BTS is scarce and is focused mostly
on benthic assemblages (e.g. Alves et al., 2020;Magalhães and Barros,
2011;Silvany and Senna, 2019;Barros et al., 2012) and fish (e.g. Dias
et al., 2011;Reis-Filho et al., 2019). Those studies show that estuaries
of BTS are diverse (Alves et al., 2020;Loureiro et al., 2016;Reis-Filho
et al., 2019), and present large variations in the distribution and abun-
dance of assemblages among habitats and salinity gradients (Mariano
and Barros, 2015). The presence of a diverse range of deposit feeder
polychaeta connecting trophic levels is a remarkable characteristic of
BTS food webs (Magalhães and Barros, 2011). Furthermore, the nesting
activity of migratory high trophic level piscivorous birds was previously
reported through the bay (Lunardi et al., 2012).
2.3. Sample collection
Twenty-one species of estuarine organisms were collected, includ-
ing mangrove cupped oyster (Crassostrea rhizophorae, n = 14), man-
grove shellfish (Mytella guyanensis, n = 63), blue crab (Callinectes
sapidus,n= 61), clam (Anomalocardia brasiliana,n= 5), polychaeta
(pooled), mangrove oyster (Crassostrea brasiliana, n = 5), stout tagelus
(Tagelus plebeius,n=60),crab(Ucides cordatus, n = 6), shrimp
(Litopenaeus sp.,n= 18), mangrove tree crab (Goniopsis cruentata,
n= 12), mullet (Mugil sp.,n= 3), torroto grunt (Genyatremus luteus,
n = 3), mojarra (Diapterus sp.,n= 9), silver jenny (Eucinostomus sp.,
n = 1), madamango sea catfish (Cathorops spixii, n = 5), fat snook
(Centropomus parallelus, n = 3), drum (Stellifer sp., n = 6), trevally
(Caranx sp., n = 5), catfish (Aspistor luniscutis, n = 3), common snook
(Centropomus undecimalis, n = 3), and barbel drum (Ctenosciaena
gracilicirrhus, n = 1). Biofilm, mangrove tree leaves, bivalves, crabs,
and polychaeta were manually collected, while shrimp, and fish were
sampled using a fixed gill net (authorization SISBIO n. 61269–3/
28885–4). Almost all sampled species (i.e. bivalves, crabs, shrimp, and
fish) are marketed locally by fishermen (Soares et al., 2011), and none
of them are listed on the IUCN red list as being in danger of extinction.
Samples were collected along a transect and pooled (i.e. sample points
from #1 to #4; Fig. 1). Composite samples of mangrove tree leaves
(Rhizophora mangle,Laguncularia racemosa, and Avicenia sp.), biofilm
(pool of species), and sediments (n= 4) were collected with a clean
spoon at the same location as the benthic organisms (i.e. from site #1
to #4) during the low tide, when sediment was exposed. The SPM was
sampled during the flood tideat three points (n = 3) between the loca-
tion of the sediment sampling sites (SPM#1: between #1 and #2;
SPM#2 between #2 and #3; and SPM#3 between #3 and #4) (Fig. 1).
For each organism, biometry information was recorded (Table S2). Sam-
ples of muscle, liver/hepatopancreas, for fish and crustaceans, soft tis-
sues for bivalves, and whole body for polychaeta were collected. After
dissection, tissue wet weight was recorded and thereafter the samples
were freeze-dried. More details on samplingand pre-treatment are pro-
vided in the Supplementary Information (SI).
2.4. Targeted PFAS analysis
Samples were processed using the method previously described in
Zabaleta et al. (2018). Further details are provided in the SI. Briefly,
Fig. 1. Sediment and bivalve sampling stations of the Subaé, Todos os Santos Bay, BA, Brazil. Suspended particulate matter (SPM) was collected between sediment sampling stations
(SPM#1 between #1 and #2, SPM#2 between #2 and #3, and SPM#3 between #3 and #4). Crustaceans and fishes were sampled between #1 and #4 sampling stations. Sampling
point #1 is the closest station to the lower estuary area, while #4 (12° 35,451′S, 38° 41,672′W) is the closest sampling point to Santo Amaro city.
3D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
isotopically labelled internal standards (2 ng) were added to 0.5 g of
freeze-dried sample (biota or sediment), which were extracted with
acetonitrile (ACN) aided by sonication. The method was repeated for
SPM samples, but with methanol as the extraction solvent. All extracts
were stored in the freezer prior to instrumental analysis.
Instrumental analysis was carried out by ultra-performance liquid
chromatography-tandem mass spectrometry (UPLC-MS/MS; Waters)
operated in negative electrospray ionization mode. Further details on
target PFASs analysis and the mobile phase gradient profile can be
found in the Table S3. Two precursor/product ion transitions, one for
quantification and the other for qualification, were monitored per ana-
lyte (Table S4). Either isotope dilution or aninternal standard approach
using a linear calibration curve with 1/x weighting was used for the
quantitative determination of target compounds. Branched isomers
were determined semi-quantitatively using the calibration curve for
the linear isomer.
2.5. Organic carbon (TOC), total nitrogen (TN) and their stable isotope
analysis
Total organic carbon (TOC), total nitrogen (TN), and stable isotopic
composition of N and C (δ
15
Nandδ
13
C) were analyzed in sediment,
muscle of fish and crustaceans, whole body of polychaeta, and soft tis-
sues of bivalves. Samples were treated with 0.1 M HCl for removing car-
bonates prior to TOC and δ
13
C analysis. Stable isotope ratios were
measured using an isotope ratio mass spectrometer (IsoPrime100,
Isoprime, Cheadle, UK) coupled in continuous flow to an elemental
analyser (vario MICRO cube, Elementar Analysensysteme GmbH,
Hanau, Germany). Carbon and nitrogen isotope ratios were expressed
as δvalues (‰) relative to the Vienna PeeDee Belemnite (vPDB) stan-
dard and to atmospheric N
2
, respectively. We usedInternational Atomic
Energy Agency (IAEA, Vienna, Austria) certified reference materials su-
crose (IAEA-C6, δ
13
C=−10.8 ± 0.5‰; mean ± SD), and ammonium
sulfate as primary standards (IAEA-N2, δ
15
N = 20.3 ± 0.2‰;mean±
SD), and sulphanilic acid as secondary analytical standard (δ
13
C=
−25.6 ± 0.4‰;δ
15
N=−0.1 ± 0.5‰; mean ± SD in each case).
2.6. Quality control
Each batch of 16 samples included blanks (n=2),duplicates(n=
1), and spiked samples (n= 3; 2 ng of individual PFAS per replicate).
Limits of detection (LODs) were estimated based on a signal-to-noise
ratio (S/N) of 3 (Table S5). LODs ranged between 0.02 and
1.11 ng mL
−1
, depending on the tissue and organism analyzed. Calibra-
tion curves (1/x weighting) were constructed using concentrations
from LOD to 150 ng mL
−1
and determination coefficients (R
2
)wereal-
ways in the range of 0.994–0.998. Solvent blanks were run between cal-
ibration curve points in order to monitor carryovers.
PFAS concentrations in blanks were below LOD in all instances.
Spike/recovery experiments showed generally good results, with most
compounds displaying recoveries from 80 to 115%. For a few target
compounds (e.g. PFPeA, PFTrDA, PFTeDA, and EtFOSA), recoveries
were below 60% or above 120% in specific batches. PFPeA was infre-
quently detected (i.e. < 30% of organisms) and was not subject to TMF
determination. Concentrations of PFHpA, PFTrDA, PFTeDA, and L-
EtFOSA, data were used as-is, but these concentrations should be con-
sidered underreported for some species (see Table S6), due to low
spike recoveries for these compounds. In addition, L-EtFOSA was not re-
ported in polychaeta and SPM due to elevated recoveries for these ma-
trices. Perfluorohexanoic acid (PFHxA) was excluded from the set of
data due to interferences with the main transition ion monitored.
2.7. Data handling and statistical analyses
δ
13
Candδ
15
N values were used to determine the trophic web struc-
ture and the trophic position (TP) of the organisms analyzed. The
following equation (Eq. (1)) was subsequently used to estimate the tro-
phic position (TP):
TPconsumer ¼2þδ15Nconsumer−δ15Nbivalve
2:3‰ð1Þ
where 2.3‰is assumed to be the δ
15
N trophic fractionation factor, ac-
cording to McCutchan Jr et al. (2003). TP of the organisms was achieved
by subtracting δ
15
N of bivalves which displayed the lowest δ
15
N values
among animal species (C. rhizophorae;δ
15
N = 8.05, secondary con-
sumer) and are assumed to have a trophic position of 2 (i.e. first con-
sumer level). δ
13
C values without lipid correction were used to predict
organic matter sources.
