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                    UNIVERSIDADE FEDERAL DE ALAGOAS
INSTITUTO DE CIÊNCIAS BIOLÓGICAS E DA SAÚDE
PROGRAMA DE PÓS-GRADUAÇÃO EM DIVERSIDADE BIOLÓGICA E
CONSERVAÇÃO NOS TRÓPICOS

FELIPE FELIX COSTA LIMA DA SILVEIRA

ECOTOXICIDADE DE PARABENOS EM ORGANISMOS AQUÁTICOS: A
INTERAÇÃO ENTRE MICROPLÁSTICOS DE POLIETILENO E METILPARABENO
NOS ESTÁGIOS INICIAIS DE DESENVOLVIMENTO DO ZEBRAFISH (Danio rerio)

MACEIÓ - ALAGOAS
11/2024

2
FELIPE FELIX COSTA LIMA DA SILVEIRA

ECOTOXICIDADE DE PARABENOS EM ORGANISMOS AQUÁTICOS: A
INTERAÇÃO ENTRE MICROPLÁSTICOS DE POLIETILENO E METILPARABENO
NOS ESTÁGIOS INICIAIS DE DESENVOLVIMENTO DO ZEBRAFISH (Danio rerio)
Dissertação apresentada ao Programa de
Pós-Graduação em Diversidade Biológica e
Conservação nos Trópicos, Instituto de Ciências
Biológicas e da Saúde. Universidade Federal de
Alagoas, como requisito para obtenção do título
de Mestre em CIÊNCIAS BIOLÓGICAS, área de
concentração em Conservação da Biodiversidade
Tropical.

Orientador:
Prof. Dr. Lázaro Wender Oliveira de Jesus
Co-orientadores:
Profa. Dra. Taciana Kramer de Oliveira
Prof. Dr. Thiago Lopes Rocha

MACEIÓ - ALAGOAS
11/2024

3

Catalogação na Fonte
Universidade Federal de Alagoas
Biblioteca Central
Divisão de Tratamento Técnico
Bibliotecário: Marcelino de Carvalho Freitas Neto – CRB-4 - 1767

4

Folha de aprovação
Felipe Felix Costa Lima da Silveira

ECOTOXICIDADE DE PARABENOS EM ORGANISMOS AQUÁTICOS: A
INTERAÇÃO ENTRE MICROPLÁSTICOS DE POLIETILENO E
METILPARABENO NOS ESTÁGIOS INICIAIS DE DESENVOLVIMENTO
DO ZEBRAFISH (DANIO RERIO)
Dissertação
apresentada
ao
Programa
de
Pós-Graduação em Diversidade Biológica e
Conservação nos Trópicos, Instituto de Ciências
Biológicas e da Saúde. Universidade Federal de
Alagoas, como requisito para obtenção do título de
Mestre em CIÊNCIAS BIOLÓGICAS, área de
concentração em Conservação da Biodiversidade
Tropical.
Dissertação aprovada em 27 de novembro de 2024

Presidente: Prof. Dr. Lázaro Wender Oliveira de Jesus (UFAL) — Orientador

Prof. Dr. Robson Guimarães dos Santos (UFAL)

Prof. Dr. Uedson Pereira Jacobina (UFAL)

Prof. Dr. Guilherme Malafaia Pinto/ Instituto Federal Goiano (IFG)
MACEIÓ - AL
Novembro/2024

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AGRADECIMENTOS
Agradeço à equipe do LMAA por todo o apoio operacional e moral e também
por toda a paciência e perseverança, em especial aos meus colegas Emilly, Lúcio, Ana
Carla e João. Agradeço também à equipe do LABAE pela ajuda com os experimentos e
com os dados, em especial à Bianca, à Lorranny e ao Felipe. Deixo aqui minha gratidão
ao professor Thiago, por ceder seu espaço e seus recursos para a realização dos
experimentos, e à professora Taciana pelo auxílio na análise dos dados, bem como pelo
apoio moral e compreensão em tempos difíceis. Além disso, agradeço aos professores
Robson, Guilherme, Uedson e César pelas relevantes contribuições para com esta
dissertação.
Sou grato às amigas que estiveram do meu lado nessa jornada, me
proporcionando momentos de alegria em meio ao caos e permanecendo sempre
próximas, ainda que fisicamente distantes: Jasmim, Raffaela e Vivian. Incluo também
meus companheiros de república, cuja companhia era sempre um alívio ao final de um
dia cheio.
Não posso deixar de agradecer à profissional que cuidou da minha saúde
mental durante todo esse tempo: Ponilla, talvez a melhor psicóloga do mundo, me
salvando semana após semana.
Embora elas não possam ler, agradeço também à coelhinha Marshmallow e à
gatinha Ramona por me lembrarem sempre que o amor puro existe e que ele é um
grande motivo para nunca desistir.

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RESUMO
Os contaminantes emergentes são substâncias de origem antrópica detectadas no
ambiente em diversas concentrações, mas sem perfil ecotoxicológico bem elucidado e
fora do escopo da legislação ambiental. A crescente industrialização introduz cada vez
mais substâncias e materiais nas matrizes ambientais, gerando preocupação pelos
possíveis efeitos deletérios à biota e à saúde humana. Destacam-se os
microplásticos, partículas com tamanho menor que 5 mm e de presença ubíqua em
ambientes aquáticos. Sua estrutura predominantemente hidrofóbica é capaz de
adsorver substâncias orgânicas e metais pesados, concentrando-os e potencialmente
aumentando sua toxicidade para organismos aquáticos. Os parabenos, ésteres do
ácido p-hidroxibenzoico, são conservantes encontrados em baixas concentrações em
meio aquático e que podem ter sua toxicidade ampliada ao serem adsorvidos por
microplásticos. A toxicidade de microplásticos e parabenos e suas possíveis
interações foram abordadas ao longo desta dissertação. Primeiramente, foi realizada
uma revisão da literatura referente ao estado da arte do conhecimento acerca da
bioacumulação e ecotoxicidade de parabenos em organismos aquáticos de diferentes
filos. Nela, evidenciou-se uma pequena quantidade de estudos de campo que
avaliavam bioacumulação e biomagnificação de parabenos em organismos aquáticos,
com subrepresentação de determinados grupos e resultados que variavam de acordo
com a localidade, com o tipo de organismo e tecido analisado. Além disso, verificou-se
grande heterogeneidade de protocolos experimentais e concentrações utilizadas,
predominância de poucas espécies-modelo e baixa utilização de controle analítico.
Também foi verificado um aumento no número de estudos ao longo do tempo, com
destaque para os anos de 2020 a 2023, bem como na variabilidade de biomarcadores
e vias metabólicas analisadas. A seguir, testou-se a partir do teste de toxicidade
embriolarval em zebrafish (ZELT) a hipótese de que a exposição de embriões e larvas
de zebrafish (Danio rerio) a uma combinação de três diferentes concentrações
ambientalmente relevantes de metilparabeno (MeP — 0,01, 0,1 e 1 µM) com
microplásticos de polietileno (MPPE — ~35 µM, 3,4 mg/L) levaria a efeitos
embriotóxicos mais pronunciados do que com a exposição isolada a cada um dos dois
contaminantes, bem como a hipótese de que a toxicidade do MeP é
concentração-dependente. Os resultados do ZELT sugerem que concentrações
ambientalmente relevantes de MeP são capazes de elicitar cardiotoxicidade em
embriões de zebrafish, embora a hipótese de que a mistura com MPPE amplifica sua
toxicidade não tenha sido confirmada. Além disso, destaca-se a relevância do
tamanho dos MPPE em seus efeitos ecotoxicológicos e a possibilidade de uma
redução na toxicidade do MeP em embriões de zebrafish com o uso de partículas de
tamanhos maiores. Chama-se atenção para a necessidade de mais estudos que
empreguem concentrações e vias de exposição realistas, buscando um panorama
mais preciso acerca das consequências da presença de tais contaminantes em meio
aquático.
Palavras-chave: poluição aquática; contaminantes emergentes; biomarcadores;
embriotoxicidade; peixes.

