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Environmental Pollution 265 (2020) 114923
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Environmental Pollution
journal homepage: www.elsevier .com/locate/envpol
Contaminant screening and tissue distribution in the critically
endangered Brazilian guitarfish Pseudobatos horkelii*
Mariana F. Martins a, *, Patrícia G. Costa a, Adalto Bianchini b
a Programa de P�os-Graduaç~ao em Ciências Fisiol�ogicas, Instituto de Ciências Biol�ogicas, Universidade Federal do Rio Grande-FURG, Av It�alia, Km 8 96203-
900, Rio Grande, Brazil
b Instituto de Ciências Biol�ogicas, Universidade Federal do Rio Grande-FURG, Av It�alia, Km 8 96203-900, Rio Grande, Brazil
a r t i c l e i n f o
Article history:
Received 11 March 2020
Received in revised form
4 May 2020
Accepted 31 May 2020
Available online 1 June 2020
Keywords:
Elasmobranch
Emerging contaminants
Pollution
Polycyclic aromatic hydrocarbons
Southwest Atlantic
* This paper has been recommended for acceptanc
* Corresponding author. Programa de P�os-Graduaç
Instituto de Ciências Biol�ogicas, Universidade Feder
Grande, Brazil.
E-mail address: marianadafmartins@gmail.com (M
https://doi.org/10.1016/j.envpol.2020.114923
0269-7491/© 2020 Elsevier Ltd. All rights reserved.
a b s t r a c t
Elasmobranchs are particularly prone to accumulating contaminants due to their life history patterns and
relatively high trophic position. However, several compounds, especially contaminants of emerging
concern, have still not been well studied in this group. Here, we aimed to determine the occurrence and
concentrations of several inorganic and organic contaminants in different tissues of the Brazilian gui-
tarfish Pseudobatos horkelii. This species is a critically endangered species, endemic from the Southwest
Atlantic which uses southern Brazilian waters as a nursery habitat. Polycyclic aromatic hydrocarbons
(PAHs), emerging pesticides, pharmaceutical and personal care products (PPCPs) and trace metals were
determined in five biological tissues in order to assess the accumulation and organotropism of these
compounds. Except for chlorothalonil and triclosan, all compounds were detected in, at least, one tissue,
mostly in liver samples. All compounds differed among tissues, with liver presenting the higher con-
centrations of several contaminants, followed by muscle and gills. PAHs and PPCPs were the most
detected analytes and presented the highest concentrations among tissues. Diclofenac levels were
determined, for the first time in elasmobranchs, and were relatively high, when compared to other
fishes. Finally, relatively high concentrations of PAHs, dichlofluanid and octocrylene in muscle might be
suggestive of chronic exposure, presenting also human health implications. Regarding trace metals,
contrary to most elasmobranch studies, Hg levels were low in all tissues, whereas Cd and Pb here higher
in liver, and gills and blood samples, respectively. Our results indicate that P. horkelii is exposed to several
organic and inorganic which might affect this species in a long-term scale. Concerning the determination
of emerging contaminants, it is likely that other elasmobranchs are also exposed to these compounds and
special attention should be given to this issue in order to predict future effects on this group.
© 2020 Elsevier Ltd. All rights reserved.
1. Introduction
For the past decades, several compounds were considered as
major contaminants in coastal regions due to their persistency and
potential deleterious effects on organisms (Fleeger et al., 2003). On
the other hand, a variety of newly synthesized compounds,
considered as of emerging concern (CE), such as pharmaceuticals
and personal care products (PPCPs), has been recently determined
in marine ecosystems (Arpin-Pont et al., 2014). These compounds
e by Sarah Harmon.
~ao em Ciências Fisiol�ogicas,
al do Rio Grande-FURG, Rio
.F. Martins).
lack of monitoring and are potentially harmful to aquatic envi-
ronments and wildlife (Sauv�e and Desrosiers, 2014; Zenker et al.,
2014). Whereas trace metals and legacy contaminants are persis-
tent in the environment and prone to accumulate in marine wild-
life, CEs are generally considered pseudo-persistent due to their
indiscriminate use and chronic release in aquatic systems (Boxall
et al., 2012; Overturf et al., 2015), despite their relatively rapid
degradation. Yet, little is known on their occurrence and potential
to accumulate in marine species. Impacts of CEs have already been
hypothesized to be associated with population declines in birds
(Oaks et al., 2004), and femininization of freshwater fishes with
consequences at a population-level (Kidd et al., 2007), but studies
conducted on marine species are mostly limited by experimental
designs on traditional model species (Dann and Hontela, 2010;
Fabbri and Franzellitti, 2016).
mailto:marianadafmartins@gmail.com
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M.F. Martins et al. / Environmental Pollution 265 (2020) 1149232
Elasmobranchs tend to accumulate high levels of contaminants
due to their higher trophic positions, acting as meso and apex
predators, and life history parameters, such as longevity. Accumu-
lation of legacy contaminants and tracemetals has been extensively
studied among this group (Gelsleichter andWalker, 2010), whereas
studies evaluating CEs are incipient, with only a few compounds
being analyzed (Gelsleichter and Szabo, 2013; Lyons et al., 2018;
Nakata, 2005; Nakata et al., 2009; Xue and Kannan, 2016; Xue et al.,
2017). However, due to their trophic proximity, as well as k-strategy
life-history patterns, with other taxa in which PPCPs have been
detected (Gago-Ferrero et al., 2013; Fair et al., 2009; Nakata, 2005),
it is highly expected to also detect these compounds and its me-
tabolites in elasmobranch tissues. Regarding tissue distribution, a
few studies have evaluated the distribution of contaminants (De
Boeck et al., 2010; Corsolini et al., 2014), but are still scarce, espe-
cially considering CEs. For this reason, studies analyzing organo-
tropism of contaminants are determinant for understanding the
kinetics and associated physiological impacts of environmental
contamination in elasmobranchs.
The Brazilian guitarfish Pseudobatos horkelii is a bottom-dweller
endemic species from the Southwestern Atlantic occurring from
southeastern Brazil to Argentina (Menni and Stehmann, 2000). This
species shows a synchronous reproductive cycle, with pregnant
females approaching shallow waters during the summer for em-
bryonic development and parturition (Lessa et al., 1986; Martins
et al., 2018). Southern Brazil represents a nursery area for south-
ern populations, which are considered resident (Vooren et al.,
2005). This species was categorized as “Critically Endangered” by
the IUCN Red List of Threatened Species (Lessa and Vooren, 2016)
and landing and commercialization are, therefore, prohibited by
the Brazilian legislation (IBAMA, 2004). Despite this, P. horkelii is
still illegally traded (De-Franco et al., 2012; Bunholi et al., 2018) and
consumed, especially in southern Brazil.
Urban, industrial, agricultural and harbor activities have been
impacting the southern Brazilian estuaries, mostly from inland
sources (Mirlean et al., 2003; Wallner-Kersanach et al., 2016). The
Patos Lagoon is the largest water body from South America and
receives most of freshwater systems input, which are possible
sources of contamination (Amado et al., 2006). This input of con-
taminants ends up in the marine environment, especially through
the Rio Grande Channel. Caldas et al. (2019) detected several
emerging compounds in surface and even drinking water from this
area, suggesting that aquatic organisms might be exposed to
emerging pesticides and PPCPs. Moreover, livestock and agricul-
tural activities might also impactorganisms in southern wetlands
(Quintela et al., 2019), which run off to the coast. Considering this, it
is likely that agricultural runoff, added to the already mentioned
activities might also impact marine ecosystems and organisms,
especially those inhabiting shallow waters.
In the light of the above, we hypothesize that P. horkeliimight be
exposed to environmental contamination of anthropogenic sources
in southern Brazil. Furthermore, considering that its meat is ille-
gally consumed, contaminant levels for this species are of great
interest regarding human health. Taking this into account, we
aimed to characterize, for the first time, the levels and tissue dis-
tribution of organic and inorganic compounds in five tissues of the
Brazilian guitarfish P. horkelii, sampled in southern Brazil.
2. Materials and methods
2.1. Study site
The Patos Lagoon is located in Rio Grande do Sul State, southern
Brazil and has been receiving an input of contaminants from
anthropogenic sources for about 15 years (Wallner-Kersanach et al.,
2016). The southernmost estuarine region has been particularly
affected due to an increase in port and naval activities, mostly in Rio
Grande city (209,378 inhabitants, IBGE, 2017), where one of the
largest harbors in southern Brazil is located. In addition, urban,
industrial (fertilizer-producing plants, oil refineries and fishing
industries), agriculture, and livestock activities, contribute to the
contamination of this estuary (Mirlean et al., 2003) and conse-
quently the coastal zone through the Rio Grande Channel. This
channel connects the estuarine complex with the Atlantic Ocean, at
Praia do Cassino, situated between Rio Grande and Chuí (31,274
habitants; IBGE, 2017) cities, Rio Grande do Sul State, Southern
Brazil (32�0200600S 52�0505500W).
2.2. Sample collection and animal handling
Eighteen pregnant females were opportunistically obtained
from fishermen at Praia do Cassino (80 Km southern from Rio
Grande), from December to February/2019. Total length ranged
from 102 to 132 cm (mean ¼ 118.0, s.d. ¼ 9.5) and total body mass
ranged from 3990 to 9780 g (mean ¼ 6618.3, s.d. ¼ 2003.8). Liver,
gonad, muscle, gills and blood samples were obtained from each
specimen and were kept under �80 �C until analysis. All proced-
ures were previously authorized (SI).