The Trophic Magnification Factor (TMF) was obtained as 10
slope
by
plotting the logarithm-transformed concentration of a given PFAS
against the TP. TMF calculations were carried out using soft tissue
PFAS concentrations for bivalves, polychaeta and shrimp, and estimated
whole-body PFAS concentrations (C
wb
) for fish and crab according to
the follow equation:
Cwb ¼∑n¼1CtissuePFAS ftissuePFAS ð2Þ
where C
tissue_PFAS
is the concentration of a given PFAS in a specifictissue
and f
tissue_PFAS
is the mass fraction of this tissue in thewhole body. A sep-
arate calculation was performed for each individual organism based on
specific tissue concentration. When the liver concentration was not
available, it was estimated that the concentrations in liver were a factor
of 10 higherthan those of muscle for each compoundbased on thefind-
ings of Nania et al. (2009) (further details are provided in the SI).
Concentrations below the LOD (i.e. non-detected data) were im-
puted as follows: after ln-transformation, the data were fittoacumula-
tive normal distribution using a Generalized Reduced Gradient
algorithm in the Solver add-in feature of Microsoft Excel. Thereafter,
the cumulative normal distribution curve was used to estimate the
missing data at the tail of the distribution as previously described by
Schultes et al. (2019a, 2019b). The imputation of missing data was pre-
viously suggested as a good practice in order to avoid bias introduced in
environmental analyses (Helsel, 2006) arising from the use of either
raw data or one-half LOD (Borga et al., 2012). However, there is no
established threshold for the percentage of left-censored data that
could be used without causing a negative effect (i.e. inflation) of TMFs.
Detection frequencies of 20 (Munoz et al., 2017a) to 40% (Simmonet-
Laprade et al., 2019) have been used in bioaccumulation studies. In
the present study, a moderate threshold (30%) was chosen in order to
include compounds previously detected in the present study area
(Gilljam et al., 2016).
To compare the effect of imputation versus substitution on calcu-
lated TMFs, a linear regression of TL versus PFAS concentration was car-
ried out, using the most frequently detected compounds (i.e. L-PFOS,
93% detection frequency [df]), a compound with a moderate detection
frequency (PFDoDA, 46% df) and the compound with lowest detection
frequency (L-EtFOSA, 30% df) (Fig. S1). In all cases (i.e. regardless of de-
tection frequency), substitution using one-half LOD produced lower
TMF estimates compared to imputation (1.19 and 1.53 for L-PFOS, re-
spectively; 0.55 and 0.94 for PFDoDA, and 1.05 and 1.80 for EtFOSA).
This is probably due to insertion of fixed values into the data set (i.e.
one-half LOD), which tends to decrease the slope of the linear regres-
sion (Borga et al., 2012;Helsel, 2006). While we conclude that substitu-
tion will produce lower TMFs than imputation, overall, the differences
were minimal and did not affect neither the observed biodilution (i.e.
TMF < 1) nor bioaccumulation (i.e. TMF >1) for any compound.
BioEstat 5.0 (Informer Technologies, Inc.), and R statistical software
(R version 3.5.2, 2018-12-20) were used for statistical analyses, with a
critical level of significance of the tests set at (α= 0.05). All PFAS con-
centrations in tissues were log-transformed in order to fit the assump-
tions of the statistical analyses. Significant relationships between TMF
4D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
and PFAS concentrations were determined with nonparametric
Kendall's τ
b
correlations (Munoz et al., 2017a). Kendall's τ
b
correlations
was used due to the slightly better fit of this method to outliers when
compared to the other nonparametric option (i.e. Spearman correla-
tion), besides their smaller gross error sensitivity (GES) and a smaller
asymptotic variance (AV) (Croux and Dehon, 2010;Khamis, 2008).
Non-detects and Data Analysis (NADA) and Linear Mixed-Effects
Models with Censored Responses (LMEC) R-packages were used to per-
form the analyses (Helsel, 2005;Munoz et al., 2017a).
3. Results and discussion
3.1. Subaé estuary food web dynamics
Using the bi-plots of δ
15
N versus δ
13
C, and the feeding habits of each
species (obtained from theliterature (Barletta et al., 2019;Ferreira et al.,
2019;Lira et al., 2018;Matich et al., 2017;Souza et al., 2018;Wellens
et al., 2015)), we projected the energy flow across the food web. δ
15
N
values increased from mangrove tree leaves to omnivore fish, ranging
from 1.06 to 13.8‰(Fig. 2) and all species connect to each other within
the isotopic mixing space. As expected, δ
13
C in mangroves was typical
(−29.3 to −26.6‰) of terrestrialplants which the first product of pho-
tosynthesis is a 3‑carbon molecule (C
3
plants) (Minihan, 1983). A simi-
lar δ
13
C composition was observed in sediment and SPM (−25.4
and −26.9‰, respectively) indicating an influence of mangrove mate-
rial in these matrices. δ
13
Cvariedfrom−25.9‰in bivalve
(C. rhizophorae) to −17.4‰in fish (Caranx sp.). Most bivalves displayed
a close relationship with SPM (i.e. an average of 1.11‰in δ
13
Cenrich-
ment compared to SPM), except for M. guyanensis with an enrichment
of 3.19‰, suggesting an additional source of carbon, probably from sed-
iment (1.67‰enrichment). Among crustaceans, U. cordatus displayed
lower δ
13
Cvalues(−24.8‰) than other crustaceans, which agrees
with their restricted leaf-feeder habit (Wellens et al., 2015), and sedi-
ment intake associated with leaves (average of 3.14‰and 0.58‰, for
carbon enrichment of leaves and sediment, respectively).
Litopenaeus sp. and G. cruentata displayed similar carbon isotopes ra-
tios. These species display generalist feeding habits, but they are found
in different locations within the mangrove forest. G. cruentata develop
an important function in mangroves, recycling carbon in sediments in
non-flooded areas (Wellens et al., 2015), while Litopenaeus sp. is an im-
portant link organism in this tropical food web (Lira et al., 2018;Vinagre
et al., 2018), being the prey of several sampled fish in this study. Among
fish, C. undecimalis showed distinctive low values of δ
13
C(−23.8‰),
even for a carnivore species showing a link to mangrove basal resource.
In general, fish had δ
15
Nandδ
13
C average values of 12.6 ± 0.84‰
and −20.7 ± 1.91‰(average ± SD), respectively. Polychaeta and
shrimp represent the main food source for these fish species (Lira
et al., 2018;Loureiro et al., 2016;Martins et al., 2017;Souza et al.,
2018), which is aligned with C and N isotope results (Fig. 2). However,
the large carbon variation between fish species may indicate an alterna-
tive source of food not considered in the presentstudy and thefact they
are not resident all the time in the studied region.
Bivalves represented the lowest TP of the food chain (TP = 2.0,
C. rhizophorae; to TP = 2.6, T. plebeius), followed by crustaceans
(TP = 2.1, C. sapidus to TP = 3.0, G. cruentata), polychaeta (TP = 3.6,
pool of species) and fish (TP = 3.2, Mugil sp. to TP = 4.5,
C. gracilicirrhus).
3.2. PFAS concentrations in biota
Sum (∑)
22
PFAS concentrations (and their isomers) above LOD in
biota ranged from 0.28 ng g
−1
ww wb (C. rhizophorae) to 3.72 ng g
−1
ww wb (Litopenaeus sp.) (Fig. 3, Table S7) which is lower than the
ranges reported for biota in subtropical (∑
21
PFASs; 0.50 to
17.5 ng g
−1
ww wb) (Loi et al., 2011), temperate (∑
23
PFASs; 0.66 to
45.0 ng g
−1
ww wb) (Munoz et al., 2017a) and polar areas (∑
5
PFASs
2.00 to 43.0 ng g
−1
ww wb) (Tomy et al., 2009). PFOS was the most
abundant PFAS among all species, with the exception of 3 bivalve sam-
ples (whichwere dominated by FOSA) and 1 fish (in which EtFOSA was
dominant). Beyond the consistent observation of PFOS, the PFAS profile
among individual species varied considerably. For example, most crus-
tacean, polychaeta and bivalve samples contained relatively high levels
of PFCAs compared to fish (Fig. 3). Also notable was that PFNA was the
dominant PFCA among most crustacean, fish and polychaeta samples,
but was generally not detected in bivalve. Differences in feed behavior,
niche and protein contents of each group of organisms can help to ex-
plain this dissimilarity. Furthermore, more than one source of contami-
nants for those organisms could also play an important role in
0
4
8
12
16
-35-30-25-20-15-10
δ
15N (‰)
δ13C (‰)
SPM (n = 3)
Sediment (n = 4)
Leaf (pool)
Biofilm (pool)
Bivalve (n = 5)
Polychaeta (pool)
Crustacean (n = 10)
Fish (n = 28)
Fig. 2. Distributionof δ
13
Candδ
15
N in biotic andabiotic samples from theSubaé estuary, NE Brazil. Where SPM represents the suspended particulate matter and n the numberof replicates
analyzed.