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ABSTRACT
Emerging contaminants are substances of anthropogenic origin detected in the
environment in various concentrations, but without a clear ecotoxicological profile and
outside the scope of environmental legislation. Growing industrialization introduces
more and more substances and materials into environmental matrices, raising concern
about the possible harmful effects on biota and human health. Of particular note are
microplastics, particles smaller than 5 mm and ubiquitous in aquatic environments.
Their predominantly hydrophobic structure is capable of adsorbing organic substances
and heavy metals, concentrating them and potentially increasing their toxicity to
aquatic organisms. Parabens, esters of p-hydroxybenzoic acid, are preservatives
found in low concentrations in aquatic environments, and their toxicity could be
increased when adsorbed by microplastics. The toxicity of microplastics and parabens,
along with their possible interactions, were addressed throughout this dissertation.
Firstly, a literature review was carried out on the state-of-the-art knowledge on the
bioaccumulation and ecotoxicity of parabens in aquatic organisms of different phyla.
This revealed a small number of field studies evaluating bioaccumulation and
biomagnification of parabens in aquatic organisms, with underrepresentation of certain
groups and results that varied according to the location, type of organism and tissue
analyzed. In addition, there was great heterogeneity in the experimental protocols and
concentrations used, a predominance of a few model species and low use of analytical
controls. There was also an increase in the number of studies over time, especially
between 2020 and 2023, as well as in the variability of biomarkers and metabolic
pathways analyzed. Subsequently, the zebrafish embryotoxicity test (ZELT) tested the
hypothesis that exposure of zebrafish (Danio rerio) embryos and larvae to a
combination of three different environmentally relevant concentrations of
methylparaben (MeP — 0. 01, 0.1 and 1 µM) with polyethylene microplastics (MPPE
— ~35 µM, 3 mg/L) would lead to more pronounced embryotoxic effects, 01, 0.1 and 1
µM) with polyethylene microplastics (MPPE, ~35 µM, 3.4 mg/L) would lead to more
pronounced embryotoxic effects than with isolated exposure to each of the two
contaminants, as well as the hypothesis that MeP toxicity is concentration-dependent.
The ZELT results suggest that environmentally relevant concentrations of MeP are
capable of eliciting cardiotoxicity in zebrafish embryos, although the hypothesis that
mixing with MPPE amplifies its toxicity has not been confirmed. In addition, the
relevance of the size of MPPE in its ecotoxicological effects and the possibility of a
reduction in the toxicity of MeP in zebrafish embryos with the use of larger particle
sizes are highlighted. Attention is drawn to the need for further studies using realistic
concentrations and exposure routes, in order to obtain a more accurate picture of the
consequences of the presence of such contaminants in aquatic environments.

Keywords: aquatic pollution; contaminants of emerging concern; biomarkers,
embryotoxicity; fish.

8
LISTA DE FIGURAS
Revisão da literatura
Figura 1. A) Processo de degradação de macroplásticos (> 5 mm) a microplásticos
(MPs — 1 µM a 5 mm) e nanoplásticos (NPs — 1 nM a 1 µM). B) Estrutura química
dos monômeros de alguns dos principais tipos de polímeros plásticos. C) Amostras de
microplásticos (MPs) coletadas do sedimento costeiro na província de Shandong,
China, incluindo: (a) MPs mistos, (b) pellets, (c) espumas, (d) fragmentos, (e) flocos,
(f)
filmes,
(g)
fibras,
e
(h)
esponjas.
.....………………………………………………………………………………….…………..19
Figura 2. Esquematização da origem, dispersão e circulação de microplásticos
primários e secundários entre as matrizes ambientais…………………………………..21
Figura 3. Resumo das rotas de degradação física e biótica de materiais plásticos em
ambientes aquáticos e de suas possíveis consequências deletérias para a biota……24
Figura 4. Efeitos biológicos de microplásticos em peixes teleósteos e fatores que
influenciam em sua toxicidade……………………………………………………………...26
Capítulo 1
Figure 1. Chemical structure of parabens commonly employed in commercial and
industrial formulations and also for p-hydroxybenzoic acid……………………………...46
Figure 2. Summarized study methodology step by step, from the scoping to the final
presentation………………………………………………………………………………...49
Figure 3. Number of articles published about ecotoxicological studies with parabens on
aquatic organisms and their distribution across the world. A) Absolute (columns) and
cumulative (line) number of articles published up to December 2023. B) Global
distribution and number of papers published by corresponding author's country……..51
Figure 4. Details of ecotoxicological studies focused on parabens in aquatic
organisms. A) Percentage of field and laboratory studies. B) Number of articles
published by groups of organisms in field studies. C) Frequency of the different
biological samples from aquatic organisms (invertebrates, vertebrates, and plants)
whose paraben s bioaccumulation was evaluated in field studies. The 21.9% include
the soft tissue of invertebrate species and also the whole body of some fish
species………………………………………………………………………………………...53
Figure 5. An overview of laboratory studies related to the ecotoxicity of parabens in
aquatic organisms. A) Number of published articles per taxon. B) Number of articles
published per fish species adopted in each study. C) Number of articles published by
fish life cycle stage assessed in each study. D) Frequency of organs and tissues

9
investigated in laboratory studies. E) Number of articles by type of paraben tested in
each study. F) Biomarkers and other parameters used in laboratory studies to
investigate the ecotoxicological effects of parabens……………………………………..65
Capítulo 2
Figure 1. A) Mortality rate of zebrafish embryos and larvae exposed to methylparaben
(MeP) in environmentally relevant concentrations either alone or in mixture with
polyethylene microplastics (MPPE) during the course of the 144h. B) Hatching rate of
zebrafish embryos exposed to MeP with or without MPPE during the course of
144h……………………………………………………………………………………….....110
Figure 2. A) Number of spontaneous movements per minute of 24 hpf zebrafish
embryos exposed to environmentally relevant concentrations of methylparaben (MeP)
with or without the presence of polyethylene microplastics (MPPE). B) Number of
heartbeats per minute of 48 hpf zebrafish embryos exposed to MeP and MPPE either
alone
or
in
mixture…………………………………………………………………………...................111
Figure 3. Photographic sheet of zebrafish embryos and larvae from five different
treatment groups (NC; SC; MPPE; MeP 1 µM; and MPPE + MeP 1 µM) at 24, 48 and
144 hpf. Scale bar represents 1000 µM for 24 and 48 hpf embryos and 4000 µM for 144
hpf
larvae.
There
are
no
visible
malformations
in
any
of
the
individuals…………..............................……116
Apêndice I
Figura 1. Captura de tela da primeira página do artigo de revisão publicado na revista
Environmental Pollution em novembro de 2024………………………………………….133

10
LISTA DE TABELAS
Capítulo 2
Table 1. Chemical and physical properties of main parabens∗………………………..134
Table 2. Summary of experimental studies with parabens in aquatic organisms
regarding the types and concentrations of parabens used, biomarkers investigated and
model
species………………………………………………………………………………...........136

11
SUMÁRIO
Apresentação​
13
Revisão da literatura​
14
Poluição ambiental e seus efeitos na biota​
14
Microplásticos: caracterização e distribuição em meio aquático​
17
Microplásticos e sua ecotoxicidade em organismos aquáticos
​
24
Parabenos: caracterização, distribuição ambiental e ecotoxicidade​
27
O zebrafish como organismo modelo​
31
Objetivos e hipóteses​
33
Referências​
34
Capítulo 1: Bioaccumulation and ecotoxicity of parabens in aquatic organisms:
Current status and trends​
44
1.1 Introduction​
44
1.2 Methodological approach​
48
1.3 Historical, geographical analysis and background​
50
1.4 Field studies on the bioaccumulation of parabens in aquatic organisms​
54
1.4.1 Green algae and aquatic angiosperms​ ​
​
​
​
​
55
1.4.2 Aquatic invertebrates​
56
1.4.3 Fishes​
57
1.4.4 Detection of paraben metabolites​
61
1.5 Laboratory studies​
62
1.5.1 Model species ​
63
1.5.2 Types of parabens tested​
67
1.5.3 Bioaccumulation in model species​
68
1.5.4 Ecotoxicological effects on aquatic organisms​
70
1.5.4.1 Effects on development and growth​
70
1.5.4.2 Effects on endocrine system​
72
1.5.4.3 Effects on reproductive system and reproduction​
74
1.5.4.4 Effects on nervous system and behavior​
75
1.5.4.5 Effects on the digestive system​
76
1.5.4.6 Effects on metabolism​
78
1.5.5 Types, exposure pathways, and analytical monitoring​
80
1.5.6 Environmental relevance of experimental studies and confounding factors​
81
1.6 Conclusions and future perspectives​
82
References​
84
Capítulo 2: Interactive effects of methylparaben and microplastics in the developing
zebrafish (Danio rerio)​
99
2.1 Introduction​
100
2.2 Materials and Methods​
103
2.2.1 Reagents​
103
2.2.2 Maintenance of adult zebrafish and collection of eggs​
104
2.2.3 Exposure​
104
2.2.4 Zebrafish embryo-larval toxicity test (ZELT)​
105
2.2.5 Behavioral analysis​
105
2.2.6 Morphological analysis​
106
2.2.7 Statistical analysis​
106

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2.3 Results and Discussion​
2.4 Conclusion​
References​
Considerações gerais​
Apêndice I​
Apêndice II​
Apêndice III​

107
108
119
132
133
134
136

13
APRESENTAÇÃO
Esta dissertação consiste de dois capítulos e trata da ecotoxicidade de
microplásticos e parabenos de forma isolada ou em conjunto através de estudos in vivo
realizados em organismos aquáticos.
No primeiro capítulo em formato de artigo, foi realizada uma extensa revisão
da literatura acerca da bioacumulação e ecotoxicidade de parabenos em organismos
aquáticos. Para isso, foi feita uma análise crítica de estudos conduzidos tanto em
campo, a partir da análise de tecidos animais e vegetais, quanto estudos experimentais
feitos com diversos modelos de microrganismos, invertebrados e vertebrados. Este
artigo foi publicado na revista Environmental Pollution em novembro de 2024.
No segundo capítulo, também em formato de artigo, foi realizado a avaliação
da toxicidade embriolarval em zebrafish (Danio rerio) de microplásticos de polietileno de
forma isolada ou em mistura com três diferentes concentrações ambientalmente
relevantes de metilparabeno. Tal estudo será futuramente acrescido de outras análises
realizadas com o mesmo modelo animal, destacando-se estudos comportamentais e
análises morfométricas, para posterior publicação em um periódico internacional da
área de ecotoxicologia e/ou conservação.