2.3. Trace metals determination
Sample preparation and determination of Cd, Cr, Cu, Fe, and Pb
followed the protocol sensu Abril et al. (2018). Subsamples of 1 mL
of blood, and 0.2 g from other tissues were dried for 48 h at 60 �C
and completely digested in 500 mL of 65% HNO3 (Suprapur, Merck,
Darmstad, Germany) for 24 h at the same temperature. Samples
had the final volume adjusted with high ultrapure water (Master
System MS-2000, Gehaka, Brazil) (resistivity of 18.2 MU/cm) at
1 mL for further dilutions (10 times for Hg and 5 times for other
metals). Determination of the previously mentioned trace metals
was done using a High-Resolution Continuum Source Graphite
Furnace Atomic Absorption Spectrometer (HR-CS GF AAS, Analitik
Jena, Jena, Germany) and mercury analysis were carried out using
an atomic fluorescence spectrometer Mercur Duo Plus (Analytik
Jena, Jena, Germany).
2.4. Organic compounds determination
All standards were purchased from Sigma-Aldrich (St. Louis,
MO, USA). Subsamples of 0.5 g from gills, liver, muscle and gonad,
were homogenized with anhydrous sodium sulfate and spikedwith
100 mL atrazine-d5 and p-Terphenyl-D14 surrogate standard for
emerging contaminants and Polycyclic Aromatic Hydrocarbons
(PAHs) analysis, respectively. Samples were further Soxhlet
extracted with 1:1 dichloromethane and n-hexane (Merck, Darm-
stadt, Germany) for 12 h. Whole blood samples were extracted by
solid extraction phase (Camacho et al., 2014). Blood samples were
applied after cartridges were activated with 3 mL of methanol
(Merck, Darmstadt, Germany) followed by three ml of high ultra-
purewater, at a rate of 1mL/min. Oneml of samplewas then passed
through the cartridge by gravity flow and the vials were further
rinsed with three aliquots of 1 mL high ultrapure water and dried
under vacuum for 15 min. After that, samples were eluted with two
aliquots of 2 mL dichloromethane, which was gently dried under
vacuum and eluted in hexane for further chromatography analysis.
The final extracts were evaporated with Nitrogen to 10 mL and
100 mL was used for gravimetric lipid determination. The remaining
extracts were evaporated to 6 mL, being further divided into three
fractions of 2 mL each for emerging contaminants, chlorinated
compounds (PCBs and pesticides, not analyzed herein) and PAHs
Table 1
Descriptive statistics of the concentrations of organic contaminants analyzed in five
tissues of female guitarfishes Pseudobatos horkelii from southern Brazil.
N Min. Max. Median Mean S.D.
Atrazine
Blood 14 0.00 0.00 0.00 0.00 0.00
Gills 17 0.00 0.00 0.00 0.00 0.00
Liver 18 0.00 31.41 9.32 11.70 10.12
Muscle 17 0.00 8.18 0.00 0.48 1.98
Ovaries 15 0.00 8.71 0.00 1.60 3.34
Chlorpyrifos
Blood 14 0.00 0.00 0.00 0.00 0.00
Gills 16 0.00 14.93 2.99 3.82 4.64
Liver 18 0.00 13.20 0.00 1.84 4.29
Muscle 17 0.00 4.93 0.00 1.18 1.92
Ovaries 15 0.00 5.23 0.00 0.82 1.74
Dichlofluanid
Blood 14 0.00 0.00 0.00 0.00 0.00
Gills 16 0.00 19.26 10.07 10.51 5.84
Liver 17 0.00 9.91 2.00 3.08 3.38
Muscle 17 0.00 14.51 8.08 8.21 3.63
Ovaries 14 0.00 21.15 5.98 7.42 5.70
Diclofenac
Blood 14 0.00 384.67 48.85 79.15 106.97
Gills 17 0.00 738.69 391.67 410.93 189.43
Liver 17 133.93 4469.22 926.89 1474.25 1466.71
Muscle 17 0.00 262.20 0.00 83.20 105.90
Ovaries 15 130.26 699.32 490.22 410.47 169.72
Diuron
Blood 14 0.00 0.32 0.00 0.03 0.09
Gills 17 0.00 1.99 0.00 0.16 0.50
Liver 18 0.00 0.00 0.00 0.00 0.00
Muscle 17 0.00 0.85 0.00 0.05 0.21
Ovaries 15 0.00 1.81 0.00 0.12 0.47
Methylparaben
Blood 14 0.00 0.00 0.00 0.00 0.00
Gills 17 0.00 42.35 0.00 2.49 10.27
Liver 16 0.00 235.80 50.16 87.85 97.50
Muscle 17 0.00 0.00 0.00 0.00 0.00
Ovaries 15 0.00 91.12 0.00 6.07 23.53
Octocrylene
Blood 14 0.00 0.00 0.00 0.00 0.00
Gills 17 0.00 35.38 19.65 18.97 10.77
Liver 17 0.00 51.35 9.64 14.80 15.60
Muscle 17 3.64 35.51 15.40 15.35 9.77
Ovaries 14 0.00 20.65 7.71 7.53 6.62
S Polycyclic Aromatic Hydrocarbons
Blood 14 304.82 2443.95 684.22 833.42 569.69
Gills 16 124.47 8487.22 1138.31 1564.74 1888.77
Liver 18 723.39 3530.94 1289.82 1452.80 710.34
Muscle 17 852.78 3651.77 1973.90 2134.82 847.99
Ovaries 14 589.92 2274.28 1184.53 1286.23 472.41
Trifluralin
Blood 14 0.00 0.00 0.00 0.00 0.00
Gills 17 0.00 27.57 0.00 1.62 6.69
Liver 18 0.00 0.00 0.00 0.00 0.00
Muscle 17 0.00 30.11 0.00 2.00 7.30
Ovaries 15 0.00 0.00 0.00 0.00 0.00
Sample size (N), minimum and maximum observed values (Min. and Max.,
respectively), median, mean, and standard deviation (S.D.) are expressed as ng g�1
wet weight.
M.F. Martins et al. / Environmental Pollution 265 (2020) 114923 3
determination, respectively. PAHs were fractionated by liquid
chromatography absorption in a silica-gel (6 g) (Merck, Darmstadt,
Germany), and neutral aluminum oxide (8 g) (Merck, Darmstadt,
Germany) column. Serial elution with 25 mL of n-hexane, and
30 mL of n-hexane/dichloromethane (9:1) followed by 25 mL of n-
hexane/dichloromethane (1:1) was used.
All organic compounds were determined by gas chromatog-
raphy (Shimadzu GC-2010 Plus, Shimadzu Corporation, Kyoto,
Japan) coupled to mass spectrometry (Shimadzu GCMS-QP 2020;
Shimadzu Corporation, Kyoto, Japan) selected in the ionmonitoring
mode. The analytes analyzed were: 18 PAHs were determined,
including low molecular weight (LMW-naphthalene, 2-
methylnaphthalene, 1-methylnaphthalene, acenaphthylene, ace-
naphthene, fluorene, phenanthrene, anthracene), and high molec-
ular weight (HMW- fluoranthene, pyrene, benzo[a]anthracene,
chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]
pyrene, indeno[1,2,3-cd]pyrene, dibenzo[a,h]anthracene, and
benzo[g,h,i]perylene); emerging pesticides (atrazine, chlor-
othalonil, chlorpyrifos, dichlofluanid, diuron, trifluralin)and PPCPs
(diclofenac, methylparaben, octocrylene, and triclosan).
Selection of analytes considered their occurrence in the study
site and surrounding areas (Garcia et al., 2010; Mirlean et al., 2003;
Caldas et al., 2019), and inwildlife tissues (Gago-Ferrero et al., 2013;
Da Silva et al., 2014; Kehrig et al., 2015), as well as their occurrence
in a global perspective (Arpin-Pont et al., 2014). Physico-chemical
proprieties of the emerging analytes and their class and usage are
provided in Table S1. Quality control and assurance included blanks
spiked with surrogate standard, which were analyzed simulta-
neously, with one blank for each group of tissue sample. Surrogate
recovery ranged from 75 to 91%. For trace metals, analytical control
was made using a certified reference material (DORM-4, National
Research Council, Canada) and recovery data was: 88.02% (Cd);
84.59% (Cr); 87.93% (Cu); 88.79% (Fe); 91.39% (Hg); and 86.94% (Pb).
2.5. Data analysis
Results are expressed as ng g�1 wet weight (w.w.) for organic
contaminants, and as mg kg�1 dry weight (d.w.) for trace metals.
Lipid percentage is provided in Table S2. Values of zero were
attributed to concentrations below detection limits (<LOD) (Cullen
et al., 2019) and were incorporated in statistical analysis. If con-
centrations of a compound were zero in all samples, this variable
was not considered in any statistical analysis.
Differences between tissues for each contaminant were tested
with the Kruskal-Wallis analysis of variance by ranks, followed by
Dunn’s Test with Bonferroni Adjustment (Table S2 and S3-Zar,
2010). Non-parametric testes were chosen due to the non-
normality distribution of data. Principal Component Analysis
(PCA) were performed for standardized trace metals and organic
compounds separately in order to explore relationships between
tissues through contamination profiles (Legendre and Legendre,
1998). Variables chosen for PCA were those with more than 40%
of detection frequency. All statistical analysis were conducted using
R (R CoreTeam, 2018). Significance level adopted was of 5%.
3. Results and discussion
Descriptive statistics of concentrations of all organic contami-
nants in each tissue analyzed are provided in Table 1 and detection
frequencies are provided in Table S4.