5D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
elucidating the PFASs accumulation patterns. While parts of the river
can be under direct effect of sewage (consequently more exposed to
AFFF and consumer products), other portions may be more affected by
the leaching of forestry carrying Sulfluramid straight in the course of
the river. Inputs of metals and nutrients have already been observed
for different segments of the Subaé estuarine system (i.e. upper and
lower estuary) (da Silva et al., 2017;Motta et al., 2018), but these data
are unknown for PFASs. While most of the sampled specimens (i.e. bi-
valve, crustacean, and polychaeta) show either sessile behavior or lim-
ited mobility, fish can be exposed to a higher range of PFAS sources by
swimming along the river and outside the estuary. Overall, the fish an-
alyzed in the present study contained lower concentrations of the major
PFCAs (PFNA, PFUnDA, PFDoDA, PFTrDA, PFTeDA) than other sampled
organisms (Fig. 3). The reason for this is unclear but may indicate differ-
ent exposure routes and/or PFAS accumulation in fish. Even though
most PFCAs were correlated with one another (p< 0.05; Table S8), cor-
relations between PFCAs and PFOS were rare, suggesting a different
source of contaminants for these organisms.
In Laguncularia racemosa (white mangrove) and Rhizophora mangle
(red mangrove) leaves, only FOSAA was observed (0.16 and
0.07 ng g
−1
ww, respectively) (Table S9). The PFASs were not detected
in the biofilm sample (i.e. concentrations < LOD). The elevated concen-
trations of PFAS, and in particular L-PFOS in shrimps (1.97 ng g
−1
ww
wb) when compared to other organisms are consistent with the results
of Carlsson et al. (2016), who proposed that the higher protein content
in shrimp, compared to fish, may lead to higher PFAS concentrations in
the former species. The detritivore habit of shrimp may also help to ex-
plain the observed profile, since the SPM showed detectable PFAS con-
centrations, and together with sediment it is an important route of
uptake of organic compounds for aquatic epibenthic organisms (Liu
Fig. 3. Sum PFAS concentrations (ng g
−1
wet weight (ww) for the whole body) measured indifferent organisms from the Subaé estuary (a); and relative PFAS profiles (normalized to
100%) (b). Number of organisms analyzed is shown in parentheses. EtFOSA was not reported for polychaeta. Br-PFOA, L-PFDS, Br-PFDS, L-MeFOSAA, Br-MeFOSAA, Br-FOSAA,
MeFOSAA, and Me-EtFOSAA were not added to the figure due to either detection <LOD or low concentrations.
6D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
et al., 2018). For fish, L-PFOS concentrations were an order of magnitude
lower than those reported for fish from a subtropical region (Loi et al.,
2011), but similar to those reported for several species of fish from the
Arctic (Butt et al., 2010;Tomy et al., 2004)andtemperate(Munoz
et al., 2018) zones. Fish and bivalve from the present work also
contained lower ƩPFAS when compared with two highly industrialized
areas in Riode Janeiro (Ʃ
8
PFAS, Guanabara Bay and Paraíba do Sul River)
(Quinete et al., 2009). While PFAS profiles were similar among the sam-
pled areas, the EtFOSA transformation product FOSA was absent in sam-
ples from the tributaries and coastal water from Rio de Janeiro, as well
as bivalves and fish muscles, whereas it was present only in the liver
of a few fish samples. This is perhaps unsurprising considering that
sales of Sulfluramid in Rio de Janeiro were ~ 100 times lower than in
Bahia state (i.e. between 2014 and 2017) (IBAMA, 2017)andthereare
no known records of Eucalyptus and Pine plantations in the area.
These previous studies together with the present work may suggest
that, although Sulfluramid is not the only source of PFASs for the
Subaé River, the use of this pesticide in the region might contribute to
greater detection of FOSA in different environmental matrices.
3.3. PFAS concentrations in abiotic samples
In SPM (n= 3), 16 PFASs were detected, including 11 linear and 5
branched isomers. Sum PFAS concentrations ranged from 1.85 ng g
−1
dw (SPM#1) to 7.25 ng g
−1
dw (SPM#2) (Fig. 4, Table S10). PFNA
was the predominant PFAS in samples closer to the lower estuary
(Fig. 1), while PFUnDA was the prevalent compound in the upper estu-
ary. In comparison, 5 linear PFAS isomers were detected in sediments
(n= 4), with sum PFAS concentrations in the range of 0.10 ng g
−1
dw
(#2) to 0.33 ng g
−1
dw (#1) (Table S11). L-PFOS and L-PFOA were de-
tected in all sediment samples at very low concentrations (L-PFOS:
0.07–0.34 ng g
−1
ww and L-PFOA: 0.35–0.79 ng g
−1
ww), while PFOS
was detected in all SPM samples ranging from 0.48 to 1.56 ng g
−1
dw.
The presence of L-EtFOSA in sediments and its degradation products
(e.g. L-FOSA) in SPM suggest that, even though the overall concentra-
tions in Subaé estuary are low, there may be a source of Sulfluramid in
the region. The use of Sulfluramid in this area was previous suggested
based on the high FOSA:PFOS detected in the Subaé river (Gilljam
et al., 2016). The formicide can be mobilized from plantations to rivers
and estuaries due to the leaching of contaminated soils. Stahl et al.
(2013) and Zabaleta et al. (2018) suggested that leaching of soils caused
by rainwater can promote the transport of EtFOSA and its degradation
products to groundwater, rivers, and coastal zones. Sources other than
Sulfluramid may also be important. For example, EtFOSAA, which was
observed in SPM and sediment samples, does not form from EtFOSA,
but rather from N-ethyl perfluorooctane sulfonamidoethanol (N-
EtFOSE) a substance used widely in consumer products (e.g., food pack-
aging) until 2002 (Benskin et al., 2013;MejiaAvendaño and Liu, 2015;
USEPA, 2002).
Previous work indicated that sediments and soils can act as a sink for
long chain PFASs, due to their sediment-water partitioning coefficient
(Nascimento et al., 2018). Several species investigated in the present
study display either benthic or demersal habits, which indicates that
sediment exposure may represent a secondary source of PFASs for
those organisms (Munoz et al., 2017a), in addition to bioaccumulation
through the food chain. Nonetheless, estuarine areas are naturally dy-
namic and dilution of organic matter (and consequently PFASs) may
occur at the site, thereby lowering observed concentrations in abiotic
matrices.
3.4. Bioaccumulation of PFASs in the food chain
Bioaccumulation of PFASs in the present food web was evaluated for
L- and Br-PFOS, PFTrDA, L-FOSA, PFUnDA, PFDoDA, PFTeDA, PFNA, and
L- and Br-EtFOSA. TMFs and biomagnification factors (BMFs) greater
than 1.0 are both considered reliable indicators of bioaccumulation
(Borga et al., 2012). However, TMF is considered the most consistent
way to characterize biomagnification of organic compounds over an en-
tire food web (Franklin, 2016). TMFs were greater than 1 for L- and Br-
PFOS, L- and Br-EtFOSA, and PFNA, indicating biomagnification across
trophic levels. The concentration of four PFASs were significantly corre-
lated with trophic level (p< 0.05; Table 1, Kendell's rank correlation;
Fig. 4. Sum of PFASconcentrationsin different samplesfrom the Subaé estuary. (a) Concentrationsare expressed in ngg
−1
dry weight(dw) for sediment and SPM, and ng g
−1
wet weight
(ww) for leave and (b) corresponding compositionprofiles (relativeabundance in % of ΣPFAS).Sed#3, Black mangroveand biofilm were excluded fromthis figure due no detected PFASs
above LOD.
7D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
Figs. S2 and S3), showing bioaccumulation of L-PFOS (n=44;τ=
0.215, p= 0.043) and biodilution of PFUnDA (n = 44; τ=−0.226,
p= 0.034), PFTrDA (n = 44; τ=−0.386, p≤0.001), PFTeDA (n =
44; τ=−0.329, p= 0.002). These data further highlight the potential
of PFOS to be biomagnified throughthe tropical food web, which is con-
sistent with prior studies in temperate and subtropical zones (Liu et al.,
2018;Loi et al., 2011;Munoz et al., 2017a).
The absence of TMFs greater than 1 for most PFCAs was initially sur-
prising, since bioaccumulation of C
9
-C
12
PFCAs has been reported else-
where (Houde et al., 2006a;Munoz et al., 2017a). However,
biodilution has also been reported for C
13
-C
14
PFCAs (Munoz et al.,
2017a), which may be associated with the limitation of PFCAs with
>10 fluorinated carbons to penetrate cell membranes due to their
large molecular size (Conder et al., 2008;Hong et al., 2015). Alterna-
tively, different sources of PFCAs for (1) bivalve, crustacean, and
polychaeta, and (2) fish may also lead to an apparentlack of bioaccumu-
lation. The low detection frequency of those compounds in organisms
(i.e. from 41% (PFNA) to 57% (PFTrDA)) should also be considered.