14
REVISÃO DA LITERATURA

Poluição ambiental e seus efeitos na biota
O conceito de “limites planetários” foi introduzido por Rockström et al (2009)
como um framework que estabelece uma “zona segura” para o desenvolvimento das
atividades humanas, dado o estado desejável dos sistemas que regem a vida no
planeta Terra durante a época do Holoceno, que abrange os últimos 11 500 anos do
período Quaternário. Tal conceito se baseia na ideia de que a estabilidade do ambiente
global é essencial para a manutenção da vida no planeta, e identifica nove limites
principais: mudanças climáticas, perda de biodiversidade genética e funcional,
interferência nos ciclos biogeoquímicos (nitrogênio e fósforo), acidificação dos oceanos,
uso da água doce, mudanças no uso do solo, novas entidades — referente à
contaminação por substâncias e materiais que não se encontravam nas matrizes
ambientais no período pré-industrial —, esgotamento da camada de ozônio e carga de
aerossóis atmosféricos. No ano de 2023, apenas três desses limites não haviam sido
ultrapassados: aerossóis atmosféricos, depleção do ozônio estratosférico e acidificação
dos oceanos (Richardson et al., 2023).
Estima-se que existam cerca de 350 mil substâncias químicas no mercado
global, com 70 mil registradas na última década. Cerca de 30 mil delas foram
registradas em economias emergentes, onde a capacidade de descarte de resíduos,
bem como a gestão de risco associado a eles, tendem a ser limitadas (Persson et al.,
2022). A produção de tais substâncias também gera subprodutos e impurezas que
geralmente não são considerados nas regulamentações e nas ações de fiscalização
ambiental. Nesse contexto, surge o princípio da cautela: conceito frequentemente
aplicado em políticas ambientais e de saúde, trata-se da ideia de que na ausência de
consenso científico sobre os riscos relacionados a determinado produto, medida,
política ou fenômeno, ações devem ser tomadas para prevenir potenciais danos ao
meio ambiente e à saúde pública (Aven, 2019). Como justificativa para a adoção de tal
princípio, é possível citar exemplos relevantes em que substâncias de toxicidade e
impacto ambiental e social desconhecidos, outrora comercializados em larga escala

15
visando a solução de problemas industriais, agropecuários e relacionados à saúde
humana, acabaram por tornar-se graves problemas (Bierbaum et al., 2020): os
clorofluorcarbonos (CFCs), amplamente utilizados em aerossóis e sistemas de
refrigeração na segunda metade do século XX, foram banidos em 1987 pelo Protocolo
de Montreal após a descoberta de sua capacidade de decompor o ozônio estratosférico
(Rowland,

1990).

O

diclorodifeniltricloroetano

(DDT),

herbicida

e

inseticida

organoclorado utilizado em larga escala no combate a pragas e vetores de doenças
entre os anos de 1940 e 1970, foi banido em diversos países devido à sua
neurotoxicidade, carcinogenicidade e alto potencial de desregulação endócrina, bem
como à sua grande persistência ambiental e efeitos deletérios no desenvolvimento e
sobrevivência de espécies de aves e peixes (Carolin C et al., 2023). Similarmente ao
DDT, as bifenilas policloradas (PCBs), compostos organoclorados cuja produção em
larga escala se iniciou no início do século XX devido às suas diversas aplicações
industriais, também foram largamente controladas e banidas em diversos países devido
à sua alta toxicidade para diversos sistemas orgânicos. Possui também alta
persistência ambiental, representando um risco para a saúde humana e para a
perpetuação de diversas espécies de vertebrados e invertebrados marinhos e terrestres
(Ngoubeyou et al., 2022). Dessa forma, entende-se atualmente que o parâmetro básico
que deve ser adotado ao se referir ao limite planetário de “novas entidades” é a
detecção em matrizes ambientais de substâncias e materiais cujo impacto sobre os
sistemas naturais e sobre a saúde humana ainda é total ou parcialmente desconhecido
(Liu et al., 2024).
Contaminantes

de

legado, definidos como substâncias amplamente

estudadas e regulamentadas devido aos seus conhecidos impactos ambientais e à
saúde humana, bem como à sua persistência ambiental e capacidade de sofrer
bioacumulação e biomagnificação (Azcune et al., 2022). Dentre eles estão os metais
pesados, metaloides tóxicos, pesticidas organoclorados e hidrocarbonetos aromáticos
policíclicos (HPAs), possuindo por vezes tratados internacionais que restringem sua
produção, uso e descarte, tal como a Convenção de Estocolmo (UN Environment
Programme, 2023). Os contaminantes de preocupação emergente (CECs), por sua vez,
são um conjunto de substâncias orgânicas e inorgânicas e de micropartículas de

16
origem

antropogênica que, com o desenvolvimento industrial e crescimento

populacional, têm sua presença recentemente documentada em matrizes ambientais.
Dentre eles estão os produtos de cuidado pessoal (PCPs) — encontrados em
cosméticos, protetores solares e produtos de higiene pessoal e limpeza doméstica —,
fármacos, meios de contraste de uso médico, drogas ilícitas, hormônios naturais e
sintéticos, alguns edulcorantes, diversos pesticidas, e os nanomateriais, com uma
extensa variedade química e introdução contínua de novas formulações no mercado
(Das et al., 2024). O relativo desconhecimento sobre seu comportamento no meio
ambiente e perfil ecotoxicológico — como bioacumulação, partição no sedimento e
coluna d’água e produtos de biodegradação — os tornam alvo de preocupação, dado o
fato de que não são usualmente incluídos pelas tecnologias de tratamento de águas
residuais nem possuem regulamentação, fiscalização e monitoramento abrangente
acerca de sua produção, uso, descarte e concentrações aceitáveis nas matrizes
ambientais (Li et al., 2024).
Em ecologia, o fitness ou aptidão de determinada população relaciona-se
diretamente com sua perpetuação, diferenciando-se entre elas de acordo com as taxas
de crescimento, sobrevivência e reprodução dos indivíduos (Laughlin e Messier, 2015).
Nesse contexto, a presença de determinadas substâncias com potencial de
desregulação endócrina (endocrine disrupting chemicals ou EDCs) pode prejudicar o
fitness reprodutivo de organismos aquáticos mesmo em baixas concentrações (Xiao et
al., 2023). A desregulação endócrina em decorrência da poluição ambiental não afeta
apenas os organismos e sua prole, mas a população da espécie como um todo e,
consequentemente, toda a comunidade biótica de um determinado local. Em um estudo
multigeracional, Kidd et al. (2007) registraram o colapso de uma população do peixe
norteamericano Pimephales promelas após a exposição ao já bem estabelecido
desregulador

endócrino

17-α-etinilestradiol,

estrógeno

sintético

utilizado

em

contraceptivos hormonais. Juntamente aos desreguladores endócrinos, a presença em
meio aquático de contaminantes com potencial de elicitar embriotoxicidade representa
um risco significativo para o desenvolvimento de organismos aquáticos, como peixes
(Escobar-Huerfano et al., 2020), anfíbios (Salla et al., 2024) e invertebrados (Ayari et
al., 2024). Tais contaminantes podem interferir diretamente nos processos iniciais de