3.1. Polycyclic Aromatic Hydrocarbons
The sum of PAHs levels (SPAHs) (Fig. 1) differed among tissues,
with muscle and liver presenting the highest mean concentrations
(2134.8, 1452.8 ng g�1 w.w.), whereas blood had the lowest mean
concentration among tissues (833.421 ng g�1 w.w.). The levels
observed here are similar to those reported for liver of carcharhinid
sharks (Cullen et al., 2019), and higher than reported for other
Fig. 1. (A) Concentrations of the sum of polycyclic aromatic hydrocarbons (PAHs) for
each tissue analyzed in Pseudobatos horkelii and possible differences tested by
Kruskal-Wallis followed by Dunn’s Test (indicated by lowercase letters, where same
letter indicate the lack of difference, i.e. p > 0.05). (B) Proportion of 18 PAHs to the sum
of all PAHs measured for each tissue analyzed.
M.F. Martins et al. / Environmental Pollution 265 (2020) 1149234
marine fish (Alani et al., 2012). Furthermore, P. horkelii presented
higher concentrations of SPAHs when compared to higher trophic
position taxa, such as spermwhales (Physeter macrocephalus) (Zhan
et al., 2019). PAHs are mainly accumulated through water and food
exposure (Lee et al., 1972) explaining the significant concentrations
observed for gills. However, trophic exposure might also play a
crucial role in the high levels observed, since guitarfishes feed on
benthic prey (Bornatowski et al., 2010), which, in some cases, are
less effective in metabolizing PAHs than other organisms (Hylland,
2006). In addition, adjacent port and estuary regions close to
P. horkelli’s geographic distribution are known to be contaminated
with PAHs (Medeiros et al., 2005; Garcia et al., 2010), explaining the
relatively high levels found herein.
PAHs tend to accumulate in lipid-rich tissues (Logan, 2007),
differently from our observations. Mashroofeh et al. (2015) found
relatively low SPAH concentrations in muscles compared to other
tissues. In this study, however, muscle samples had atypical higher
concentrations among tissues. Since muscle concentrations are
indicative of chronic exposure (Daley et al., 2014), our results sug-
gest that these guitarfishes might be chronically exposed to these
contaminants in the sampling area. Furthermore, considering that
P. horkelii is illegally traded for human consumption (Bunholi et al.,
2018), risks to human health due to meat contamination might be
considered.
Congener profile among tissues was similar (Fig. 1) and low
molecular weight (LMW) PAHs were more abundant in all tissues
(52.0e78.6%), with 2-methylnaphtalene presenting the higher
mean concentrations among congeners (12.1e25.3%). This preva-
lence is a common feature among fishes, possibly due to a more
effective metabolism of high molecular weight (HMW) congeners
rather than LMW ones, as observed for fishes (Mashroofeh et al.,
2015; Marsili et al., 2016; Jafarabadi et al., 2019). High molecular
weight PAHs, on the other hand, were less abundant, except in
muscle samples, where they contributed with 48% of SPAH, with
pyrene being the predominant HMW congener. However, due to
their potential toxicity (Henner et al., 1997), the low levels of HMW
PAHs should not exclude possible deleterious effects to this species.
3.2. Emerging contaminants
Differences between tissues were observed for chlorpyrifos,
dichlofluanid, diclofenac, and octocrylene (Fig. 2). Atrazine
(11.696 ng g�1 w.w.), diclofenac (1474.251 ng g�1 w.w.), and
methylparaben (87.850 ng g�1 w.w.) were predominant in liver
samples, possibly due to this organ’s functional role in xenobiotics
metabolism (Van der Oost et al., 2003; Miller et al., 2018) and also
due to its high lipid content (Table S1). However, studies analyzing
emerging contaminants in wildlife mostly analyze liver samples
(Arpin-Pont et al., 2014) and studies on the organotropism of these
compounds are not available.
Liver methylparaben mean concentration (87.850 ng g�1 w.w.)
observed for P. horkelii was higher than those observed for coastal
fishes, the Atlantic sharpnose shark Rhizoprionodon terranovae
(13e71 ng g�1 w.w.) and even black-footed albatrosses (Phoebastria
nigripes) (9.55e10.3 ng g�1 w.w.) (Xue and Kannan, 2016). Despite
the relatively high levels observed, metabolites of methylparaben
might be more prone to accumulate and to consequently distribute
among tissues, explaining the low detection frequency of this
compound in other tissues. The same explanation can be applied
for the low detection of triclosan, in which it’s metabolites might
accumulate at higher rates than intact triclosan. In this study,
however, the null detection of triclosan in continental superficial
waters (Caldas et al., 2019) is suggestive of little or no exposure of
P. horkelii to this compound.
Octocrylenewas detected at a frequency of 71.4e100% (Table S2)
in all tissues except blood, and concentrations were lower than
those reported for Franciscana dolphins (Pontoporia blainvillei)
from southern Brazil (Gago-Ferrero et al., 2013). Tissue distribution
also differed from the observed for the lebranchemullet (Mugil liza)
(Molins-Delgado et al., 2018) in southeastern Brazil. Whereas
muscle contained the higher concentrations of octocrylene in
mullet, no differences were observed among liver, muscle and gill
samples for P. horkelii. Differences in habitat might drive the
observed disparities, as the fish M. liza is an estuarine species and
P. horkelii is an exclusive marine species. Considering the differ-
ences between the results from Gago-Ferrero et al. (2013) and the
present ones, migratory capacities might be also taken into account,
since P. horkelii is considered a resident species in southern Brazil
(Vooren et al., 2005).
Our study is the first to report diclofenac accumulationin
elasmobranchs. Here, diclofenac had the highest levels and fre-
quency rates among the emerging contaminants analyzed, ranging
Fig. 2. Concentrations of (A) chlorpyrifos, (B) dichlofluanid, (C) diclofenac, and (D) octocrylene for each tissue analyzed for Pseudobatos horkelii. Differences tested by Dunn’s Test
are indicated by different lowercase letters (p < 0.05).
M.F. Martins et al. / Environmental Pollution 265 (2020) 114923 5
from 83.185 to 1471.251 ng g�1 w.w. Ojemaye and Petrik (2019)
found similar diclofenac levels for south African fishes, whereas
most of the diclofenac studies inmarine environments report lower
concentrations (Liu et al., 2015; Omar et al., 2019). The high
detection frequency in all tissues might be linked to diclofenac’s
ability to cross biological membranes, as other pharmaceuticals
(Miller et al., 2018). The tissue’s distribution followed the observed
by Schwaiger et al. (2004) with liver presenting the highest values,
followed by gills and, finally, muscle. However, contrarily to this
study, the authors did not analyzed gonads.
Diclofenac as other pharmaceuticals, is still scarcely determined
in wildlife (Bonnefille et al., 2017; Miller et al., 2018), despite being
detected even in remote areas such as Arctic (Gonzales-Alonso
et al., 2017). Indeed, 75% of diclofenac used in either domestic or
veterinary compartments enters the environment (He et al., 2017).
In marine compartments, diclofenac has been detected up to
10.2 mg L�1 in seawater (Ali et al., 2018) and 13.8 ng g�1 in sediment
(Omar et al., 2019). In southeastern Brazil, environmental levels
were lower than in previous studies, with concentrations of 19.4 ng
L1 and 1.06 ng g�1 for seawater and sediment, respectively (Pereira
et al., 2016; Beretta et al., 2014). Despite the lack of studies
analyzing the presence of diclofenac in the sampling area, both
sources might be responsible for the high levels observed in all
tissues of P. horkelii, since the studied area is situated among three
municipalities where livestock activities are intense. In addition,
the high detection of diclofenac in all tissues analyzed herein,
including less metabolic tissues such as muscle and gonads, sug-
gests that the organism might not be efficiently metabolizing and
excreting this compound, especially considering that organisms are
chronically exposed to this pharmaceutical due to its continuous
input in the environment. Considering the high levels found for
P. horkelii, physiological implications as a result from this chronic
exposure might be a concern, since diclofenac has been associated
with oxidative damage, cytotoxicity, and genotoxicity
(Sathishkumar et al., 2020). In addition, these results are important
for monitoring issues since diclofenac is one of the compounds
listed in the watch-list of the European Union (European Union,
2013).
Despite its worldwide use as antifouling biocide and weed
control, as well as its occurrence in areas of high boating activity
(Konstantinou and Albanis, 2004), diuron was only detected in a
few samples (5.9e14.3%) and at low concentrations (0.131e0.159 ng
g��1 w.w.), whereas chlorothalonil was not detected in any tissue.
Low or null concentrations of these compounds observed herein
could be explained by the low usage of these compounds in the
study area . However, Caldas et al. (2019) detected diuron in surface
water of continental compartments, indicating that diuron con-
centrations, if inputted to the marine environment, might be
diluted. Atrazine, on the other hand, is one of the most used active
compounds of pesticides in Rio Grande do Sul State and had the
highest concentration among emerging biocides analyzed
(0.481e11.696 ng g�1 w.w.), but dichlofluanid (3.084e10.505 ng g�1
w.w.) was the predominant biocide (0.0e94.1%) followed by
chlorpyrifos (0.0e62.5%). The prevalence of dichlofluanid in gills
and muscles, differently from atrazine, which was more abundant
in liver samples, suggests that biocides do not follow a distribution
pattern in P. horkelii. In fact, the null detection of atrazine in gills
suggests that this compound might be uptaken by trophic transfer
rather than waterborne exposure. In contrast, the high levels of
dichlofluanid found in gill samples are indicative of waterborne
exposure, especially considering that the gills are the main
absorbing organ in fishes (Barron, 2003).
M.F. Martins et al. / Environmental Pollution 265 (2020) 1149236
3.3. Inorganic contaminants
Descriptive statistics of trace metals concentrations per tissue
are provided in Table 2.
Trace metal levels differed among tissues, with liver, gills and
muscle presenting the highest concentrations, except for Fe.
Detection frequency was of 100% for all tissues except for Cd, where
47.6% of the samples were <LOQ.