The TMF determined here for PFOS was lower than previously re-
ported for polar food webs, but similar to thos e in subtropical regions
(Table S12) (Boisvert et al., 2019;Loi et al., 2011;Munoz et al.,
2017a;Tomy et al., 2004). Variations in PFAS sources and/or proxim-
ity to sources may play a role in explaining these differences. For ex-
ample, biota from the Northern Hemisphere tends to contain higher
PFAS concentrations due to its proximity to point sources (Benskin
et al., 2012). On the other hand, the ongoing use of EtFOSA in South
America is unique, and exposure tothiscompoundmaycausePFOS
accumulation in the Subaé estuarine wildlife due to its biotransfor-
mation (Gilljam et al., 2016;Nascimento et al., 2018). In addition,
the upper trophic levels of polar food webs is occupied by top pred-
ator mammals, which means higher energy flow compared to tropi-
cal environments (Hobson et al., 2002). The lower TMFs in tropical
aquatic ecosystems possibly reflects not only the lower input of
PFASs, but also the structure of food webs that generally larger num-
ber of organisms, which may dilute the energy flows (and by exten-
sion, contaminants) among them (Borga et al., 2012). Also, variances
in protein content in the blood and tissues of distinct species can re-
sult in variability in the accumulation of proteinophilic co ntaminants
(Goeritz et al., 2013). Finally, climatic factors cannot be excluded
(e.g. water temperature) and water chemistry, which may also
have an impact on PFAS accumulation at different trophic levels
(Vidal et al., 2019;Xu et al., 2014).
Surprisingly, the TMF for L-EtFOSA was above 1 (1.80), suggesting
that this compound is bioaccumulating in the Subaé estuarine food
chain. However, since EtFOSA was rarely detected throughout the food
chain and also in low concentrations (0.01 to 0.21 ng g
−1
ww wb, 30%
detection frequency), TMFs should be interpreted cautiously, and fur-
ther studies are warranted to corroborate these results. To the best of
our knowledge, TMFs for EtFOSA have not been previously reported,
but Tomy et al. (2004) showed Trophic Position-adjusted BMFs above
1 for species from the Arctic web food and suggested that organisms
from different TP will metabolize/accumulate EtFOSA in different pro-
portions. It is well-known that EtFOSA is readily biotransformed into
FOSA in a few types of soil under aerobic conditions (Avendaño and
Liu, 2015;Yin et al., 2018;Zabaleta et al., 2018), and also in earthworms
(Zhao et al., 2018) but no studies have been carried out in estuarine en-
vironments. The presence of PreFOS (i.e. L-EtFOSA, L-EtFOSAA, and
FOSA) in the Subaé biota may contribute indirectly to the occurrence
of PFOS in top-level organisms via biotransformation, which may in
turn contribute to increasing the TMF of PFOS (Simonnet-Laprade
et al., 2019). This is supported by the observed biodilution of FOSA
through the food web. Nonetheless, an in-depth evaluation of a larger
range of PreFOS (and long-chain PFCAs precursors) will be necessary
to confirm if TMF inflation is occurring due to PFAS precursors
biotransformation.
Interestingly, when EtFOSA TMFs were re-calculated after split-
ting the organisms into 2 groups based on their trophic level (i.e.
group 1: n= 18, TP 2.00 to 3.22; and group 2: n= 26, TP 3.56 to
4.49), a much higher TMF was observed for group 1 (2.65) compared
to group 2 (0.22), suggesting that dilution of this compound is occur-
ring in top organisms (Table S13). Besides the differences related to
habitats and trophic level between the aforementioned groups, the
first group (TP 2.00 to 3.22) is composed mainly by filter-feeding or-
ganisms, while the second group (TP 3.56 to 4.49) comprises mostly
omnivorous organisms. The different feeding habits of those organ-
isms can imply dissimilar contaminant exposure and hence uptake.
Higher TMFs for bottom species were previously reported for other
PFASs (Munoz et al., 2017a), and may point to greater efficiency in
metabolizing and/or excreting xenobiotics in higher trophic level or-
ganisms. This metabolic difference in the accumulation/depuration
of PFAA precursors between fish and invertebrates has been previ-
ously reported (Langberg et al., 2019).
The divergence in TMFs observed for EtFOSA upon splitting the
food web into two groups was not observed for PFNA, L-PFOS, or
Br-PFOS (Table S13). These substances all displayed TMFs >1 regard-
less of whether the food web was split into two groups or considered
together, suggesting a linear trend in bioaccumulation among the
sampled organisms. However, this was not the case for L-FOSA and
C
11
-C
14
PFCAs. These substances all displayed apparent biodilution
(i.e. TMF < 1) when the entire food web was considered but upon
splitting the food web into two groups, L-FOSA, PFUnDA, and
PFDoDA displayed bioaccumulation (i.e. TMFs>1) for both groups,
while PFTrDA and PFTeDA displayed biodilution (i.e. TMF < 1) in
the first group and bioaccumulation (i.e. TMF > 1) in the second
group, which was the opposite of what was observed for EtFOSA.
The factors controlling these phenomena are currently unclear and
warrant further investigation, once the metabolization of PFASs by
different organisms can help to explain the bioaccumulation pattern
of these compounds through the food chain.
Table 1
Regressions of Log Concentration (concentrations expressed in ng g
−1
ww whole body,
min-max) vs TP (trophic position) in the whole food web, obtained with the NADA R-
package
a
, including the p-value ofthe regression, and Kendall's τcorrelation coefficient
(Helsel,2005;Munoz et al.,2017a), compoundsdetection frequency
a
, minimumand max-
imum whole-body wet weight concentrations, and Kendall's τcorrelation coefficient
b
.
EtFOSA was not reported for polychaeta.
Compound Detection
frequency (%)
Concentration average
(min-max)
τTMF
FW
(min–max)
L-PFOS 93 0.49
(0.05–1.97)
0.215 1.53
(1.49–1.57)
Br-PFOS 73 0.09
(0.01–0.35)
0.052 1.53
(1.45–1.61)
PFTrDA 57 0.06
(0.01–0.28)
−0.386 0.39
(0.38–0.40)
L-FOSA 54 0.06
(0.01–0.32)
−0.177 0.64
(0.62–0.66)
Br-FOSA 27 0.02
(0.01–0.24)
−0.119 0.83
(0.78–0.88)
PFUnDA 48 0.04
(0.01–0.17)
−0.226 0.41
(0.39–0.43)
PFDoDA 46 0.03
(0.01–0.15)
−0.075 0.94
(0.88–1.00)
PFTeDA 46 0.05
(0.01–0.26)
−0.329 0.62
(0.48–0.80)
PFNA 41 0.03
(0.01–0.42)
0.029 1.34
(1.27–1.42)
L-EtFOSA 30 0.03
(0.01–0.21)
0.112 1.80
(1.72–1.89)
Br-EtFOSA 30 0.09
(0.01–0.58)
−0.021 1.60
(1.38–1.86)
a
Note that for detection frequency percentage of individual tissue (muscle, liver, he-
patopancreas and soft tissue of bivalves and shrimp, n=56)wasused;
b
It was obtained with the NADA R-package, including Kendall's τcorrelation coeffi-
cient,the resulting trophicmagnificationfactor (TMF = 10
slope
) with theconfidence inter-
val of 95% (min –max). Bold numbers indicate significant regressions (p<0.05).
8D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
4. Conclusions
The bioaccumulation of PFASs was accessed in a tropical estuary in
northeastern Brazil. PFASs were detected in all organisms analyzed,
alongside with abiotic matrices (SPM and sediment). Unique accumula-
tion profiles and low levels of PFASs were generally found in biotic and
abiotic matrices when compared to subtropical, temperate and polar
environments. Different patterns of bioaccumulation were observed
through the sampled group of organisms (i.e. bivalve, crustacean,
polychaeta, and fish), suggesting a setof factors (physical, chemical, bi-
ological and spatial) must be influencing on the PFAS assimilation.
Moreover, PFOS, PFNA and EtFOSA were found to be bioaccumulating
in this food chain, even though the bioaccumulation of the latter com-
pound should be considered cautiously due to the low detection of
EtFOSA in organisms. Nonetheless, the occurrence of PFASs in coastal
marine biota is particularly concerning due to the potential for human
exposure via consumption of these organisms. Further investigations
into the sources of PFASs to Subaé estuary are warranted.