17
desenvolvimento

embrionário

e

larval,

causando

malformações,

atrasos

no

crescimento, alterações no desenvolvimento sexual e no comportamento e aumento na
mortalidade dos embriões e larvas (Escobar-Huerfano et al., 2020; Dourdin et al.,
2023). A exposição a essas substâncias durante as fases mais sensíveis do ciclo de
vida dos organismos pode resultar em prole menos numerosa e resiliente,
comprometendo a biodiversidade e a estabilidade dos ecossistemas aquáticos. A partir
disso, é possível compreender a importância de se investigar substâncias com
potencial

de

causar

desregulação

endócrina,

alterações

reprodutivas

e

embriotoxicidade em organismos aquáticos, considerando sua capacidade de se
dispersar na coluna d’água e sua presença ubíqua em tais ambientes.
Microplásticos: caracterização e distribuição em meio aquático
A poluição por materiais plásticos é um problema ambiental de crescente
preocupação na atualidade, visto que eles são os principais detritos de origem
antrópica encontrados em ambientes aquáticos e estão presentes até mesmo em áreas
remotas e sem presença humana (Chassignet et al., 2021). Os polímeros plásticos são
leves, duráveis, maleáveis, resistentes e com baixo custo de produção, e sua utilização
perpassa os mais variados produtos, desde vestimentas, embalagens, cosméticos e
produtos de limpeza até componentes de veículos e equipamentos industriais (Boucher
e Friot, 2017). Apesar de suas vantagens práticas para as atividades humanas, seu
lado negativo torna-se cada vez mais evidente: estima-se que entre 8,8 e 11 milhões de
toneladas de plástico são lançadas aos oceanos todos os anos (Fava, 2022).
A partir dos anos 1960, começou a ser detectada a presença de
macroplásticos no estômago de aves marinhas e de outros animais aquáticos, bem
como o aumento na quantidade desses detritos em regiões costeiras. Os problemas
causados por esses materiais possuem relação com interferência física, podendo levar
à morte de animais por asfixia, ferimentos, emaranhamento, constrição ou ingestão —
o que pode obstruir seu trato digestório, bem como dar-lhes uma falsa sensação de
saciedade e fazê-los entrar em inanição (Schmid et al., 2021).
O termo “microplásticos” (MPs) foi mencionado pela primeira vez em 2004,
em um estudo de Thompson e colaboradores. Desde então, a presença de tais

18
partículas em águas superficiais, sedimentos e em organismos, assim como seus
possíveis efeitos prejudiciais nos ecossistemas aquáticos, tornou-se uma área de
pesquisa relevante (Rezania et al., 2018). Microplásticos são definidos como partículas
poliméricas sintéticas com tamanho entre 1 µM e 5 mm (Fig. 1A), podendo ter origem
primária ou secundária. Os MPs primários incluem pellets plásticos usados como
matéria-prima pela indústria, agentes esfoliantes em produtos de cuidado pessoal
(PCPs), além de partículas liberadas pelo desgaste de pneus ou pela lavagem de
tecidos sintéticos. Já os MPs secundários são formados pela degradação de
macroplásticos, através de processos químicos, exposição à luz ou ação biológica
(Rezania et al., 2018; Schmid et al., 2020). Além de diferentes origens, os
microplásticos também são materiais com características altamente diversificadas,
podendo variar quanto a: 1. Tipo de polímero, como polietileno (PE), polipropileno (PP),
poliestireno (PS), poliamida (PA), poliuretano (PU), acetato de polivinila (PVA), cloreto
de polivinila (PVC), polietileno tereftalato (PET), polimetil metacrilato (PMMA), dentre
outras (Fig. 1B); 2. Cor; 3. Formato, apresentando-se como esferas, fibras isoladas ou
em emaranhados, fragmentos, pellets, espumas, filmes, dentre outros (Fig. 1C); 4.
Presença de aditivos associados, como estabilizantes, plastificantes, corantes e
retardantes de chama (Rochman et al., 2019).

19

Figura 1. A) Processo de degradação de macroplásticos (> 5 mm) a microplásticos (MPs — 1
µM a 5 mm) e nanoplásticos (NPs — 1 nM a 1 µM). Fonte: Adaptado de Belmaker et al.
(2024). B) Estrutura química dos monômeros de alguns dos principais tipos de polímeros
plásticos. Fonte: adaptado de Urbanek et al. (2018). C) Amostras de microplásticos (MPs)
coletadas do sedimento costeiro na província de Shandong, China, incluindo: (a) MPs mistos,
(b) pellets, (c) espumas, (d) fragmentos, (e) flocos, (f) filmes, (g) fibras, e (h) esponjas. Fonte:
Zhou et al. (2018).

Segundo

a

Associação

de

Produtores

de

Plástico

da

Europa

(PlasticsEurope, 2020), o PE de baixa e de alta densidade (LDPE e HDPE) são,
respectivamente, os polímeros termoplásticos com segunda e terceira maior demanda
de mercado no continente, estando atrás apenas do PP. Tais polímeros, devido a seu
baixo custo e grande versatilidade, são frequentemente usados em produtos plásticos
descartáveis e de uso único, de forma que sua produção e descarte ocorrem de forma

20
acelerada e em larga escala (Erni-Cassola et al., 2019). Como exemplo, análises de
duas amostras de águas residuais da cidade de Karlsruhe, na Alemanha, Majewski et
al. (2016) relataram o polietileno como sendo o polímero mais frequentemente
detectado, compondo 34% (81 mg/m3) e 17% (257 mg/m3) de cada uma delas. Para
além da quantidade absoluta produzida e descartada, a abundância relativa de
diferentes tipos de polímeros plásticos no ambiente aquático depende do local
analisado: polímeros de baixa densidade (< 1 g/cm³), como PE e PP, conseguem flutuar
na coluna d’água e são frequentemente encontrados em amostras coletadas em zonas
superficiais de mar aberto, porém estão presentes em menor quantidade em zonas
intertidais e subtidais, bem como em regiões mais profundas da coluna d’água.
Polímeros mais densos, como o poliéster, PA e o acrílico, por sua vez, são mais
frequentemente encontrados em zonas subsuperficiais e em sedimentos marinhos
(Boucher e Friot, 2017; Erni-Cassola et al., 2019). Os rios são uma importante fonte
carreadora de detritos plásticos para ambientes estuarinos e marinhos (Fig. 2),
coletando rejeitos industriais, águas de escoamento urbano e efluentes domésticos
com quantidades mensuráveis de polímeros plásticos como PE, PP, PS, PA, PVA e PU
em tamanhos de <100 a 3500 µM (Stanton et al., 2020). Sugere-se que partículas
plásticas com tamanho entre 0,333 e 4,75 mm sejam a vasta maioria, compreendendo
mais de 90% dos plásticos encontrados em águas superficiais marinhas (Trevisan et
al., 2020).

21

Figura 2. Esquematização da origem, dispersão e circulação de microplásticos primários e
secundários entre as matrizes ambientais. Fonte: Lee et al. (2022)

Foi relatada a presença de MPs de diferentes tipos e tamanhos em
ambientes marinhos de diversos países entre os anos de 2013 e 2018. Em uma
revisão, Rezania et al. (2018) listaram estudos que verificaram a presença de MPs na
coluna d’água e sedimentos marinhos em diferentes concentrações e em numerosas
partes do mundo. Em praias da Coreia do Sul, encontrou-se uma média de 27 606/m2
partículas de MPs, com predominância de PS. Na baía de Jinhae, também na Coreia
do Sul, foram encontradas entre 33-83 partículas/L de MPs, sendo, à época (2015) a
maior abundância de micropartículas plásticas flutuantes já relatada em águas
superficiais. No Canadá, relatou-se a presença de 20 a 80 micropartículas plásticas a
cada de 10 g de sedimento analisado no estudo. Estudos conduzidos em diversas
regiões costeiras da China entre os anos de 2018 e 2020 encontraram concentrações
variáveis de microplásticos, com as menores médias sendo referentes a regiões de mar
aberto (0,31 itens/m³ no Mar do Leste da China, 0,40 a 5.2 itens/m³ no Mar Bohai) e, as
maiores, a regiões estuarinas (67,5 ± 94,4 itens/m³ no Estuário Yangtze). Foram