Mercury is the most studied metal in elasmobranchs, usually
observed at higher concentrations in liver and muscle (Gelsleichter
and Walker, 2010; Bezerra et al., 2019). For P. horkelii, however, Cd
and Pb were observed at higher concentrations in liver, and gills
and blood, respectively, whereas Hg levels were relatively low
(0.01e0.037 mg kg�1 d.w.) (Fig. 3). Cadmium accumulates prefer-
entially in liver, as observed for the silky shark Carcharhinus falci-
formis (Terrazas-L�opez et al., 2016), explaining differences in tissue
detection frequency and the higher mean levels of this metal
(0.226 mg kg�1 d.w.), comparable to apex predators. As also
observed for the smoothtooth blacktip shark Carcharhinus leiodon,
Cd levels in muscle were predominantly below detection limits
(Moore et al., 2015), whereas liver samples presented the higher
Table 2
Descriptive statistics of the concentrations of trace metals analyzed in five tissues of
female guitarfishes Pseudobatos horkelii from southern Brazil.
N Min. Max. Median Mean S.D.
Cd
Blood 12 0.00 0.00 0.00 0.00 0.00
Gills 15 0.00 0.02 0.00 0.01 0.01
Liver 13 0.03 1.01 0.15 0.23 0.29
Muscle 17 0.00 0.00 0.00 0.00 0.00
Ovaries 12 0.02 0.05 0.02 0.03 0.01
Cr
Blood 12 0.00 0.15 0.03 0.03 0.04
Gills 14 0.12 0.92 0.39 0.43 0.22
Liver 12 0.02 0.09 0.03 0.04 0.02
Muscle 17 0.07 1.34 0.57 0.61 0.36
Ovaries 12 0.01 0.08 0.02 0.03 0.02
Cu
Blood 12 0.68 1.48 1.11 1.09 0.23
Gills 15 1.34 2.25 1.78 1.79 0.26
Liver 13 1.05 9.75 1.91 2.65 2.37
Muscle 17 0.11 0.98 0.25 0.30 0.22
Ovaries 12 1.40 3.77 1.85 1.96 0.61
Fe
Blood 12 757.43 1038.64 848.91 853.26 76.85
Gills 15 125.62 348.14 259.24 229.94 74.67
Liver 13 26.02 221.54 117.98 113.04 61.73
Muscle 17 26.01 94.82 39.63 46.99 19.58
Ovaries 12 2.32 51.47 18.62 22.14 15.49
Hg
Blood 12 0.00 0.01 0.00 0.00 0.00
Gills 15 0.01 0.17 0.02 0.04 0.04
Liver 13 0.00 0.03 0.00 0.01 0.01
Muscle 17 0.00 0.11 0.02 0.03 0.03
Ovaries 12 0.00 0.00 0.00 0.00 0.00
Pb
Blood 15 0.08 0.16 0.12 0.12 0.02
Gills 15 0.05 0.14 0.07 0.08 0.03
Liver 13 0.01 0.10 0.01 0.02 0.02
Muscle 17 0.00 0.04 0.01 0.01 0.01
Ovaries 12 0.01 0.04 0.01 0.01 0.01
Sample size (N), minimum and maximum observed values (Min. and Max.,
respectively), mean, standard deviation (S.D.) and median are expressed as mg kg�1
dry weight.
levels of this metal. Such differences might be a result of differences
in the availability of reactive groups responsible for the binding of
trace metals to organic molecules (e.g. metallothionein) (Storelli
et al., 2011), which are more abundant in liver and kidney.
Lead levels, on the other hand, were higher in gills (0.078 mg
kg�1 d.w.) and blood (0.121 mg kg�1 d.w.) samples, possibly
because gills are the main organ exposed to the environment
(Lopes et al., 2019). In addition, Pd tends to bind to erythrocytes
(Wood et al., 2012), explaining the high relative concentrations in
this tissue. Pb accumulation occurs mostly for species occurring in
highly urbanized and therefore impacted areas (De Boeck et al.,
2010), indicating that P. horkelii might be exposed to this contam-
inant and that physiological effects should be expected. Further-
more, due to their lower potential to be transferred through the
food chain,Cd and Pb levels observed for P. horkelii might be
associated with waterborne rather than dietborne exposure.
Liver, muscle and blood samples presented higher concentra-
tions of specific metals (Cu, Cr and Fe, respectively), whereas gill
samples had the highest levels of all essential metals analyzed.
Except for Cu, lower concentrations of all metals were observed for
ovaries samples. Accumulation of Cu in liver and gills was also
observed for the Atlantic dogfish Scilyorhinus canicula (De Boeck
et al., 2010) but the high concentration observed in ovaries of
P. horkeliiwas atypical. Maternal offloading of tracemetals has been
suggested for elasmobranchs (Lopes et al., 2019; Hauser-Davis et al.,
2020), but not for Cu. In this study, however, considerable levels
were only observed for Cu, suggesting possible maternal transfer of
this metal to offspring.
All metals were detected at relatively low concentrations when
compared to similar studies carried out in the same sampling area
for green turtles (Chelonia mydas) (Da Silva et al., 2014) and seabirds
(Kehrig et al., 2015), which are migratory groups. Therefore,
contaminant levels of these organisms might not indicate specif-
ically the environmental pollution in southern Brazil. P. horkelii, on
the other hand, is a resident species, presenting longitudinal mi-
grations, and spending approximately three months in shallow
waters during the summer (Lessa et al., 1986; Vooren et al., 2005).
Despite the lack of studies reporting diet composition of P. horkelii,
this species possibly feeds mainly on benthic preys, as observed for
other sympatric Pseudobatos species (Bornatowski et al., 2010; Do
Carmo et al., 2015). In this context, Baraj et al. (2003) reported
high levels of trace metals in bivalves from the same sampling area,
indicating that other benthic preys might also present considerable
burdens which can potentially be transferred to P. horkelii during
the summer.
3.4. Contaminant distribution and tissue profile
Two PCAs were carried out for organic contaminants and trace
metals (Fig. 4) and the two principal components (PC1 and PC2)
were effective in summarizing the variability, explaining 88.7 and
70.6% of the total variance, respectively. Differences in tissues
contaminant profile could be observed for both PCAs, especially
blood samples.
Regarding organic compounds, a clear separation was observed
for liver and blood samples, with liver clustering associated with
higher concentrations of diclofenac in this organ, whereas blood
samples were clustered due to the lowest concentrations of any
organic contaminant in this tissue. The remaining tissues were not
distinguished by the contaminants profile considered in the PCA
although a slight association between gills and dichlofluanid and
octocrylene could be observed.
Differences in accumulation of organic contaminants per tissue
might be related to the compound’s physicochemical proprieties,
physiological and life-history patterns, as well as the organism’s
Fig. 3. Concentrations of (A) Cd, (B) Cr, (C) Cu, (D) Fe, (E) Hg, and (F) Pb for each tissue analyzed for Pseudobatos horkelii. Differences tested by Dunn’s Test are indicated by different
lowercase letters (p < 0.05).
Fig. 4. Principal Component Analysis (PCA) performed on standardized concentrations of (A) organic contaminants and (B) trace metals for Pseudobatos horkelii. Different colors
were arbitrarily attributed for each tissue and do not represent any cluster. (For interpretation of the references to color in this figure legend, the reader is referred to the Web
version of this article.)
M.F. Martins et al. / Environmental Pollution 265 (2020) 114923 7
xenobiotic elimination capacity (Corsolini et al., 2014; Mashroofeh
et al., 2015). For P. horkelii, the higher concentrations observed for
diclofenac, as well as for the other compounds not included in the
PCA can be explained by two factors: the high lipid content of
elasmobranchs liver (Gelsleichter and Walker, 2010, Table S2) and
its metabolic capacities. As liver concentrations are commonly
associated with acute exposure, diclofenac is likely transformed by
the liver and therefore, occurs in lower concentrations or as me-
tabolites in other organs. In addition, individuals analyzed in this
study were possibly under stress due to their concomitant preg-
nancy and vitellogenesis, which demands high lipid-mobilization
capacities, provided by the liver (Sheridan, 1988). In this context,
bioamplification (e.g. increase in contaminant concentrations due
to body mass loss, Daley et al., 2014) might, possibly, also explain
M.F. Martins et al. / Environmental Pollution 265 (2020) 1149238
the observed concentrations in liver samples analyzed for
P. horkelii.
The organic contaminants used as descriptors for the PCA failed
in differentiating gills, muscles and ovaries, contrary to the sepa-
ration of liver and blood samples described above.Mashroofeh et al.
(2015) detected lower concentrations of PAHs in muscle and gills
and attributed this to the low lipid content of this organs, contrary
to the lipid content of livers. Besides, fish muscle, for example, has
low CYP1A activity, indicating low xenobiotic metabolism effec-
tiveness (Daley et al., 2014; Beaudry et al., 2015), which can also be
associated with differential contaminant levels found among
tissues.
Regarding trace metals, a clearer separation of tissues was
observed, with blood clustering due to the higher concentrations of
Fe and Pb only, possibly due to the high association of red blood
cells with these compounds. Whereas Fe is essential for oxygen
transportation, Pb uptake occurs through gills, which might
transport this metal to blood (Wood et al., 2012). Blood concen-
trations are, however, temporary and usually indicate a momen-
taneous panorama.
Clustering of ovaries and liver samples was driven by Cu con-
centrations, which were higher in these tissues, whereas gills and
muscle were distributed along the Hg and Cr vectors. This distri-
bution showed that Cu concentrations were inversely low between
these tissues, whereas Hg and Cr were high. In fact, higher con-
centrations of Cr in gills and muscle in comparison to liver are
unusual, once Lopes et al. (2019) observed no difference between
these organs for the Brazilian electric ray Narcine brasiliensis.