CRediT authorship contribution statement
Daniele A. de Miranda: Conceptualization, Investigation, Writing -
original draft, Writing - review & editing. Jonathan P. Benskin: Writing
- review & editing, Visualization, Resources. Raed Awad: Investigation,
Writing - review & editing. Gilles Lepoint: Formal analysis. Juliana
Leonel: Writing - review & editing, Visualization. Vanessa Hatje: Con-
ceptualization, Writing - review & editing, Resources, Visualization,
Supervision.
Declaration of competing interest
There are no conflicts of interest to declare.
Acknowledgments
This work was supported by Fundação de Amparo à Pesquisa do
Estado da Bahia, Brazil (FAPESB) (PET0034/2012), CNPQ (441829/
2014-7), and The Rufford Foundation (n° 2992-1). The authors were
supported by FAPESB (D. Miranda, n° BOL0122/2017), Coordenação de
Aperfeiçoamento de Pessoal de Nível Superior, Brazil (Ph.D. Sandwich,
D. Miranda, n° 88881.188589/2018-01), and CNPq (V. Hatje - 304823/
2018-0). J. Leonel were sponsored by CNPq (401443/2016-7 and
310786/2018-5). G. Lepoint is supported by the F.R.S.-FNRS. Thanks to
Krishna Das for the collaboration with the Isotope Data. Thanks to
Taiana A. Guimarães for the sample collection and sample preparation
for the isotopic analyses.
Appendix A. Supplementary data
Supplementary data to this article can be found online at https://doi.
org/10.1016/j.scitotenv.2020.142146.
References
Almeida, M., Nascimento, D.V., Mafalda, P.O., Patire, V.F., Albergaria-Barbosa, A.C.R.,
2018. Distribution and sources of polycyclic aromatic hydrocarbons (PAHs) in
surface sediments of a Tropical Bay influenced by anthropogenic activities
(Todos os Santos Bay, BA, Brazil). Mar. Pollut. Bull. 137, 399–407. https://doi.
org/10.1016/j.marpolbul.2018.10.040.
Alves, A.T., Petsch, D.K., Barros, F., 2020. Drivers of benthic metacommunity structure
along tropical estuaries. Sci. Rep. 10, 1739. https://doi.org/10.1038/s41598-020-
58631-1.
Avendaño, S., Liu, J., 2015. Production of PFOS from aerobic soil biotransformation of two
perfluoroalkyl sulfonamide derivatives. Chemosphere 119, 1084–1090. https://doi.
org/10.1016/j.chemosphere.2014.09.059.
Barclay, W.R., Meager, K.M., Abril, J.R., 1994. Heterotrophic production of long chain
omega-3 fatty acids utilizing algae and algae-like microorganisms. J. Appl. Phycol. 6,
123–129. https://doi.org/10.1007/BF02186066.
Barletta, M., Lima, A.R.A., Costa, M.F., 2019. Distribution, sources and consequences of
nutrients, persistent organic pollutants, metals and microplastics in South American
estuaries. Sci. Total Environ. 651, 1199–1218. https://doi.org/10.1016/j.
scitotenv.2018.09.276.
Barros, F., de Carvalho, G.C., Costa, Y., Hatje, V., 2012. Subtidal benthic macroinfaunal
assemblages in tropical estuaries: generality amongst highly variable gradients.
Mar. Environ. Res. 81, 43–52. https://doi.org/10.1016/j.marenvres.2012.08.006.
Becker, A.M., Gerstmann, S., Frank, H., 2010. Perfluorooctanoic acid and perfluorooctane
sulfonate in two fish species collected from the Roter Main River, Bayreuth,
Germany. Bull. Environ. Contam. Toxicol. 84, 132–135. https://doi.org/10.1007/
s00128-009-9896-0.
Benskin, J.P., Muir, D.C.G., Scott, B.F., Spencer, C., De Silva, A.O., Kylin, H., Martin, J.W.,
Morris, A., Lohmann, R., Tomy, G., Rosenberg, B., Taniyasu, S., Yamashita, N., 2012.
Perfluoroalkyl acids in the Atlantic and Canadian Arctic oceans. Environ. Sci. Technol.
46, 5815–5823. https://doi.org/10.1021/es300578x.
Benskin, J.P., Ikonomou, M.G., Gobas, F.A.P.C., Begley, T.H., Woudneh, M.B., Cosgrove, J.R.,
2013. Biodegradation of N-ethyl perfluorooctane sulfonamido ethanol (EtFOSE) and
EtFOSE-based phosphate diester (SAmPAP Diester) in marine sediments. Environ.
Sci. Technol. 47, 1381–1389. https://doi.org/10.1021/es304336r.
Boisvert, G., Sonne, C., Rigét, F.F., Dietz, R., Letcher, R.J., 2019. Bioaccumulation and
biomagnification of perfluoroalkyl acids and precursors in East Greenland polar
bears and their ringed seal prey. Environ. Pollut. 252, 1335–1343. https://doi.org/
10.1016/j.envpol.2019.06.035.
Borga, K., Kidd, K.A., Muir, D.C.G., Berglund, O., Conder, J.M., Gobas, F.A.P.C., Kucklick, J.,
Malm, O., Powell, D.E., 2012. Trophicmagnification factors: considerations of ecology,
ecosystems, and study design. Integr. Environ. Assess. Manag. 8, 64–84. https://doi.
org/10.1002/ieam.244.
Brown, J.H., 2014. Why are there so many species in the tropics? J. Biogeogr. 41, 8–22.
https://doi.org/10.1111/jbi.12228.
Buck, R.C., Franklin, J., Berger, U., Conder, J.M., Cousins, I.T., de Voogt, P., Jensen, A.A.,
Kannan, K., Mabury, S.A., van Leeuwen,S.P.J., 2011. Perfluoroalkyl and polyfluoroalkyl
substances in the environment: terminology, classification, and origins. Integr. Envi-
ron. Assess. Manag. 7, 513–541. https://doi.org/10.1002/ieam.258.
Butt, C.M., Berger, U., Bossi, R., Tomy, G.T., 2010. Levels and trends of poly- and
perfluorinated compounds in the arctic environment. Sci. Total Environ. 408,
2936–2965. https://doi.org/10.1016/j.scitotenv.2010.03.015.
Carlsson, P., Crosse, J.D., Halsall, C., Evenset, A., Heimstad, E.S., Harju, M., 2016.
Perfluoroalkylated substances (PFASs) and legacy persistent organic pollutants
(POPs) in halibut and shrimp fromcoastal areas in the far northof Norway: small sur-
vey of important dietary foodstuffs for coastal communities. Mar. Pollut. Bull. 105,
81–87. https://doi.org/10.1016/j.marpolbul.2016.02.053.
Cirano, M., Lessa, G.C., 2007. Oceanographic characteristics of BaÃ\-a de Todos os Santos,
Brazil. Rev. Bras. GeofÃ\-sica 25, 363–387.
Conder, J.M., Hoke, R.A., Wolf, W. de, Russell, M.H., Buck, R.C., 2008. Are PFCAs
bioaccumulative? A critical review and comparison with regulatory criteria and per-
sistent lipophilic compounds. Environ. Sci. Technol. 42, 995–1003. https://doi.org/
10.1021/es070895g.
Croux, C., Dehon, C., 2010. Influence functions of the Spearman and Kendall correlation
measures. Stat. Methods Appt. 19, 497–515. https://doi.org/10.1007/s10260-010-
0142-z.
da Silva, G.S., Gloaguen, T.V.,Couto, C.F., Motta, P.N.S.D., 2017. Persistence and mobility of
metals in an estuarine environment 25 years after closure of a lead smelter, Bahia
State, Brazil. Environ. Earth Sci. 76, 548. https://doi.org/10.1007/s12665-017-6886-0.
da Silva, S.C., Pusceddu, F.H., dosSantos Barbosa Ortega, A., de Souza Abessa, D.M., Pereira,
C.D.S., Maranho, L.A., 2019. Aqueous film-forming foams (AFFFs) are very toxic to
aquatic microcrustaceans. Water Air Soil Pollut. 230, 260. doi:https://doi.org/
10.1007/s11270-019-4291-x.
Dauchy, X., Boiteux, V., Colin, A., Bach, C., Rosin, C., Munoz, J.-F., 2019. Poly- and
perfluoroalkyl substances in runoff water and wastewater sampled at a firefighter
training area. Arch. Environ. Contam. Toxicol. 76, 206–215. https://doi.org/10.1007/
s00244-018-0585-z.
Dias, J.F.,Gonçalves,A.M., Fernandez, W.S., Silbiger, H.L.N., Fiadi, C.B., Schmidt, T.C. dos S.,
2011. Ichthyofauna in an estuary of the Mataripe area, Todos os Santos Bay, Bahia,
Brazil. Braz. J. Oceanogr. 59 (1). https://doi.org/10.1590/S1679-87592011000100006.