22
encontradas altas concentrações de microplásticos nos sedimentos da Baía Xiangshan
(1740 ± 2150 itens/kg, peso seco) e do Mar Amarelo (2580 ± 1140 itens/kg, peso seco),
contrastando com baixas concentrações em suas respectivas colunas d’água — 8.91 ±
4.7 itens/m³ e 0.33 ± 0.28 itens/m³, respectivamente (Wang et al., 2020). Em 2019,
Zheng et al. (2019) encontraram microplásticos em uma concentração média de 46 ± 28
itens/m³ na coluna d´água da Baía Jiaozhou, na China, bem como uma média de 15 ± 6
itens/kg, peso seco, nos sedimentos locais. Dentre os tipos de polímeros encontrados,
predominavam o PET, o PP e o PE, formando respectivamente 56,25%, 34,38% e
3,13% das partículas detectadas nas amostras de água e 51,35%, 21,62% e 8,11%
daquelas detectadas nas amostras de sedimentos. Em geral, a heterogeneidade nas
concentrações de microplásticos detectadas em diversos estudos reflete variações
geográficas nas regiões estudadas, como conformação espacial da zona costeira,
correntes marítimas, condições climáticas, variações sazonais, nível de troca de água,
atividade fluvial e nível de atividade humana (Erni-Cassola et al., 2019; Wang et al.,
2020).
Além da presença na coluna d’água e em sedimentos de regiões costeiras e
marinhas de diferentes regiões do mundo, também já foi detectada a presença de
micropartículas plásticas em tecidos de diversas espécies de peixes e invertebrados
marinhos. De acordo com uma revisão publicada por Sequeira et al. (2020), de
amostras coletadas de 198 espécies de peixes em 24 países, sendo que apenas 14%
destes eram provenientes de aquicultura, 60% continham microplásticos em seus
órgãos e tecidos. Dentre os principais polímeros encontrados estavam PE (16%), PP
(14%), PS (24%), PA, (11%) e PET (5%). O nível trófico e os hábitos alimentares dos
animais influenciaram significativamente o número de microplásticos por indivíduo, com
valores mais altos para peixes carnívoros e predadores, dando suporte à hipótese de
transferência trófica para tais contaminantes. Dentre os órgãos mais frequentemente
avaliados pelos estudos estavam os órgãos do sistema digestivo, as brânquias, o
músculo esquelético e a pele. Em um estudo conduzido por Wang et al. (2020) na Baía
Hangzhou, China, foram encontrados 70 microplásticos em 92 amostras extraídas de
quatro espécies de peixes e três espécies de crustáceos. Todas as espécies de peixes
analisadas possuíam maior abundância média de MPs em comparação às espécies de

23
crustáceos, provavelmente refletindo seus hábitos predatórios e demersais — além da
transferência trófica, a maior proximidade com o sedimento, um dos principais
sumidouros de MPs, poderia facilitar a ingestão de tais partículas.
A superfície dos microplásticos possui capacidade de realizar diversas
interações intermoleculares com substâncias presentes nas matrizes ambientais,
destacando-se as interações hidrofóbicas, as forças de Van Der Waals, as interações
pi-pi, as interações eletrostáticas, as ligações covalentes, a complexação iônica e as
ligações de hidrogênio. Isso permite a adsorção de diversos tipos de contaminantes em
sua superfície, tais como moléculas orgânicas apolares, polares e ionizadas, bem como
substâncias inorgânicas como metais pesados (Fig. 3) (Rafa et al. 2024). A capacidade
de adsorção e dessorção de tais moléculas da superfície dos microplásticos depende
de diversos fatores, como o tipo de polímero, o tamanho, formato e nível de
degradação das partículas, a carga e polaridade do contaminante e a temperatura,
salinidade e pH do ambiente onde se encontram. Sendo de tamanho diminuto e,
portanto, de fácil ingestão por parte de organismos aquáticos, os microplásticos podem
agir como um “Cavalo de Troia” ao funcionar como potenciais carreadores de poluentes
de alta toxicidade e de microorganismos patogênicos adsorvidos em sua superfície
(Hartmann et al., 2017; Rafa et al., 2024). Dentre os exemplos de contaminantes
ambientais com capacidade de sofrer sorção pela superfície de microplásticos, estão
HPAs (Tan et al., 2019), antibióticos (Li et al., 2018), pesticidas (Li et al., 2021),
nanopartículas (Singh et al., 2021), fungicidas (Fang et al., 2019), per- e
polifluoroalquilados (PFAS) (Dai et al, 2022), hormônios (Hu et al, 2020), PCPs (Zhou et
al. 2020) e metais pesados como cádmio, níquel e chumbo (Tenea et al., 2024).

24

Figura 3. Resumo das rotas de degradação física e biótica de materiais plásticos em ambientes
aquáticos e de suas possíveis consequências deletérias para a biota.Fonte: Urbanek et al.
(2018).

Microplásticos e sua ecotoxicidade em organismos aquáticos
Diversos estudos experimentais relataram a ocorrência de consequências
negativas causadas por MPs e nanoplásticos (NPs) em embriões, larvas, juvenis e
indivíduos adultos de espécies de peixes como o zebrafish, Oryzias spp., Tigriopus
japonicas, Cyprinus carpio, Carassius spp., Pimephales promelas, Pomatoschistus
microps, Sparus aurata, Acanthochromis polyacanthus, dentre outros. Dentre elas,
estavam processos inflamatórios, danos ao sistema digestório, prejuízos ao
crescimento e desenvolvimento, estresse oxidativo, neurotoxicidade, alterações
comportamentais e menor viabilidade e sobrevivência de embriões e larvas (Yong et al.,
2020; Hasan et al., 2024).
Barboza et al. (2018) observaram uma inibição na atividade da
acetilcolinesterase (AChE) cerebral e na isocitrato desidrogenase (IDH) muscular de

25
indivíduos juvenis do peixe Dicentrarchus labrax expostos a uma concentração de 0,69
mg/L de microplásticos ao longo de 96 horas, bem como aumento na peroxidação
lipídica no cérebro e músculo dos animais testados. Malafaia et al. (2020) registraram
efeitos negativos da exposição a micropartículas de polietileno (MPPE — 6,2 a 100
mg/L) sobre a sobrevivência e desenvolvimento de embriões e larvas de zebrafish
expostos por 144h sob um regime semi-estático. Dentre estes efeitos, estavam a
eclosão precoce de embriões, a maior mortalidade e alterações morfométricas de
larvas, como maior altura da cabeça, maior área da vesícula óptica e do saco vitelínico
e, nas maiores concentrações (50 a 100 mg/L), maior distância interocular e maior
distância entre os mioseptos. Tais alterações poderiam ocasionar em perda da
acuidade visual e problemas na contratilidade do músculo esquelético nos animais,
prejudicando sua capacidade de nado e sua adaptabilidade comportamental. Em
indivíduos juvenis do peixe marinho Sebastes schlegelii expostos a microplásticos de
PS (MPPS) na concentração de 106 microesferas/L durante 14 dias, verificou-se
alterações comportamentais como redução no tempo de forrageamento e de
alimentação, além de inibição na atividade locomotora quanto à distância percorrida e
velocidade de nado. Foram observadas também alterações histopatológicas, como
congestão e hiperemia hepáticas e escurecimento biliar, além de uma menor taxa de
crescimento e menor concentração de conteúdo proteico e lipídico total (Yin et al.,
2018).
Em adultos de zebrafish, a exposição a micropartículas de PE e PS (100 e
1000 µg/L) durante 20 dias foi capaz de induzir alterações pró-inflamatórias, estresse
oxidativo, danos na mucosa intestinal e no epitélio branquial, alterações morfológicas,
redução na expressão de genes relacionados ao metabolismo energético e à função
imune e mudanças no ritmo circadiano, com significativo aumento de atividade
locomotora no período noturno (Limonta et al., 2019). A exposição a micropartículas de
polietileno e polipropileno naturalmente desgastadas durante 21 dias (0,1 e 1 mg/L)
levou adultos de zebrafish a apresentarem comportamentos sugestivos de ansiedade,
além de alterações na função mitocondrial cerebral e aumento na atividade de enzimas
antioxidantes hepáticas (Félix et al., 2023).

26

Figura 4. Efeitos biológicos de microplásticos em peixes teleósteos e fatores que influenciam
em sua toxicidade. Fonte: Liu et al. (2024).