Despite the comparable levels of Hg and Cr observed for P. horkelii,
the detection of these contaminants in less metabolic tissues, such
asmuscle indicate that the organismsmight be exposed chronically
to these compounds (Daley et al., 2014; Beaudry et al., 2015).
4. Conclusions
Our results indicate that P. horkelii is exposed to several con-
taminants in southern Brazil. Acute exposure was hypothesized to
be related to liver levels whereas high concentrations in less
metabolic tissues (e.g. muscle), especially of PAHs and emerging
pesticides might reflect chronic exposure. However, high liver
levels can also indicate chronic exposure and should not be only
associated with acute exposure, since elasmobranchs are less effi-
cient in xenobiotics metabolism in comparison with other verte-
brates. Considering that K-strategist species are particularly
vulnerable to effects of chronic exposure (Rowe, 2008) and that
P. horkelii presents a moderate growth rate (Caltabellotta et al.,
2019) ecotoxicological studies are crucial for assessing the poten-
tial vulnerability of this species to anthropogenic pollutants.
Moreover, guitarfishes are particularly vulnerable to population
declines (Dulvy and Forrest, 2010) and are considered a target
group for conservation efforts. In addition, the relative high
pollutant concentrations observed herein are also the first report
on muscle contamination in the illegally consumed Brazilian gui-
tarfish and might have human health implications.
The relatively high concentrations of some emerging contami-
nants are of particular interest as some of the compounds analyzed
here (diclofenac and octocrylene, for example) have never beenreported in any elasmobranch species and are still scarcely deter-
mined in wildlife. Due to the ubiquitous and pseudo-persistent
characteristic of emerging contaminants, we suggest that other
elasmobranch species might also be exposed to these compounds.
Furthermore, considering that some of them are known to impair
physiological functions in chronic exposure, it is possible that
exposure to constant concentrations of these compounds might
have some impact on elasmobranch populations.
Main findings
Pseudobatos horkelii is exposed to trace metals, polycyclic aro-
matic hydrocarbons and emerging contaminants and the highest
levels were found in liver samples.
Notes
The authors declare no competing financial interest.
Declaration of competing interest
The authors declare that they have no known competing
financial interests or personal relationships that could have
appeared to influence the work reported in this paper.
CRediT authorship contribution statement
Mariana F. Martins: Conceptualization, Formal analysis, Inves-
tigation, Writing - original draft, Writing - review & editing, Visu-
alization, Funding acquisition. Patrícia G. Costa: Conceptualization,
Methodology, Validation, Investigation, Writing - review & editing.
Adalto Bianchini: Conceptualization, Resources, Writing - review
& editing, Supervision.
Acknowledgements
This study was financed in part by the Coordenaç~ao de Aper-
feiçoamento de Pessoal de Nível Superior e Brasil (CAPES) e
Finance Code 001 and partially funded by the Save Our Seas
Foundation (SOSF 422). A. Bianchini is a research fellow from the
Brazilian Conselho Nacional de Desenvolvimento Científico e Tec-
nol�ogico (CNPq # 307647/20161). The authors thank the artisanal
and recreational fisherman from Rio Grande, southern Brazil for
donating the samples.
Appendix A. Supplementary data
Supplementary data to this article can be found online at
https://doi.org/10.1016/j.envpol.2020.114923.
References
Abril, S.I.M., Costa, P.G., Bianchini, A., 2018. Metal accumulation and expression of
genes encoding for metallothionein and copper transporters in a chronically
exposed wild population of the fish Hyphessobrycon luetkenii. Comp. Biochem.
Physiol. C 211, 25e31. https://doi.org/10.1016/j.cbpc.2018.05.008.
Alani, R., Ken, D., Kehinde, O., Babajide, A., 2012. Bioaccumulation of polycyclic ar-
omatic hydrocarbons in fish and invertebrates of lagos Lagoon, Nigeria.
J. Emerg. Trends Eng. Appl. Sci. 3 (2), 287e296.
Ali, A.M., Rønning, H.T., Sydnes, L.K., Alarif, W.M., Kallenborn, R., Al-Lihaibi, S.S.,
2018. Detection of PPCPs in marine organisms from contaminated coastal wa-
ters of the Saudi Red Sea. Sci. Total Environ. 621, 654e662. https://doi.org/
10.1016/j.scitotenv.2017.11.298.
Amado, L.L., Rosa, C.E., Leite, A.M., Moraes, L., Pires, W.V., Pinho, G.L.L.,
Martins, C.M.G., Robaldo, R.B., Nery, L.E.M., Monserrat, J.M., Bianchini, A.,
Martinez, P.E., Geracitano, L.A., 2006. Biomarkers in croakers Micropogonias
furnieri (Teleostei: sciaenidae) from polluted and non-polluted areas from the
Patos Lagoon estuary (Southern Brazil): evidences of genotoxic and immuno-
logical effects. Mar. Pollut. Bull. 52, 199e206.
Arpin-Pont, L., Bueno, M.J.M., Gomez, E., Fenet, H., 2014. Occurrence of PPCPs in the
marine environment: a review. Environ. Sci. Pollut. Res. 23, 4978e4991. https://
doi.org/10.1007/s11356-014-3617-x.
Baraj, B., Niencheski, L.F., Corradi, C., 2003. Trace metal content trend of mussel
Perna perna (linnaeus, 1758) from the Atlantic coast of southern Brazil. Water
Air Soil Pollut. 145, 205e214.
Barron, M.G., 2003. Bioaccumulation and bioconcentration in aquatic organisms. In:
Hoffman, D.J., Rattner, B.A., Burton, G.A., Cairns, J. (Eds.), Handbook of Ecotox-
icology. CRC Press, USA, pp. 877e892.
Beaudry, M.C., Hussey, N.E., McMeans, B., McLeod, A.M., Wintner, S.P., Cliff, G.,
Dudley, S.F.J., Fisk, A.T., 2015. Comparative organochlorine accumulation in two
https://doi.org/10.1016/j.envpol.2020.114923
https://doi.org/10.1016/j.cbpc.2018.05.008
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref2
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref2
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref2
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref2
https://doi.org/10.1016/j.scitotenv.2017.11.298
https://doi.org/10.1016/j.scitotenv.2017.11.298
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref4
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref4
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref4
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref4
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref4
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref4
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref4
https://doi.org/10.1007/s11356-014-3617-x
https://doi.org/10.1007/s11356-014-3617-x
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref6
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref6
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref6
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref6
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref7
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref7
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref7
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref7
M.F. Martins et al. / Environmental Pollution 265 (2020) 114923 9
ecologically similar shark species (Carcharodon carcharias and Carcharhinus
obscurus) with divergent uptake based on different life history. Environ. Tox-
icol. 9999, 1e10. https://doi.org/10.1002/etc.3029.
Beretta, M., Britto, V., Tavares, T.M., da Silva, S.M.T., Pletsch, A.L., 2014. Occurrence of
pharmaceutical and personal care products (PPCPs) in marine sediments in the
Todos of Santos Bay and the north coast of Salvador, Bahia, Brazil. J. Soils Sed-
iments 14, 1278e1286. https://doi.org/10.1007/s11368-014-0884-6.
Bezerra, M.F., Lacerda, L.D., Lai, C., 2019. Trace metals and persistente organic pol-
lutants contamination in batoids (Chrondrichthyes:Batoidea): a systematic re-
view. Environ. Pollut. 248, 684e695, 0.1016/j.envpol.2019.02.070.
Bonnefille, B., Gomez, H., Courant, F., Escande, A., Fenet, H., 2017. Diclofenac in the
marine environment: a review of its occurrence and effects. Mar. Pollut. Bull.
131, 496e506. https://doi.org/10.1016/j.marpolbul.2018.04.053.
Bornatowski, H., Robert, M.C., Costa, L., 2010. Feeding of guitarfish Rhinobatos
percellens (Walbaum, 1972) (Elasmobranchii, Rhinobatidae), the target of arti-
sanal fishery in southern Brazil. Braz. J. Oceanogr. 58 (1), 45e52.
Boxall, A.B.A., Rudd, M.A., Brooks, B.w., Caldwell, D.J., Choi, K., Hickman, S., Innes, E.,
Ostapyk, K., Staveley, J.P., Verslycke, T., Ankley, G.T., Beazley, K.F., Belanger, S.E.,
Berninger, J.P., et al., 2012. Pharmaceuticals and personal care products in the
environment: what are the big questions? Environ. Health Perspect. 120 (9),
1221e1229. https://doi.org/10.1289/ehp.1104477.
Bunholi, I.V., Ferrette, B.L.S., De Biasi, J.B., Magalh~aes, C.O., Rotundo, M.M.,
Oliveira, C., Foresti, F., Mendonça, F.F., 2018. The fishing and illegal trade of the
angelshark: DNA barcoding against misleading identifications. Fish. Res. 206,
193e197 doi: j.fishres.2018.05.018.
Caldas, S.S., Arias, J.L.O., Rombaldi, C., Mello, L.L., Cerqueira, M.B.R., Martins, A.F.,
Primel, E.G., 2019. Occurrence of pesticides and PPCPs in surface and drinking
water in southern Brazil: data on 4-year monitoring. J. Braz. Chem. Soc. 30 (1),
71e80. https://doi.org/10.21577/0103-5053.20180154.
Caltabellotta, F.P., Siders, Z.A., Murie, D.J., Motta, F.S., Cailliet, G.M., Gadig, O.B.F.,
2019. Age and growth of three endemic threatened guitarfishes Pseudobatos
horkelii, P. percellens and Zapteryx brevirostris in the western South Atlantic
Ocean. J. Fish. Biol. 95 (5), 1236e1248. https://doi.org/10.1111/jfb.14123.