Dorneles,P.R., Lailson-Brito, J., Azevedo, A.F., Meyer, J., Vidal, L.G., Fragoso, A.B., Torres, J.P.,
Malm, O., Blust, R., Das, K., 2008. High accumulation of perfluorooctane sulfonate
(PFOS) in marine tucuxi dolphins (Sotalia guianensis) from the Brazilian coast. Envi-
ron. Sci. Technol. 42, 5368–5373. https://doi.org/10.1021/es800702k.
Ferreira, G.V.B., Barletta, M., Lima, A.R.A., Morley, S.A., Costa, M.F., 2019. Dynamics of ma-
rine debris ingestion by profitable fishes along the estuarine Ecocline. Sci. Rep. 9,
13514. https://doi.org/10.1038/s41598-019-49992-3.
Franklin, J., 2016. How reliable are field-derived biomagnification factors and trophic
magnification factors as indicators of bioaccumulation potential? Conclusions from
a case study on per- and polyfluoroalkyl substances. Integr. Environ. Assess. Manag.
12, 6–20. https://doi.org/10.1002/ieam.1642.
Giesy, J.P., Kannan, K., 2001. Global distribution of perfluorooctane sulfonate in wildlife.
Environ. Sci. Technol. 35, 1339–1342. https://doi.org/10.1021/es001834k.
Gilljam, J.L., Leonel, J., Cousins,I.T., Benskin, J.P., 2016. Is ongoing Sulfluramid use in South
America a significant source of perfluorooctanesulfonate (PFOS)? Production inven-
tories, environmental fate, and local occurrence. Environ. Sci. Technol. 50, 653–659.
https://doi.org/10.1021/acs.est5b04544.
Goeritz, I., Falk, S., Stahl, T., Schaefers, C., Schlechtriem, C., 2013. Biomagnification and tis-
sue distribution of perfluoroalkyl substances (PFASs) in market-size rainbow trout
(Oncorhynchus mykiss). Environ. Toxicol. Chem. 32, 2078–2088. https://doi.org/
10.1002/etc.2279.
Gribble,M.O., Karimi, R., Feingold, B.J., Nyland, J.F., O’Hara,T.M., Gladyshev, M.I.,Chen, C.Y.,
2016. Mercury, selenium and fish oils in marine food webs and implications for
9D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
human health. J. Mar. Biol. Assoc. 96, 43–59. https://doi.org/10.1017/
S0025315415001356 UNITED KINGDOM.
Hatje, V., Barros, F., 2012. Overview of the 20th century impact of trace metal contamina-
tion in the estuaries of Todos os Santos Bay: past, present and future scenarios. Mar.
Pollut. Bull. 64, 2603–2614. https://doi.org/10.1016/j.marpolbul.2012.07.009.
Hatje, V., Barros, F., Figueiredo, D.G., Santos, V.L.C.S., Peso-Aguiar, M.C., 2006. Trace metal
contamination and benthic assemblages in Subaé estuarine system, Brazil. Mar.
Pollut. Bull. 52, 982–987. https://doi.org/10.1016/j.marpolbul.2006.04.016.
Helsel, D.R., 2005. Nondectects and Data Analysis; Statistics for Censored Environmental
Data. John Wiley and Sons.
Helsel, D.R., 2006. Fabricating data: how substituting values for nondetects can ruin re-
sults, and what can be done about it. Chemosphere 65, 2434–2439. https://doi.org/
10.1016/j.chemosphere.2006.04.051.
Hobson, K.A., Fisk, A., Karnovsky, N., Holst, M., Gagnon, J.-M., Fortier, M., 2002. A stable
isotope (δ13C, δ15N) model for the North Water food web: implications for evaluat-
ing trophodynamics and the flow of energy and contaminants. Deep Sea Res. Part II
Top. Stud. Oceanogr. 49, 5131–5150. https://doi.org/10.1016/S0967-0645(02)
00182-0.
Hong, S., Khim, J.S., Wang, T., Naile, J.E., Park, J., Kwon, B.-O., Song, S.J., Ryu, J., Codling, G.,
Jones, P.D., Lu, Y., Giesy, J.P., 2015. Bioaccumulation characteristics of perfluoroalkyl
acids (PFAAs) in coastal organisms from the west coast of South Korea. Chemosphere
129, 157–163. https://doi.org/10.1016/j.chemosphere.2014.06.023.
Houde, M., Bujas, T.A.D., Small, J., Wells, R.S., Fair, P.A., Bossart, G.D., Solomon, K.R., Muir,
D.C.G., 2006a. Biomagnification of perfluoroalkyl compounds in the bottlenose dol-
phin (Tursiops truncatus) food web. Environ. Sci. Technol. 40, 4138–4144. https://
doi.org/10.1021/es060233b.
Houde, M., Martin, J.W., Letcher, R.J., Solomon, K.R., Muir, D.C.G., 2006b. Biological moni-
toring of polyfluoroalkyl substances: a review. Environ. Sci. Technol. 40, 3463–3473.
https://doi.org/10.1021/es052580b.
Houde, M., De Silva, A.O., Muir, D.C.G., Letcher, R.J., 2011. Monitoring of perfluorinated
compounds in aquatic biota: an updated review. Environ. Sci. Technol. 45,
7962–7973. https://doi.org/10.1021/es104326w.
IBAMA, Institute of Environment and Natural Resources, 2017. Sales of Active
Ingredientsby State Brasilia.
Khamis, H., 2008. Measures of association: how to choose? J. Diagnostic Med. Sonogr. 24,
155–162. https://doi.org/10.1177/8756479308317006.
Krull, M., Abessa, D.M.S., Hatje, V., Barros, F., 2014. Integrated assessment of metal con-
tamination in sediments from two tropical estuaries. Ecotoxicol. Environ. Saf. 106,
195–203. https://doi.org/10.1016/j.ecoenv.2014.04.038.
Langberg, H.A., Breedveld, G.D., Grønning, H.M., Kvennås, M., Jenssen, B.M., Hale, S.E.,
2019. Bioaccumulationof fluorotelomer sulfonates and perfluoroalkyl acidsin marine
organisms living in aqueous film-forming foam impacted waters. Environ. Sci.
Technol. 53, 10951–10960. https://doi.org/10.1021/acs.est.9b00927.
Lau, C., Anitole, K., Hodes, C., Lai, D., Pfahles-Hutchens, A., Seed, J., 2007. Perfluoroalkyl
acids: a review of monitoring and toxicological findings. Toxicol. Sci. 99, 366–394.
Leonel, J., Kannan, K., Tao, L., Fillmann, G., Montone, R.C., 2008. A baseline study of
perfluorochemicals in Franciscanadolphin and Subantarctic fur seal from coastal wa-
ters of Southern Brazil. Mar. Pollut. Bull. 56, 778–781. https://doi.org/10.1016/j.
marpolbul.2008.01.012.
Lira, A., Angelini, R., Le Loc’h, F., Menard, F., Lacerda, C., Fredou, T., Fredou, F.L., 2018. Tro-
phic flow structure of a neotropical estuary in northeastern Brazil and the compari-
son of ecosystem model indicators of estuaries. J. Mar. Syst. 182, 31–45. https://doi.
org/10.1016/j.jmarsys.2018.02.007.
Liu, W., He, W., Wu, J., Qin, N., He, Q., Xu, F., 2018. Residues, bioaccumulations and
biomagnification of perfluoroalkyl acids (PFAAs) in aquatic animals from Lake
Chaohu, China. Environ. Pollut. 240, 607–614. https://doi.org/10.1016/j.
envpol.2018.05.001.
Loi, E., Zhao, S., Wang, B., Zhu, L., Liang, T., Chen, M., Yang, L., Lv, J., Liu, L., 2011. Trophic
magnification of poly- and perfluorinated compounds in a subtropical food web. En-
viron. Sci. Technol. 45, 598–678. https://doi.org/10.1080/10643389.2011.627016.
Loureiro, S.N., Reis-Filho, J.A., Giarrizzo, T., 2016. Evidence for habitat-driven segregation
of an estuarine fish assemblage. J. Fish Biol. 89, 804–820. https://doi.org/10.1111/
jfb.13017.
Lunardi, V.O., Macedo, R.H., Granadeiro, J.P., Palmeirim, J.M., 2012. Migratory flows and
foraging habitat selection by shorebirds along the northeastern coast of Brazil: the
case of Baía de Todos os Santos. Estuar. Coast. Shelf Sci. 96, 179–187. https://doi.
org/10.1016/j.ecss.2011.11.001.
Mackay, D., Boethling, R.S., 2000. Bioconcentration and biomagnification in the aquatic
environment. Handbook of Property Estimation Methods for Chemicals. CRC Press,
pp. 211–254.
Magalhães, W.F., Barros, F., 2011. Structural and functional approaches to describe poly-
chaete assemblages: ecological implications for estuarine ecosystems. Mar. Freshw.