Em 2019, Wang et al. verificaram diversos marcadores de toxicidade na
exposição do teleósteo medaka marinho (Oryzias melastigma) a MPPS (2, 20 e 200
µg/L) durante 60 dias, tais como alterações histopatológicas nas brânquias e gônadas,
aumento nos marcadores de estresse oxidativo nas brânquias, intestinos, fígado e
testículos de machos, desregulação nos hormônios sexuais das fêmeas e redução de
sua fecundidade. Além disso, os efeitos deletérios foram transgeracionais, afetando
negativamente o desenvolvimento da prole dos animais expostos aos tratamentos com
MPPS. Qiang e Cheng (2021), em um estudo também com MPPS (10, 100 e 1000
µg/L), encontraram alterações histopatológicas e aumento no estresse oxidativo e nos
marcadores apoptóticos nas gônadas de zebrafish expostos aos microplásticos por 21
dias.
A partir de uma análise integrada de dados de biomarcadores relacionados a
estresse oxidativo e neurotoxicidade em juvenis de D. labrax expostos a microplásticos
e a mercúrio de forma isolada ou em mistura, Barboza et al. (2018) encontraram
evidência de interação significativa entre os dois contaminantes, levando a efeitos
biológicos diferentes dos exibidos na exposição isolada. Em um estudo conduzido por

27
Bihanic et al. (2020), embriões e larvas de Oryzias melastigma foram expostos a MPPE
(4 to 6 µM, 10 mg/L) durante 12 dias, isoladamente ou sorvidos com três
contaminantes:

benzo(a)pireno

(BaP

—

0,01

e

16,64

µg/g

MP),

ácido

perfluoro-octanosulfônico (PFOS — 0,12 e 55,65 µg/g MP) e benzofenona-3 (BP3 —
0,14 e 24 ng/g MP). Na concentração mais alta de PFOS sorvidos em MPPE, houve
redução na taxa de eclosão e no comprimento total final, ao passo que tais alterações
não foram observadas na exposição ao PFOS de forma isolada. Enquanto a exposição
isolada ao BaP na maior concentração não foi capaz de induzir toxicidade, a exposição
conjunta com MPPE levou a uma maior incidência de anormalidades morfológicas e
uma redução na atividade locomotora quanto à distância percorrida. Na exposição ao
BP3 em maior concentração sorvido ao MPPE, observou-se redução no comprimento
final total e no comprimento da cabeça, o que não foi observado na exposição isolada
ao BP3. Não houve evidência de ingestão dos MPPE pelos animais, porém foi
visualizado um acúmulo de partículas ao longo da superfície do córion, sugerindo um
contato direto dos MPPE com os embriões através deste.
Em 2019, Roje e colaboradores realizaram um estudo in vitro realizado com
linhagens celulares de adenocarcinoma mamário sensível a estrógenos expostas a
uma associação de nanopartículas de poliestireno (NPPS — 1, 10 e 100 ppm) com
uma mistura de parabenos (PBmix — 0,01 a 1 µg/mL). Observou-se um efeito sinérgico
sobre a atividade dessas células, com a proporção de células MCF-7 em proliferação
ativa mais do que dobrando ao serem expostas à concentração mais alta de NPPS em
combinação com PBmix em uma concentração constante de 1 µg/mL. Foi sugerido que
tais achados poderiam ser resultantes de uma interação dos parabenos com a estrutura
dos NPPS, de forma a concentrar tais moléculas em sua superfície e carreá-las para o
interior das células, intensificando sua atividade pró-estrogênica.
Parabenos: caracterização, distribuição ambiental e ecotoxicidade
Os parabenos, considerados como CECs, são um conjunto de substâncias
orgânicas de baixo peso molecular, resultantes da reação de esterificação do ácido
4-hidroxibenzóico (4-HB). São utilizadas desde 1920 como conservantes em
preparações farmacêuticas, e na atualidade são encontradas também em cosméticos,

28
alimentos e outros produtos industrializados (Nowak et al., 2018; Bolujoko et al., 2022).
Dentre eles, os mais comumente empregados são o metilparabeno (MeP) e o
propilparabeno (PrP), sozinhos ou em associação. São estáveis em soluções aquosas
de baixo pH e suas propriedades variam de acordo com o tamanho de seu grupamento
alquila: as maiores cadeias possuem maior atividade antimicrobiana e maior resistência
à hidrólise, mas menor solubilidade em água — o MeP e o PrP possuem cadeias de um
e três carbonos, respectivamente (Wei et al., 2021; Pereira et al., 2023). Devido ao seu
influxo contínuo em baixas concentrações nas águas superficiais através dos efluentes
antropogênicos, os parabenos podem ser considerados como contaminantes
pseudopersistentes, bem como os outros PCPs em geral (Garric, 2013).
Como resultado de pesquisas conduzidas em diversos países, foi constatada
a presença de parabenos em diferentes ambientes aquáticos, na água potável, em
efluentes urbanos e em solos agrícolas (Feng et al., 2019). A presença de parabenos
no esgoto urbano é bem documentada e, mesmo com sua baixa estabilidade, alta
biodegradabilidade em condições aeróbias e uma eficácia de remoção de acima de
90% durante o processo de tratamento de água (Haman et al., 2015), eles ainda são
encontrados nos efluentes tratados e na água e sedimentos de corpos d’água, assim
como o seu metabólito 4-HB (Feng et al., 2019). Os parabenos também são capazes
de gerar derivados clorados mais estáveis e persistentes do que as substâncias
originais, principalmente devido ao processo de cloração utilizado no tratamento de
água (Bolujoko et al., 2021).
No que diz respeito às concentrações de parabenos encontradas em
ambientes naturais e antropizados, há grande variabilidade de acordo com o local
estudado. As maiores concentrações são encontradas em efluentes urbanos, chegando
a 76 900 ng/L em uma estação de tratamento de água do sul da Califórnia (Błędzka et
al., 2014). No Brasil, o MeP foi detectado em córregos da cidade de Rio Grande/RS em
concentrações entre 7,6 e 29,8 µg/L, e na cidade de Morro Redondo/RS, entre <1 e
134 µg/L (Penha et al., 2021). Derisso et al. (2020), ao analisar sete pontos do Rio
Monjolinho na cidade de São Carlos/SP, detectaram concentrações de MeP que
variavam entre 0,11 e 0,98 µg/L. Na avaliação de três rios da região de Curitiba/PR,
Santos et al. (2016) encontraram concentrações de MeP de até 2875 ng/L.

29
Além de sua presença constatada em águas superficiais, também há
evidências de bioacumulação de parabenos em tecidos animais e humanos. Xue e
Kannan (2016) relataram a presença de MeP e seu metabólito 4-HB no rim, fígado e
tecido

muscular de águias carecas (Haliaeetus leucocephalus) e albatrozes

(Phoebastria spp.) em concentrações que variavam de 580 ng/g (MeP) a 35 - 300 ng/g
(4-HB), dependendo do tecido e da espécie. No mesmo estudo, relatou-se a presença
de MeP, PrP e 4-HB nos tecidos hepático e cerebral de peixes da costa da Flórida, em
concentrações que variaram de 11,2 ng/g para o MeP a 1130 ng/g para o 4-HB. As
concentrações teciduais maiores do que as plasmáticas são sugestivas de
bioacumulação, principalmente no fígado (Xue e Kannan, 2016). A bioacumulação de
4-HB e parabenos também já foi relatada em mamíferos marinhos como lontras e
golfinhos, e a relação entre sua concentração e a de MeP nestes animais sugere que a
fonte de tais substâncias tenha sido, de fato, antropogênica (Xue e Kannan, 2016).
Chiesa et al. (2018) encontraram diversos parabenos e 4-HB em peixes pelágicos
como Salmo trutta, Salmo solar e Thunnus albacares, bem como em bivalves, que são
os invertebrados com a maior capacidade de bioacumulação dessas substâncias. A
análise de invertebrados marinhos por Xue et al. (2017) demonstrou a presença de
MeP em 82% das amostras, em concentrações que variavam de 9,43 a 322 ng/g, a
depender da espécie. Neste estudo, chegou-se ao valor de 1,83 para o fator de
magnificação trófica (TMF) do MeP a partir da análise de uma teia alimentar subtropical
composta por 13 espécies, sugerindo um potencial significativo de biomagnificação
para este composto.
A partir do final do século XX, com a publicação de estudos sugerindo a
atividade estrogênica e antiandrogênica dos parabenos, cresceu a preocupação com
seus possíveis efeitos negativos sobre o equilíbrio ecológico (Błędzka et al., 2014). A
sua potencial atividade estrogênica é associada com o tamanho e o peso molecular de
sua cadeia alquila ou arila, de forma que o butilparabeno (BuP), o heptilparabeno (HeP)
e o benzilparabeno (BzP) possuem um maior potencial de desregulação endócrina
(Routledge et al., 1998). O 4-HB é considerado o metabólito final da degradação biótica
dos parabenos e também possui atividade estrogênica relatada in vitro e in vivo
(Błędzka et al., 2014; Raja et al., 2019). Geração de estresse oxidativo,