Camacho, M., Boada, L.D., Or�os, J., L�opez, P., Zumbado, M., Almeida-Gonz�alez, M.,
Luzardo, O.P., 2014. Monitoring organic and inorganic pollutants in juvenile live
sea turtles: resultsfrom a study of Chelonia mydas and Eretmochelys imbricate
in Cape Verde. Sci. Total Environ. 481, 303e310. https://doi.org/10.1016/
j.scitotenv.2014.02.051.
Corsolini, S., Ancora, S., Bianchi, N., Mariotti, G., Leonzio, C., Christiansen, J.S., 2014.
Organotropism of persistent organic pollutants and heavy metals in the
Greenland shark Somniosus microcephalus in NE Greenland. Mar. Pollut. Bull.
87, 381e387. https://doi.org/10.1016/j.marpolbul.2014.07.021.
Cullen, J.A., Marshall, C.D., Halla, D., 2019. Integration of multi-tissue PAH and PCB
burdens with biomarker activity in three coastal shark species from the
northwestern Gulf of Mexico. Sci. Total Environ. 650, 1158e1172. https://
doi.org/10.1016/j.scitotenv.2018.09.128.
Da Silva, C.C., Varela Jr., A.S., Barcaroli, I.F., Bianchini, A., 2014. Concentrations and
distributions of metals in tissues of stranded green sea turtles (Chelonia mydas)
from the Southern Atlantic coast of Brazil. Sci. Total Environ. 466e467, 109e118.
https://doi.org/10.1016/j.scitotenv.2013.06.094.
Daley, J.M., Paterson, G., Drouillard, K.G., 2014. Bioamplification as a bio-
accumulation mechanism for persistent organic pollutants (POPs) in wildlife.
In: Whitacre, D.M. (Ed.), Reviews of Environmental Contamination and Toxi-
cology. Springer, Switerland, pp. 107e155. https://doi.org/10.1007/978-3-319-
01327-5_4.
Dann, A.B., Hontela, A., 2010. Triclosan: environmental exposure, toxicity and
mechanisms of action. J. Appl. Toxicol. 31, 285e311. https://doi.org/10.1002/
jat.1660.
De Boeck, G., Eyckmans, M., Lardon, I., Bobbaers, R., Sinha, A.K., Blust, R., 2010. Metal
accumulation and metallotionein induction in the spotted dogfish Scyliorhinus
canicula. Comp. Biochem. Physiol. A 155, 503e508.
De-Franco, B.A., Mendonça, F.F., Oliveira, C., Foresti, F., 2012. Illegal trade of the
guitarfish Rhinobatos horkelii on the coasts of central and Southern Brazil:
genetic identification to aid conservation. Aquat. Conserv. 22 (2), 272e276.
https://doi.org/10.1002/aqc.2229.
Do Carmo, W.P.D., Bornatowski, H., Oliveira, E.C., F�avaro, L.L., 2015. Diet of the chola
guitarfish Rhinobatos percellens (Rhinobatidae) in the Paranagu�a Estuarine
complex. Natl. Acad. Bras. Ciên. 87 (2), 721e731. https://doi.org/10.1590/0001-
3765201520140121.
Dulvy, N.K., Forrest, R.E., 2010. Life histories, population dynamics and extinction
risks in Chondrichthyans. In: Carrier, J.C., Musick, J.A., Heithaus, M.R. (Eds.),
Sharks and Their Relatives II: Biodiversity, Adaptive Physiology and Conserva-
tion. CRC Press, USA, pp. 639e679.
Fabbri, E., Franzellitti, S., 2016. Human pharmaceuticals in the marine environment:
focus on exposure and biological effects in animal species. Environ. Toxicol.
Chem. 35 (4), 799e812. https://doi.org/10.1002/etc.3131.
Fair, P.A., Lee, H., Adams, J., Darling, C., Pacepavicius, G., Alaee, M., Bossart, G.D.,
Henry, N., Muir, D., 2009. Occurrence of triclosan in plasma of wild Atlantic
bottlenose dolphins (Tursiops truncates) and in their environment. Environ.
Pollut. 157, 2248e2254. https://doi.org/10.1016/j.envpol.2009.04.002.
Fleeger, J.W., Carman, K.R., Nisbet, R.M., 2003. Indirect effects of contaminants in
aquatic ecosystems. Sci. Total Environ. 317, 207e233. https://doi.org/10.1016/
S0048-9697(03)00141-4.
Gago-Ferrero, P., Alonso, M.B., Bertozzi, C.P., Marigo, J., Barbosa, L., Cremer, M.,
Secchi, E.R., Azevedo, A., Lailson-Brito Jr., J., Torres, J.P.M., Malm, O., Eljarrat, E.,
Di�az-Cruz, M.S., Barcel�o, D., 2013. First determination of UV filters in marine
mammals. Octocrylene levels in Franciscana dolphins. Environ. Sci. Technol. 47,
5619e5625, 0.1021/es400675y.
Garcia, M.R., Mirlean, N., Baisch, P.R., Caram~ao, E.B., 2010. Assessment of polycyclic
aromatic hydrocarbon influx and sediment contamination in an urbanized es-
tuary. Environ. Monit. Assess. 168, 269e276. https://doi.org/10.1007/s10661-
009-1110-7.
Gelsleichter, J., Szabo, N.J., 2013. Uptake of human pharmaceuticals in bull sharks
(Carcharhinus leucas), inhabiting a wastewater-impacted river. Sci. Total Envi-
ron. 456e457, 196e201. https://doi.org/10.1016/j.scitotenv.2013.03.078.
Gelsleichter, J., Walker, C.J., 2010. Pollutant exposure and effects in sharks and their
relatives. In: Carrier, J.C., Musick, J.A., Heithaus, M.R. (Eds.), Sharks and Their
Relatives II: Biodiversity, Adaptive Physiology, and Conservation. CRC Press,
USA, pp. 491e537.
Gonzales-Alonso, S., Merino, L.M., Esteban, S., de Alda, M.L., Barcel�o, D., Dur�an, J.J.,
L�opez-Martínez, J., Ace~na, J., P�erez, S., Mastroianni, N., Silva, A., Catal�a, M.,
Valc�arcel, Y., 2017. Occurrence of pharmaceutical, recreational and psychotropic
drug residues in surface water on the northern Antarctic Peninsula region.
Environ. Pollut. 229, 241e254. https://doi.org/10.1016/j.envpol.2017.05.060.
Hauser-Davis, R.A., Pereira, C.F., Pinto, F., Torres, J.P.M., Malm, O., Vianna, M., 2020.
Mercury contamination in the recently described Brazilian White-tail dogfish
Squalus albicaudus (Squalidae, Chondrichthyes). Chemosphere 250, 126228.
https://doi.org/10.1016/j.chemosphere.2020.126228.
He, B., Wang, J., Liu, J., Hu, X., 2017. Eco-pharmacovigilance of non-steroidal anti-
inflammatory drugs: necessity and opportunities. Chemosphere 181, 178e189.
https://doi.org/10.1016/j.chemosphere.2017.04.084.
Henner, P., Schiavon, M., morel, J., Lichtfouse, E., 1997. Polycyclic aromatic hydro-
carbons (PAH) occurrence and remediation methods. Analusis 25, M56eM59.
Hylland, K., 2006. Polycyclic aromatic hydrocarbon (PAH) ecotoxicology in marine
ecosystems. J. Toxicol. Environ. Health 69, 109e123. https://doi.org/10.1080/
15287390500259327.
IBAMA, 2004. Instruç~ao normativa do minist�erio do meio ambiente n 5, de 21 de
maio de 2004. Di�ario Oficial da Uni~ao 32, 136e142.
IBGE, 2017. Estimativas da Populaç~ao Residente no Brasil e Unidades da Federaç~ao
com Data de Referência em 1 de Julho de 2017. ftp://ftp.ibge.gov.br/Estimativas_
de_ Populacao/Estimativas_2017/estimativa_dou_2017.pdf/. (Accessed 28
March 2018).
Jafarabadi, A.R., Bakhtiari, A.R., Yaghoobi, Z., King Yap, C., Maisano, M., Cappello, T.,
2019. Distributions and compositional patterns of polycyclic aromatic hydro-
carbons (PAHs) and their derivatives in three edible fishes from Kharg coral
Island, Persian Gulf, Iran. Chemosphere 215, 835e845. https://doi.org/10.1016/
j.chemosphere.2018.10.092.
Kehrig, H.A., Hauser-Davis, R.A., Seixas, T.G., Fillman, G., 2015. Trace-elements,
methylmercury and metallothionein levels in Magellanic penguin (Spheniscus
magellanicus) found stranded on the Southern Brazilian coast. Mar. Pollut. Bull.
96 (1e2), 450e455. https://doi.org/10.1016/j.marpolbul.2015.05.006.
Kidd, K.A., Blanchfield, P.J., Mills, K.H., Palace, V.P., Evans, R.E., Lazorchak, J.M.,
Flick, R.W., 2007. Collapse of a fish population after exposure to a synthetic
estrogen. Proc. Nat. Acad. Sci. USA 104, 8897e8901. https://doi.org/10.1073/
pnas.0609568104.
Konstantinou, I.K., Albanis, T.A., 2004. Worldwide occurrence and effects of anti-
fouling paint booster biocides in the aquatic environment: a review. Environ.
Int. 30, 235e248.
Lee, R.F., Sauerheber, R., Dobbs, G.H., 1972. Uptake, metabolism and discharge of
polycyclic aromatic hydrocarbons by marine fish. Mar. Biol. 17, 201e208.