Res. 62, 918–926.
Mariano, D.L.S., Barros, F., 2015. Intertidal benthic macrofaunal assemblages: changes in
structure along entire tropical estuarine salinity gradients. J. Mar. Biol. Assoc.
United Kingdom 95, 5–15. https://doi.org/10.1017/S002531541400126X.
Martin, J.W., Mabury, S.A., Solomon, K.R., Muir, D.C.G., 2003. Bioconcentration and tissue
distribution of perfluorinated acids in rainbow trout (Oncorhynchus mykiss). Envi-
ron. Toxicol. Chem. 22, 196–204.
Martins, M.M., Mendonça, H.S., Rodrigues, S.S., Araújo, F.G., 2017. Trophic ecology of two
syntopic sciaenid species (Micropogonias furnieri (Desmarest, 1823) and
Ctenosciaena gracilicirrhus) (Metzelaar, 1919) in a tropical bay in south-eastern
Brazil. J. Appl. Ichthyol. 33, 740–745. https://doi.org/10.1111/jai.13360.
Matich, P., Ault, J.S., Boucek, R.E., Bryan, D.R., Gastrich, K.R., Harvey, C.L., Heithaus, M.R.,
Kiszka, J.J., Paz, V., Rehage, J.S., Rosenblatt, A.E., 2017. Ecological niche partitioning
within a large predator guild in a nutrient-limited estuary. Limnol. Oceanogr. 62,
934–953. https://doi.org/10.1002/lno.10477.
McCutchan Jr., J.H., Lewis Jr., W.M., Kendall, C., McGrath, C.C., 2003. Variation in trophic
shift for stable isotope ratios of carbon, nitrogen, and sulfur. Oikos 102, 378–390.
https://doi.org/10.1034/j.1600-0706.2003.12098.x.
MDIC, 2019.Bureau of Foreign Trade [WWWdocument]. URL. http://comexstat.mdic.gov.
br/pt/geral.
Minihan, S., 1983. Principles of biochemistry by Albert L Lehninger. pp 1011. 11. Worth
Publishers, New York, pp. 41–42. https://doi.org/10.1016/0307-4412(83)90020-1
1982. £13. ISBN 0-87901-136-X. Biochem. Educ.
(Ministério do Meio Ambiente) MMA, 2015a. Inventário Nacional de Novos Poluentes
Orgânicos Persistentes (Novos POPs) de uso industrial - Convenção de Estocolmo.
vol 166. Ministério do Meio Ambiente, Brasília.
(Ministério do Meio Ambiente) MMA, 2015b. Plano Nacional de Implementação da
Convenção de Estocolmo no Brasil. Ministério do Meio Ambiente.
Motta, P.N.D.S., Gloaguen, T.V., Couto, C.F., Merdy, P., Lucas, Y., 2018. Hydrogeochemistry
of the Subaé River Impacted by Lead Smelting Activities, Bahia State, Brazil: Geo-
chemical Characterization and Fluxes of Metals. https://doi.org/10.1007/s11270-
018-4032-6.
Munoz, G., Budzinski, H., Babut, M., Drouineau, H., Lauzent, M., Menach, K. Le, Lobry, J.,
Selleslagh, J., Simonnet-Laprade, C., Labadie, P., 2017a. Evidence for the trophic trans-
fer of perfluoroalkylated substances in a temperate macrotidal estuary. Environ. Sci.
Technol. 51, 8450–8459. https://doi.org/10.1021/acs.est.7b02399.
Munoz, G., Labadie, P., Botta, F., Lestremau, F., Lopez, B., Geneste, E., Pardon, P., Dévier, M.-
H., Budzinski, H., 2017b. Occurrence survey and spatial distribution of perfluoroalkyl
and polyfluoroalkyl surfactants in groundwater, surface water, and sediments from
tropical environments. Sci. Total Environ. 607–608, 243–252. https://doi.org/
10.1016/j.scitotenv.2017.06.146.
Munoz, G., Fechner, L.C., Geneste, E., Pardon, P., Budzinski, H., Labadie, P., 2018. Spatio-
temporal dynamics of per and polyfluoroalkyl substances (PFASs) and transfer to
periphytic biofilm in an urban river: case-study on the River Seine. Environ. Sci.
Pollut. Res. 25, 23574–23582. https://doi.org/10.1007/s11356-016-8051-9.
Nagamoto, N.S., Forti, L.C., Andrade, A.P.P., Boaretto, M.A.C., Wilcken, C.F., 2004. Method
for the evaluation of insecticidal activity over time in Atta sexdens rubropilosa
workers (Hymenoptera: Formicidae). SOCIOBIOLOGY 44, 413–431.
Nania, V., Pellegrini, G.E., Fabrizi, L., Sesta, G., De Sanctis, P., Lucchetti, D., Di Pasquale, M.,
Coni, E., 2009. Monitoring of perfluorinated compounds in edible fish from the
Mediterranean Sea. Food Chem. 115, 951–957. https://doi.org/10.1016/j.
foodchem.2009.01.016.
Nascimento, R.A., Nunoo, D.B.O., Bizkarguenaga, E., Schultes, L., Zabaleta, I., Benskin, J.P.,
Spanó, S., Leonel, J., 2018. Sulfluramid use in Brazilian agriculture: a source of per-
and polyfluoroalkyl substances (PFASs) to the environment. Environ. Pollut. 242,
1436–1443. https://doi.org/10.1016/J.ENVPOL.2018.07.122.
Olivero-Verbel, J., Tao, L., Johnston-Restrepo, B., Guette-Fernandez, J., Baldiris-Avila, R.,
O’byrne-Hoyos, I., Kannan, K., 2006. Perfluorooctanesulfonate and related
fluorochemicals in biological samples from the north coast of Colombia. Environ.
Pollut. 142, 367–372. https://doi.org/10.1016/j.envpol.2005.09.022.
Pérez, F., Llorca, M., Köck-Schulmeyer, M., Škrbić, B., Oliveira, L.S., da Boit Martinello, K.,
Al-Dhabi,N.A., Antić, I., Farré, M., Barceló,D., 2014. Assessment of perfluoroalkyl sub-
stances in food items at global scale. Environ. Res. 135, 181–189. https://doi.org/
10.1016/j.envres.2014.08.004.
Quinete, N., Wu, Q., Zhang, T., Yun, S.H., Moreira, I., Kannan, K., 2009. Specificprofiles of
perfluorinated compounds in surfaceand drinking waters and accumulation in mus-
sels, fish, and dolphins from southeastern Brazil. Chemosphere 77, 863–869. https://
doi.org/10.1016/j.chemosphere.2009.07.079.
Reis-Filho, J.A., Schmid, K., Harvey, E.S., Giarrizzo, T.,2019. Coastal fish assemblages reflect
marine habitat connectivity and ontogenetic shifts in an estuary-bay-continental
shelf gradient. Mar. Environ. Res. 148, 57–66. https://doi.org/10.1016/j.
marenvres.2019.05.004.
Ribeiro, L.F., Eça, G.F., Barros, F., Hatje, V., 2016. Impacts of shrimp farming cultivation cy-
cles on macrobenthic assemblages and chemistry of sediments. Environ. Pollut. 211,
307–315. https://doi.org/10.1016/j.envpol.2015.12.031.
Rossato,F.G.F.S., Susaeta,A., Adams, D.C., Hidalgo, I.G.,de Araujo, T.D., de Queiroz, A., 2018.
Comparison of revealed comparative advantage indexes with application to trade
tendencies of cellulose production from planted forests in Brazil, Canada, China,
Sweden, Finland and the United States. For. Policy Econ. 97, 59–66. https://doi.org/
10.1016/j.forpol.2018.09.007.
Routti, H., Aars, J., Fuglei, E., Hanssen, L., Lone, K., Polder, A., Pedersen, A.O., Tartu, S.,
Welker, J.M., Yoccoz, N.G., 2017. Emission changes dwarf the influence of feeding
habits on temporal trends of per- and polyfluoroalkyl substances in two Arctic top
predators. Environ. Sci. Technol. 51, 11996–12006. https://doi.org/10.1021/acs.
est.7b03585.
Sanganyado, E., Rajput, I.R., Liu, W., 2018. Bioaccumulation of organic pollutants in Indo-
Pacific humpback dolphin: a review on current knowledge and future prospects. En-
viron. Pollut. 237, 111–125. https://doi.org/10.1016/j.envpol.2018.01.055.
Schaider, L., Balan, S., Blum, A., Andrews, D., Strynar, M., Dickinson, M., Lunderberg, D.,
Lang, J., Peaslee, G., 2018. Fluorinated compounds in U.S. fast foodpackaging. Environ.
Sci. Technol. 4, 105–111. https://doi.org/10.1002/etc.3500.Molecular.