30
embriotoxicidade, genotoxicidade, neurotoxicidade, desregulação endócrina, mudanças
funcionais na microbiota intestinal e alterações metabólicas, histopatológicas e
comportamentais também são possíveis efeitos deletérios destes compostos, de
acordo com diversos estudos experimentais conduzidos em organismos como peixes
teleósteos e crustáceos aquáticos (Merola et al., 2020; Penha et al., 2021; Lin et al.,
2022; Eghan et al., 2023; Hu et al., 2023b).
Em 2017, Ateş et al. demonstraram a ocorrência de embriotoxicidade e
estresse oxidativo em embriões de zebrafish (Danio rerio) expostos a uma
concentração de 50 mg/L de MeP por 68 horas. Verificou-se aumento da mortalidade,
da incidência de efeitos teratogênicos subletais e do dano oxidativo nos embriões
expostos a 50 mg/L de MeP. Tais alterações também foram encontradas no estudo
conduzido em zebrafish por Merola et al. (2020), sendo observadas em concentrações
a partir de 30 mg/L de MeP. Dambal et al. (2017) verificaram anomalias no
desenvolvimento de embriões e larvas de zebrafish expostos a concentrações de MeP
a partir de 200 µM, tais como redução da frequência cardíaca, edema pericárdico,
curvatura anormal da coluna e acúmulo de células sanguíneas. Observou-se também
indução na expressão de vitelogenina (vtg), um biomarcador de desregulação
endócrina, em larvas expostas a concentrações de 100 µM de MeP. Em 2019, Raja et
al. observaram redução da frequência cardíaca e atraso no desenvolvimento e eclosão
de embriões de zebrafish expostos a concentrações de MeP de 10 e 100 partes por
bilhão (ppb). Em concentrações de MeP de 0,1 e 1 ppb, observou-se aumento dos
níveis de cortisol e da exibição de comportamentos de ansiedade nos animais, bem
como redução na atividade da enzima acetilcolinesterase. Bereketoglu e Pradhan
(2019) registraram efeitos deletérios da exposição de embriões de zebrafish a MeP
(100 e 200 µM) e PrP (10 e 25 µM) por 120 horas, como malformações embrionárias e
menor expressão de genes relacionados à proteção contra o estresse oxidativo, como
a superóxido dismutase 1 (sod1) e glutationa-S-transferase (gst). Também verificaram
expressão alterada de genes relacionados à função endócrina, com aumento da
expressão do receptor de estrógeno 2-α (esr2a) e redução do receptor de andrógeno
(ar). Adicionalmente, Penha et al. (2021) verificaram em zebrafish adultos, expostos a
50 mg/L de metilparabeno por 96 horas, o aumento da peroxidação lipídica, um

31
biomarcador

de

dano

oxidativo,

e

de micronúcleos , um biomarcador de

genotoxicidade.
O zebrafish como organismo-modelo
O Danio rerio (Hamilton 1822), popularmente conhecido como zebrafish, é
um peixe teleósteo de água doce da família dos ciprinídeos, nativo do sudeste asiático
e do subcontinente indiano (Ribas e Piferrer, 2013). Há mais de 100 anos tem sido
empregado como modelo bem estabelecido em pesquisas experimentais, tendo uma
crescente importância em estudos de diversas áreas no Brasil (Trigueiro et al., 2020). É
utilizado como um como um organismo-modelo para estudos de embriotoxicidade,
teratogenicidade, neurotoxicidade, genotoxicidade, estresse oxidativo e alterações
morfológicas e comportamentais devido a diversos fatores (Ribas e Piferrer, 2013;
Roper e Tanguay, 2018). Os estudos realizados nesta espécie de peixe são de
crescente importância para a avaliação de possíveis consequências da exposição de
vertebrados a contaminantes encontrados em águas superficiais (Busch et al., 2011;
Garcia et al., 2016). Embriões e larvas de zebrafish, devido à sua transparência,
também têm sido cada vez mais utilizados em estudos de localização com MPs e NPs
marcados por fluorocromos (Bhagat et al., 2020).
Em primeiro lugar, o genoma do zebrafish foi completamente sequenciado
em 2013, demonstrando ortologia de 71% em relação ao genoma humano, e os genes
relacionados a doenças chegam a 84% de homologia de sequência (Roper e Tanguay,
2018). A morfologia e fisiologia da espécie possuem similaridades significativas em
relação às de outros vertebrados, bem como o seu desenvolvimento embrionário —
que é altamente conservado neste subfilo e, parcialmente, em alguns outros cordados
(Roper e Tanguay, 2018).
Além disso, o desenvolvimento dos embriões e larvas é de fácil observação
devido à transparência dos ovos e do epitélio dos indivíduos, que permanece até sete
dias após a fertilização. Além disso, seu processo de desenvolvimento é bem
conhecido e documentado, o que torna a observação e análise de possíveis alterações
mais simples e eficiente: as fases de gastrulação, neurulação e organogênese estão
completas em cerca de 48 a 72 horas após a fertilização dos ovos, quando tem início a

32
eclosão. Em 72 a 96 horas após a fertilização, inicia-se o período larval (Kimmel et al.,
1995).
Os peixes adultos são de pequeno tamanho (3 a 4 cm de comprimento),
sendo possível manter uma grande quantidade de animais em um pequeno espaço e
sem a necessidade de estrutura laboratorial de alta complexidade. Oferecem também
ampla disponibilidade de embriões, com fêmeas capazes de colocar centenas de ovos
a cada evento reprodutivo, que podem ocorrer mais de uma vez por semana. Além
disso, os animais possuem um curto tempo geracional, alcançando a maturidade
reprodutiva após cerca de três a quatro meses de vida (Roper e Tanguay, 2018). Por
todas essas características, o zebrafish se apresenta como um modelo animal
vantajoso em relação à questão de praticidade, facilidade de manejo e viabilidade
econômica dos estudos, além de ter alto valor translacional em relação a outras
espécies de vertebrados (Kalueff et al., 2013; Ribas e Piferrer, 2013).

33
Objetivos e hipóteses
Tendo em vista a revisão da literatura feita ao longo desta seção, tem-se
como objetivo geral do trabalho investigar como a presença de microplásticos de
polietileno (MPPE) afeta a toxicidade de metilparabeno (MeP) sobre os estágios iniciais
de desenvolvimento do zebrafish.
Os objetivos específicos, por sua vez, são:
1. Realizar uma ampla revisão da literatura acerca da bioacumulação e
toxicidade de parabenos e seus metabólitos sobre organismos aquáticos de múltiplos
táxons;
2. Avaliar a toxicidade de três concentrações ambientalmente relevantes de
MeP (0,01 µM, 0,1 µM e 1 µM) de forma isolada ou em mistura com MPPE (3,4 mg/L)
sobre embriões e larvas de zebrafish a partir do Teste de Toxicidade Embriolarval
(ZELT).

34
REFERÊNCIAS
AVEN, Terje. The cautionary principle in risk management: foundation and practical use.
Reliability Engineering & System Safety,, v. 191, p. 106585, nov. 2019.
http://dx.doi.org/10.1016/j.ress.2019.106585.
AYARI, Tahani El; AHMED, Raja Ben; BOURIGA, Nawzet; GRAVATO, Carlos; CHELBI,
Emna; NECHI, Salwa; MENIF, Najoua Trigui El. Florfenicol induces malformations of
embryos and causes altered lipid profile, oxidative damage, neurotoxicity, and
histological effects on gonads of adult sea urchin, Paracentrotus lividus. Environmental
Toxicology
And
Pharmacology,
v.
110,
p.
104533,
set.
2024.
http://dx.doi.org/10.1016/j.etap.2024.104533.
BARBOZA, Luís Gabriel Antão; VIEIRA, Luís Russo; BRANCO, Vasco; FIGUEIREDO,
Neusa; CARVALHO, Felix; CARVALHO, Cristina; GUILHERMINO, Lúcia. Microplastics
cause neurotoxicity, oxidative damage and energy-related changes and interact with the
bioaccumulation of mercury in the European seabass, Dicentrarchus labrax (Linnaeus,
1758).
Aquatic
Toxicology,
v.
195,
p.
49-57,
fev.
2018.
http://dx.doi.org/10.1016/j.aquatox.2017.12.008.
BELMAKER, Ilana; ANCA, Evelyn D.; RUBIN, Lisa P.; MAGEN-MOLHO, Hadas;
MIODOVNIK, Anna; HAL, Noam van Der. Adverse health effects of exposure to plastic,
microplastics and their additives: environmental, legal and policy implications for Israel.
Israel Journal Of Health Policy Research, v. 13, n. 1, p. 1-25, 10 set. 2024.
http://dx.doi.org/10.1186/s13584-024-00628-6.
BHAGAT, Jacky; ZANG, Liqing; NISHIMURA, Norihiro; SHIMADA, Yasuhito. Zebrafish:
an emerging model to study microplastic and nanoplastic toxicity. Science Of The Total
Environment,
v.
728,
p.
138707,
ago.
2020.
http://dx.doi.org/10.1016/j.scitotenv.2020.138707.
BIERBAUM, Rosina; LEONARD, Sunday A.; REJESKI, David; WHALEY, Christopher;
BARRA, Ricardo O.; LIBRE, Christina. Novel entities and technologies: environmental
benefits and risks. Environmental Science & Policy, v. 105, p. 134-143, mar. 2020.
http://dx.doi.org/10.1016/j.envsci.2019.11.002.
BIHANIC, Florane Le; CLÉRANDEAU, Christelle; CORMIER, Bettie; CREBASSA,
Jean-Claude; KEITER, Steffen H.; BEIRAS, Ricardo; MORIN, Bénédicte; BÉGOUT,
Marie-Laure; COUSIN, Xavier; CACHOT, Jérôme. Organic contaminants sorbed to
microplastics affect marine medaka fish early life stages development. Marine
Pollution
Bulletin,
v.
154,
p.
111059,
maio
2020..