Legendre, P., Legendre, L., 1998. Numerical Ecology. Elsevier Science BV,
Netherlands, p. 852.
Lessa, R., Vooren, C.M., 2016. Pseudobatos Horkelii. The IUCN Red List of Threatened
Species 2016. https://doi.org/10.2305/IUCN.UK.2016-
3.RLTS.T41064A103933918.en. Retrieved from.
Lessa, R., Vooren, C.M., Lahaye, J., 1986. Desenvolvimento e ciclo sexual das fêmeas,
migraç~oes e fecundidade da viola Rhinobatos horkelli (Müller & Henle, 1841) do
Sul do Brasil. Atlantica 8, 5e34.
Liu, J., Lu, G., Xie, Z., Zhang, Z., Li, S., Yan, Z., 2015. Occurrence, bioaccumulation and
risk assessment of lipophilic pharmaceutically active compounds in the
downstream rivers of sewage treatment plants. Sci. Total Environ. 511, 54e62.
https://doi.org/10.1016/j.scitotenv.2014.12.033.Logan, D.T., 2007. Perspective on ecotoxicology of PAHs to fish. Hum. Ecol. Risk
Assess. 13, 302e316. https://doi.org/10.1080/10807030701226749.
Lopes, C.A., Araujo, N.L.F., Rocha, L., Monteiro, F., Rocha, R.C.C., Saint’Pierre, T.D.,
Lufti, D.S., Vianna, M., Hauser-Davis, R.A., 2019. Toxic and essential metals in
Narcine brasiliensis (Elasmobranchii: narcinidae): a baseline ecotoxicological
study in Southeast Atlantic and preliminary maternal transfer implications.
Mar. Pollut. Bull. 149, 110606. https://doi.org/10.1016/j.marpolbul.2019.110606.
Lyons, K., Burket, S.R., Brooks, B.W., 2018. Sex may influence environmental
diphenhydramine accumulation in Round Stingrays. Mar. Pollut. Bull. 135,
648e653. https://doi.org/10.1016/j.marpolbul.2018.07.042.
Marsili, L., Coppola, D., Giannetti, M., Casini, S., Fossi, M.C., van Wyk, J.H., Sperone, E.,
Tripepi, S., Micarelli, P., Rizzuto, S., 2016. Skin biopsies as a sensitive non-lethal
technique for the ecotoxicological studies of great white shark (Carcharodon
carcharias) sampled in South Africa. Exp. Op. Environ. Biol. 4, 1. https://doi.org/
10.4172/2325-9655.1000126.
Martins, M.F., Pasquino, A.F., Gagig, O.B.F., 2018. Reproductive biology of the Bra-
zilian guitarfish, Pseudobatos horkelii (Müller & Henle, 1841) from southeastern
Brazil, western South Atlantic. J. Appl. Ichthyol. 34, 646e652. https://doi.org/
https://doi.org/10.1002/etc.3029
https://doi.org/10.1007/s11368-014-0884-6
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref10
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref10
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref10
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref10
https://doi.org/10.1016/j.marpolbul.2018.04.053
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref12
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref12
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref12
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref12
https://doi.org/10.1289/ehp.1104477
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref14
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref14
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref14
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref14
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref14
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref14
https://doi.org/10.21577/0103-5053.20180154
https://doi.org/10.1111/jfb.14123
https://doi.org/10.1016/j.scitotenv.2014.02.051
https://doi.org/10.1016/j.scitotenv.2014.02.051
https://doi.org/10.1016/j.marpolbul.2014.07.021
https://doi.org/10.1016/j.scitotenv.2018.09.128
https://doi.org/10.1016/j.scitotenv.2018.09.128
https://doi.org/10.1016/j.scitotenv.2013.06.094
https://doi.org/10.1007/978-3-319-01327-5_4
https://doi.org/10.1007/978-3-319-01327-5_4
https://doi.org/10.1002/jat.1660
https://doi.org/10.1002/jat.1660
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref24
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref24
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref24
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref24
https://doi.org/10.1002/aqc.2229
https://doi.org/10.1590/0001-3765201520140121
https://doi.org/10.1590/0001-3765201520140121
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref27
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref27
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref27
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref27
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref27
https://doi.org/10.1002/etc.3131
https://doi.org/10.1016/j.envpol.2009.04.002
https://doi.org/10.1016/S0048-9697(03)00141-4
https://doi.org/10.1016/S0048-9697(03)00141-4
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref31
https://doi.org/10.1007/s10661-009-1110-7
https://doi.org/10.1007/s10661-009-1110-7
https://doi.org/10.1016/j.scitotenv.2013.03.078
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref34
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref34
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref34
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref34
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref34
https://doi.org/10.1016/j.envpol.2017.05.060
https://doi.org/10.1016/j.chemosphere.2020.126228
https://doi.org/10.1016/j.chemosphere.2017.04.084
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref38
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref38
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref38
https://doi.org/10.1080/15287390500259327
https://doi.org/10.1080/15287390500259327
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref40
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref40
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref40
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref40
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref40
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref40
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref40
ftp://ftp.ibge.gov.br/Estimativas_de_%20Populacao/Estimativas_2017/estimativa_dou_2017.pdf/
ftp://ftp.ibge.gov.br/Estimativas_de_%20Populacao/Estimativas_2017/estimativa_dou_2017.pdf/
https://doi.org/10.1016/j.chemosphere.2018.10.092
https://doi.org/10.1016/j.chemosphere.2018.10.092
https://doi.org/10.1016/j.marpolbul.2015.05.006
https://doi.org/10.1073/pnas.0609568104
https://doi.org/10.1073/pnas.0609568104
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref45
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref45
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref45
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref45
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref46
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref46
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref46
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref47
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref47
https://doi.org/10.2305/IUCN.UK.2016-3.RLTS.T41064A103933918.en
https://doi.org/10.2305/IUCN.UK.2016-3.RLTS.T41064A103933918.en
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref49
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref49
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref49
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref49
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref49
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref49
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref49
https://doi.org/10.1016/j.scitotenv.2014.12.033
https://doi.org/10.1080/10807030701226749
https://doi.org/10.1016/j.marpolbul.2019.110606
https://doi.org/10.1016/j.marpolbul.2018.07.042
https://doi.org/10.4172/2325-9655.1000126
https://doi.org/10.4172/2325-9655.1000126
https://doi.org/10.1111/jai.13652
M.F. Martins et al. / Environmental Pollution 265 (2020) 11492310
10.1111/jai.13652.
Mashroofeh, A., Bakhtiari, A.R., Pourkazemi, M., 2015. Distribution and composition
pattern of polycyclic aromatic hydrocarbons in different tissues of sturgeons
collected from Iranian coastline of the Caspian Sea. Chemosphere 120,
575e583. https://doi.org/10.1016/j.chemosphere.2014.09.071.
Medeiros, P.M., Bícego, M.C., Castelao, R.M., Del Rosso, C., Fillmand, G., Zamboni, A.J.,
2005. Natural and anthropogenic hydrocarbon inputs to sediments of Patos
Laggon Estuary, Brazil. Environ. Int. 31, 77e87. https://doi.org/10.1016/
j.envint.2004.07.001.
Menni, R.C., Stehmann, M., 2000. Distribution, environment and biology of batoid
fishes off Argentina, Uruguay and Brazil. A review. Rev. Mus. Argentina Cienc
Nat. n.s. 2, 69e109. https://doi.org/10.22179/REVMACN.
Miller, T.H., Bury, N.R., Owen, S.F., MacRae, J.I., Barron, L.P., 2018. A review of the
pharmaceutical exposome in aquatic fauna. Environ. Pollut. 239, 129e146.https://doi.org/10.1016/j.envpol.2018.04.012.
Mirlean, N., Andrus, V.E., Baisch, P., 2003. Mercury pollution sources in sediments of
Patos Lagoon estuary, southern Brazil. Mar. Pollut. Bull. 46, 331e334.
Molins-Delgado, D., Mu~noz, R., Nogueira, S., Alonso, M.B., Torres, J.P., Malm, O.,
Ziolli, R.L., Hauser-Davis, R.A., Eljarrat, E., Barcel�o, D., Díaz-Cruz, M.S., 2018.
Occurrence of organic UV filters and metabolites in lebranche mullet (Mugil
liza) from Brazil. Sci. Total Environ. 618, 451e459. https://doi.org/10.1016/
j.scitotenv.2017.11.033.
Moore, A.B.M., Bolam, T., Lyons, B.P., Ellis, J.R., 2015. Concentrations of trace ele-
ments in a rare and threatened coastal shark from the Arabian Gulf (smooth-
tooth blacktip Carcharhinus leiodon). Mar. Pollut. Bull. 30, 646e650. https://
doi.org/10.1016/j.marpolbul.2015.06.005.
Nakata, H., 2005. Occurrence of synthetic musk fragrances in marine mammals and
sharks from Japanese coastal waters. Environ. Sci. Technol. 39 (10), 3430e3434.
https://doi.org/10.1021/es050199l.
Nakata, H., Mutara, S., Filatreau, J., 2009. Occurrence and concentrations of ben-
zotriazole UV stabilizers in marine organisms and sediments from the Ariake
sea, Japan. Environ. Sci. Technol. 43, 6920e6926. https://doi.org/10.1021/
es900939j.
Oaks, L.J., Gilbert, M., Virane, M.Z., Watson, R.T., Meteyers, C.U., Rideout, B.A.,
Shivaprasad, H.L., Ahmed, S.I., Chaudhry, M.J.I., Arshad, M., Mahmood, S., Ali, A.,
Khan, A.A., 2004. Diclofenac residues as the cause of vulture population decline
in Pakistan. Nature 427, 630e633.