Schultes, L., Vestergren, R., Volkova, K., Westberg, E., Jacobson, T., Benskin, J.P., 2018. Per-
and polyfluoroalkyl substances and fluorine mass balance in cosmetic products from
the Swedish market: implications for environmental emissions and human exposure.
Environ Sci Process Impacts 20, 1680–1690. https://doi.org/10.1039/c8em00368h.
Schultes, L., Benskin, J.P., Majumdar, A., Peaslee, G.F., Brockman, J.D., McGuinness, S.R.,
Ngwenyama, R.A., Wilkinson, J.T., Sandblom, O., 2019a. Total fluorine measurements
in food packaging: how do current methods perform? Environ. Sci. Technol. Lett. 6,
73–78. https://doi.org/10.1021/acs.estlett.8b00700.
10 D.A. Miranda et al. / Science of the Total Environment 754 (2021) 142146
Schultes, L., Sandblom, O., Broeg, K., Bignert, A., Benskin, J., 2019b. Temporal trends
(1981–2013) of per- and polyfluoroalkyl substances and total fluorine in Baltic cod
(Gadus morhua). Environ. Toxicol. Chem. 39 (2), 300–309. https://doi.org/10.1002/
etc.4615.
Silvany, L., Senna, A.R., 2019. Three new species of Colomastix Grube, 1861 (Amphipoda:
Colomastigidea) from Todos-os-Santos Bay, northeastern Brazilian coast, with identi-
fication keys to Atlantic Ocean species. Zootaxa 4629 (4). https://doi.org/10.11646/
zootaxa.4629.4.3 9 Jul. 2019.
Simmonet-Laprade, C., Budzinski, H., Babut, M., Le Menach, K., Munoz, G., Lauzent, M.,
Ferrari, B.J.D., Labadie, P., 2019. Investigation of the spatial variability of poly- and
perfluoroalkyl substance trophic magnification in selected riverine ecosystems. Sci.
Total Environ. 686, 393–401. https://doi.org/10.1016/j.scitotenv.2019.05.461.
Simonnet-Laprade, C., Budzinski, H., Maciejewski, K., Le Menach, K., Santos, R., Alliot, F.,
Goutte, A., Labadie, P., 2019. Biomagnification of perfluoroalkyl acids (PFAAs) in the
food web of an urban river: assessment of the trophic transfer of targeted and un-
known precursors and implications. Environ Sci Process Impacts https://doi.org/
10.1039/C9EM00322C.
Soares, L.S.H., Lopez, J.P., Muto, E.Y., 2011. Giannini. Capture fishery in northern Todos os
Santos Bay, tropical southwestern Atlantic, Brazil. Brazilian J. Oceanogr, R.
Song, X., Vestergren, R., Shi, Y., Huang, J., Cai, Y., 2018. Emissions, transport, and fate of
emerging per- and polyfluoroalkyl substances from one of the major fluoropolymer
manufacturing facilities in China. Environ. Sci. Technol. https://doi.org/10.1021/acs.
est.7b06657.
Souza, M.M., Windmöller, C.C., Hatje, V., 2011. Shellfish from Todos os Santos Bay, Bahia,
Brazil: treat or threat? Mar. Pollut. Bull. 62, 2254–2263. https://doi.org/10.1016/j.
marpolbul.2011.07.010.
Souza, M.M., Windmöller, C.C., Hatje, V., 2014. Corrigendum to “shellfish from Todos os
Santos Bay, Bahia, Brazil: treat or threat?”[mar. Pollut. Bull. 62 (2011) 2254–2263].
Mar. Pollut. Bull. 88, 401. https://doi.org/10.1016/j.marpolbul.2014.07.027.
Souza, I. da C., Arrivabene, H.P., Craig, C.A., Midwood, A.J., Thornton, B., Matsumoto, S.T.,
Elliott, M., Wunderlin, D.A., Monferrán, M.V., Fernandes, M.N., 2018. Interrogating
pollution sources in a mangrove food web using multiple stable isotopes. Sci. Total
Environ. 640–641, 501–511. https://doi.org/10.1016/j.scitotenv.2018.05.302.
Stahl, T., Riebe, R.A., Falk, S., Failing, K., Brunn, H., 2013. Long-term lysimeterexperiment
to investigate the leaching of perfluoroalkyl substances (PFASs) and the carry-over
from soil to plants: results of a pilot study. J. Agric. Food Chem. 61, 1784–1793.
https://doi.org/10.1021/jf305003h.
Tavares, T.M., Beretta,M., Costa, M.C., 1999. Ratioof DDT/DDE in the all saints bay, Brazil
and its use in environmental management. Chemosphere 38, 1445–1452. https://doi.
org/10.1016/S0045-6535(98)00546-3.
Tomy, G.T., Budakowski, W., Halldorson, T., Helm, P.A., Stern, G.A., Friesen, K., Pepper, K.,
Tittlemier, S.A., Fisk, A.T., 2004. Fluorinated organic compounds in an eastern arctic
marine food web. Environ. Sci. Technol. 38, 6475–6481. https://doi.org/10.1021/
es049620g.
Tomy, G.T.,Pleskach, K., Ferguson, S.H.,Hare, J., Stern, G.,MacInnis, G., Marvin, C.H., Loseto,
L., 2009. Trophodynamics of some PFCs and BFRs in a Western Canadian Arctic ma-
rine food web. Environ. Sci. Technol. 43, 4076–4081. https://doi.org/10.1021/
es900162n.
UNSCPOPs, 2009. No Title [WWW Document]. Stock. Conv. Persistent Org. Pollut URL.
http://chm.pops.int/.
USEPA, 2002. Perfluoroalkyl Sulfonates; Significant New Use Rule (SNUR) (Washington,
DC).
Vidal, A., Lafay, F., Daniele, G., Vulliet, E., Rochard, E., Garric,J., Babut, M., 2019. Does water
temperature influence the distribution and elimination of perfluorinated substances
in rainbow trout (Oncorhynchus mykiss)? Environ. Sci. Pollut. Res. 26,
16355–16365. https://doi.org/10.1007/s11356-019-05029-w.
Vinagre, C., Mendonça, V., Flores,A.A.V., Baeta, A., Marques, J.C., 2018. Complex food webs
of tropical intertidal rocky shores (SE Brazil) –an isotopic perspective. Ecol. Indic. 95,
485–491. https://doi.org/10.1016/j.ecolind.2018.07.065.
Wang, Z., Dewitt, J.C., Higgins, C.P., Cousins, I.T., 2017. A never-ending story of per- and
polyfluoroalkyl substances (PFASs)? Environ. Sci. Technol. 51, 2508–2518. https://
doi.org/10.1021/acs.est.6b04806.
Wang, Y., Zhong, Y., Li, J., Zhang, J., Lyu, B., Zhao, Y., Wu, Y., 2018. Occurrence of
perfluoroalkyl substances in matched human serum, urine, hair and nail. J. Environ.
Sci. 67, 191–197. https://doi.org/10.1016/j.jes.2017.08.017.
Wellens, S., Sandrini-Neto, L., González-Wangüemert, M., Lana, P., 2015. Do the crabs
Goniopsis cruentata and Ucides cordatus compete for mangrove propagules? A
field-based experimental approach. Hydrobiologia 757, 117–128. https://doi.org/
10.1007/s10750-015-2245-x.
Xu, J., Guo, C.-S., Zhang, Y., Meng, W., 2014. Bioaccumulation and trophic transfer of
perfluorinated compounds in a eutrophic freshwater food web. Environ. Pollut. 184,
254–261. https://doi.org/10.1016/j.envpol.2013.09.011.
Yin, T., Te, S.H., Reinhard, M., Yang, Y., Chen, H., He, Y., Gin, K.Y.-H., 2018. Biotransforma-
tion of Sulfluramid (N-ethyl perfluorooctane sulfonamide) and dynamics of associ-
ated rhizospheric microbial community in microcosms of wetland plants.
Chemosphere 211, 379–389. https://doi.org/10.1016/j.chemosphere.2018.07.157.
Zabaleta, I., Bizkarguenaga, E., Nunoo, D.B.O., Schultes, L., Leonel, J., Prieto, A., Zuloaga, O.,
Benskin, J.P., 2018. Biodegradation and uptake of the pesticide sulfluramid in a soil-
carrot Mesocosm. Environ. Sci. Technol. 52, 2603–2611. https://doi.org/10.1021/acs.
est.7b03876.
Zhao, S., Zhou, T., Zhu, L., Wang, B., Li, Z., Yang, L., Liu, L., 2018. Uptake, translocation and
biotransformation of N-ethyl perfluorooctanesulfonamide (N-EtFOSA) by hydropon-
ically grown plants. Environ. Pollut. 235, 404–410. https://doi.org/10.1016/j.
envpol.2017.12.053.
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