Ojemaye, C.Y., Petrik, L., 2019. Occurrences, levels and risk assessment studies of
emerging pollutants (pharmaceuticals, perfluoroalkyl and endocrine disrupting
compounds) in fish samples from Kalk Bay harbor, South America. Environ.
Pollut. 252, 562e572. https://doi.org/10.1016/j.envpol.2019.05.091.
Omar, T.F.T., Aris, A.Z., Yusoff, F.M., Mustafa, S., 2019. Occurrence and level of
emerging organic contaminant in fish and mollusk from Klang River estuary,
Malaysia and assessment on human health risk. Environ. Pollut. 248, 763e773.
https://doi.org/10.1016/j.envpol.2019.02.060.
Overturf, M.D., Anderson, J.C., Pandelides, Z., Beyger, L., Holdway, D.A., 2015. Phar-
maceuticals and personal care products: a critical review of the impacts on fish
reproduction. Crit. Rev. Toxicol. 45 (6), 469e491. https://doi.org/10.3109/
10408444.2015.1038499.
Pereira, C.D.S., Maranho, L.A., Cortez, F.S., Pusceddu, F.H., Santos, A.R., Ribeiro, D.A.,
Cesar, A., Guimaraes, L.L., 2016. Occurrence of pharmaceuticals and cocaine in a
Brazilian coastal zone. Sci. Total Environ. 548e549, 148e154. https://doi.org/
10.1016/j.scitotenv.2016.01.051.
Quintela, F.M., Lima, G.P., Silveira, M.L., Costa, P.G., Bianchini, A., Loebmann, D.,
Martins, S.E., 2019. High arsenic and low lead concentrations in fish and reptiles
from Taim wetlands, a Ramsar site in southern Brazil. Sci. Total Environ. 660,
1004e1014. https://doi.org/10.1016/j.scitotenv.2019.01.031.
R Core Team, 2018. R: A Language and Environment for Statistical Computing. R
Foundation for Statistical Computing, Vienna, Austria. URL. https://www.R-
project.org/.
Rowe, C.L., 2008. “The calamity of so long life”: life histories, contaminants, and
potential emerging threats to long-lived vertebrates”. Bioscience 58 (7),
623e631. https://doi.org/10.1641/B580709.
Sathishkumar, P., Meena, R.A.A., Palanisami, T., Ashokkumar, V., Palvannan, T.,
Gu, F.L., 2020. Occurrence, interactive effects and ecological risk of diclofenac in
environmental compartments and biota- a review. Sci. Total Environ. 698,
134057, 0.1016/j.scitotenv.2019.134057.
Sauv�e, S., Desrosiers, M., 2014. A review of what is an emerging contaminant. Chem.
Cent. J. 8e15. https://doi.org/10.1186/1752-153X-8-15.
Schwaiger, J., Ferling, H., Mallow, U., Wintermayr, H., Negele, R.D., 2004. Toxic ef-
fects of the non-steroidal anti-inflammatory drug diclofenac. Aquat. Toxicol. 68
(2), 141e150. https://doi.org/10.1016/j.aquatox.2004.03.014.
Sheridan, M.A., 1988. Lipid dynamics in fish: aspects of absorption, transportation,
deposition and mobilization. Comp. Biochem. Physiol. B 90 (4), 679e690.
Storelli, M.M., Cuttone, G., Marcotrigiano, G.O., 2011. Distribution of trace elements
in the tissues of smooth hound Mustelus mustelus (Linneaus, 1758) from the
southern-eastern waters of Mediterranean Sea (Italy). Environ. Monit. Assess.
174, 271e281. https://doi.org/10.1007/s10661-010-1456-x.
Terrazas-L�opez, R., Arreola-Mendoza, L., Galv�an-Maga~na, F., Anguiano-Zamora, M.,
Sujitha, S.B., Johanthan, M.P., 2016. Cadmium concentration in liver and muscle
of silky shark (Carcharhinus falciformis) in the tip of Baja California south,
M�exico. Mar. Pollut. Bull. 107, 389e392. https://doi.org/10.1016/
j.marpolbul.2016.03.035.
European Union, 2013. Directive 2013/39 of the European parliament and of the
Council of 12 August 2013 Amending directives 2000/60/EC and 2008/105/EC as
regards priority substances in the field of water policy. Off. J. Eur. Union. L 226/1.
Van der Oost, R., Beyer, J., Vermeulen, N.P.E., 2003. Fish bioaccumulation and bio-
markers in environmental risk assessment: a review. Environ. Toxicol. Phar-
macol. 13, 57e149.
Vooren, C.M., Lessa, R.P., Klippel, S., 2005. Biologia e status de conservaç~ao da viola
Rhinobatos horkelii. In: Vooren, C.M., Klippel, S. (Eds.), Aç~oes para conservaç~ao
de tubar~oes e raias no sul do Brasil. Igar�e: Brazil, pp. 33e56.
Wallner-Kersanach, M., Mirlean, N., Baumgarten, M.G.Z., Costa, L.D.F., Baisch, P.,
2016. Temporal evolution of the contamination in the southern area of the
Patos Lagoon estuary, RS, Brazil. J. Integ. Coast. Zone Manag. 16 (3), 263e279.
https://doi.org/10.5894/rgci596.
Wood, C.M., Farrell, A.P., Brauner, C.J., 2012. Homeostasis and Toxicology of Non-
essential Metals. Elsevier Inc: Canada, p. 507.
Xue, J., Kannan, K., 2016. Accumulation profiles of parabens and their metabolites in
fish, black bear, and birds, including bald eagles and albatrosses. Environ. Int.
94, 546e533.
Xue, X., Xue, J., Liu, W., Adams, D.H., Kannan, K., 2017. Trophic magnification of
parabens and their metabolites in a subtropical marine food web. Environ. Sci.
Technol. 51, 780e789. https://doi.org/10.1021/acs.est.6b05501.
Zar, J.H., 2010. Biostatistical Analyses, fifth ed. Prentice Hall, USA, p. 944.
Zenker, A., Cicero, M.R., Prestinaci, F., Bottoni, P., Carere, M., 2014. Bioaccumulation
and biomagnification potential of pharmaceuticals with a focus to the aquatic
environment. J. Environ. Manag. 133, 378e387. https://doi.org/10.1016/
j.jenvman.2013.12.017.
Zhan, F., Yu, X., Zhang, X., Chen, L., Sun, X., Yu, R., Wu, Y., 2019. Tissue distribution of
organic contaminants in stranded pregnant sperm whale (Physeter micro-
cephalus) from the Huizhou coast of the South China Sea. Mar. Pollut. Bull. 144,
181e188. https://doi.org/10.1016/j.marpolbu.
https://doi.org/10.1111/jai.13652
https://doi.org/10.1016/j.chemosphere.2014.09.071
https://doi.org/10.1016/j.envint.2004.07.001
https://doi.org/10.1016/j.envint.2004.07.001
https://doi.org/10.22179/REVMACN
https://doi.org/10.1016/j.envpol.2018.04.012
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref60
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref60
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref60
https://doi.org/10.1016/j.scitotenv.2017.11.033
https://doi.org/10.1016/j.scitotenv.2017.11.033
https://doi.org/10.1016/j.marpolbul.2015.06.005
https://doi.org/10.1016/j.marpolbul.2015.06.005
https://doi.org/10.1021/es050199l
https://doi.org/10.1021/es900939j
https://doi.org/10.1021/es900939j
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref65
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref65
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref65
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref65
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref65
https://doi.org/10.1016/j.envpol.2019.05.091
https://doi.org/10.1016/j.envpol.2019.02.060
https://doi.org/10.3109/10408444.2015.1038499https://doi.org/10.3109/10408444.2015.1038499
https://doi.org/10.1016/j.scitotenv.2016.01.051
https://doi.org/10.1016/j.scitotenv.2016.01.051
https://doi.org/10.1016/j.scitotenv.2019.01.031
https://www.R-project.org/
https://www.R-project.org/
https://doi.org/10.1641/B580709
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref73
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref73
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref73
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref73
https://doi.org/10.1186/1752-153X-8-15
https://doi.org/10.1016/j.aquatox.2004.03.014
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref76
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref76
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref76
https://doi.org/10.1007/s10661-010-1456-x
https://doi.org/10.1016/j.marpolbul.2016.03.035
https://doi.org/10.1016/j.marpolbul.2016.03.035
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref79
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref79
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref79
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref80
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref80
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref80
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref80
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref81
https://doi.org/10.5894/rgci596
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref83
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref83
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref84
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref84
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref84
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref84
https://doi.org/10.1021/acs.est.6b05501
http://refhub.elsevier.com/S0269-7491(20)31707-3/sref86
https://doi.org/10.1016/j.jenvman.2013.12.017
https://doi.org/10.1016/j.jenvman.2013.12.017
https://doi.org/10.1016/j.marpolbu
	Contaminant screening and tissue distribution in the critically endangered Brazilian guitarfish Pseudobatos horkelii
	1. Introduction
	2. Materials and methods
	2.1. Study site
	2.2. Sample collection and animal handling
	2.3. Trace metals determination
	2.4. Organic compounds determination
	2.5. Data analysis
	3. Results and discussion
	3.1. Polycyclic Aromatic Hydrocarbons
	3.2. Emerging contaminants
	3.3. Inorganic contaminants
	3.4. Contaminant distribution and tissue profile
	4. Conclusions
	Main findings
	Notes
	Declaration of competing interest
	CRediT authorship contribution statement
	Acknowledgements
	Appendix A. Supplementary data
	References