35
http://dx.doi.org/10.1016/j.marpolbul.2020.111059.
BŁĘDZKA, Dorota; GROMADZIńSKA, Jolanta; WĄSOWICZ, Wojciech. Parabens. From
environmental studies to human health. Environment International, v. 67, p. 27-42,
jun. 2014. http://dx.doi.org/10.1016/j.envint.2014.02.007.
BOLUJOKO, Nathaniel B.; UNUABONAH, Emmanuel I.; ALFRED, Moses O.;
OGUNLAJA, Aemere; OGUNLAJA, Olumuyiwa O.; OMOROGIE, Martins O.;
OLUKANNI, Olumide D.. Toxicity and removal of parabens from water: a critical review.
Science Of The Total Environment, v. 792, p. 148092, out. 2021.
http://dx.doi.org/10.1016/j.scitotenv.2021.148092.
BUSCH, Wibke; DUIS, Karen; FENSKE, Martina; MAACK, Gerd; LEGLER, Juliette;
PADILLA, Stephanie; STRÄHLE, Uwe; WITTERS, Hilda; SCHOLZ, Stefan. The
zebrafish embryo model in toxicology and teratology, September 2–3, 2010, Karlsruhe,
Germany. Reproductive Toxicology, v. 31, n. 4, p. 585-588, maio 2011.
http://dx.doi.org/10.1016/j.reprotox.2011.02.010.DAS, Shubhasikha; PARIDA, Vishal
Kumar; TIWARY, Chandra Sekhar; GUPTA, Ashok Kumar; CHOWDHURY, Shamik.
Emerging Contaminants in the Aquatic Environment: fate, occurrence, impacts, and
toxicity.
Acs
Symposium
Series,
p.
1-32,
16
set.
2024.
http://dx.doi.org/10.1021/bk-2024-1475.ch001.
CAROLIN C, Femina; KAMALESH, T.; KUMAR, P. Senthil; RANGASAMY, Gayathri. An
insights of organochlorine pesticides categories, properties, eco-toxicity and new
developments in bioremediation process. Environmental Pollution, v. 333, p. 122114,
set. 2023. http://dx.doi.org/10.1016/j.envpol.2023.122114.
CHASSIGNET, Eric P.; XU, Xiaobiao; ZAVALA-ROMERO, Olmo. Tracking Marine Litter
With a Global Ocean Model: where does it go? Where does it come from?. Frontiers In
Marine Science, v. 8, p. 1-15, 23 abr. 2021. Frontiers Media SA.
http://dx.doi.org/10.3389/fmars.2021.667591.
DAI, Yanhui; ZHAO, Jian; SUN, Chunxiao; LI, Diying; LIU, Xia; WANG, Zhenyu; YUE,
Tongtao; XING, Baoshan. Interaction and combined toxicity of microplastics and perand polyfluoroalkyl substances in aquatic environment. Frontiers Of Environmental
Science & Engineering, v. 16, n. 10, p. 1-17, 26 abr. 2022.
http://dx.doi.org/10.1007/s11783-022-1571-2.
DOURDIN, Thomas Sol; RIVIÈRE, Guillaume; CORMIER, Alexandre; POI, Carole di;
GUYOMARD, Killian; RABILLER, Manuella; AKCHA, Farida; SADIALIOU, Thierno Bah;
MONIER, Pauline Le; SUSSARELLU, Rossana. Molecular and phenotypic effects of early

36
exposure to an environmentally relevant pesticide mixture in the Pacific oyster,
Crassostrea gigas. Environmental Pollution, v. 326, p. 121472, jun. 2023.
http://dx.doi.org/10.1016/j.envpol.2023.121472.
EGHAN, Kojo; LEE, Sangwoo; KIM, Woo-Keun. Cardio- and neuro-toxic effects of four
parabens on Daphnia magna. Ecotoxicology And Environmental Safety, v. 268, p.
115670, dez. 2023. http://dx.doi.org/10.1016/j.ecoenv.2023.115670.
ESCOBAR-HUERFANO,
Francisco;
GÓMEZ-OLIVÁN,
Leobardo
Manuel;
LUJA-MONDRAGÓN, Marlenee; SANJUAN-REYES, Nely; ISLAS-FLORES, Hariz;
HERNÁNDEZ-NAVARRO, María Dolores. Embryotoxic and teratogenic profile of
tetracycline at environmentally relevant concentrations on Cyprinus carpio.
Chemosphere,
v.
240,
p.
124969,
fev.
2020.
http://dx.doi.org/10.1016/j.chemosphere.2019.124969.
FANG, Song; YU, Weisong; LI, Chengliang; LIU, Yuedong; QIU, Jun; KONG, Fanyu.
Adsorption behavior of three triazole fungicides on polystyrene microplastics. Science
Of
The
Total
Environment,
v.
691,
p.
1119-1126,
nov.
2019.
http://dx.doi.org/10.1016/j.scitotenv.2019.07.176.
FÉLIX, Luís; CARREIRA, Paulo; PEIXOTO, Francisco. Effects of chronic exposure of
naturally weathered microplastics on oxidative stress level, behaviour, and mitochondrial
function of adult zebrafish (Danio rerio). Chemosphere, v. 310, p. 136895, jan. 2023.
http://dx.doi.org/10.1016/j.chemosphere.2022.136895.
GARCIA, Gloria R.; NOYES, Pamela D.; TANGUAY, Robert L.. Advancements in
zebrafish applications for 21st century toxicology. Pharmacology & Therapeutics, v.
161, p. 11-21, maio 2016. http://dx.doi.org/10.1016/j.pharmthera.2016.03.009.
GARRIC, Jeanne. Emerging Issues in Ecotoxicology: Pharmaceuticals and Personal
Care Products (PPCPs). In: FÉRARD, Jean-François; BLAISE, Christian (ed.).
Encyclopedia of Aquatic Ecotoxicology. Dordrecht: Springer, 2013. Cap. 38. p.
407-428.
HAMAN, Camille; DAUCHY, Xavier; ROSIN, Christophe; MUNOZ, Jean-François.
Occurrence, fate and behavior of parabens in aquatic environments: a review. Water
Research, v. 68, p. 1-11, jan. 2015. http://dx.doi.org/10.1016/j.watres.2014.09.030.
HASAN, A K M Munzurul; HAMED, Mohamed; HASAN, Jabed; MARTYNIUK,
Christopher J.; NIYOGI, Som; CHIVERS, Douglas P.. A review of the neurobehavioural,

37
physiological, and reproductive toxicity of microplastics in fishes. Ecotoxicology And
Environmental
Safety,
v.
282,
p.
116712,
set.
2024.
http://dx.doi.org/10.1016/j.ecoenv.2024.116712.
HU, Baiyang; LI, Yanxia; JIANG, Linshu; CHEN, Xingcai; WANG, Lin; AN, Siyu;
ZHANG, Fengsong. Influence of microplastics occurrence on the adsorption of
17β-estradiol in soil. Journal Of Hazardous Materials, v. 400, p. 123325, dez. 2020.
http://dx.doi.org/10.1016/j.jhazmat.2020.123325.
KALUEFF, Allan V.; STEWART, Adam Michael; GERLAI, Robert. Zebrafish as an
emerging model for studying complex brain disorders. Trends In Pharmacological
Sciences, v. 35, n. 2, p. 63-75, fev. 2014. http://dx.doi.org/10.1016/j.tips.2013.12.002.
HU, Chenyan; BAI, Yachen; LI, Jing; SUN, Baili; CHEN, Lianguo. Endocrine disruption
and reproductive impairment of methylparaben in adult zebrafish. Food And Chemical
Toxicology, v. 171, p. 113545, jan. 2023. http://dx.doi.org/10.1016/j.fct.2022.113545.
KIMMEL, Charles B.; BALLARD, William W.; KIMMEL, Seth R.; ULLMANN, Bonnie;
SCHILLING, Thomas F.. Stages of embryonic development of the zebrafish.
Developmental
Dynamics,
v.
203,
n.
3,
p.
253-310,
jul.
1995.
http://dx.doi.org/10.1002/aja.1002030302.
LEE, Minha; KIM, Heejung; RYU, Han-Sun; MOON, Jinah; KHANT, Naing Aung; YU,
Chaerim; YU, Ji-Hee. Review on invasion of microplastic in our ecosystem and
implications. Science Progress, v. 105, n. 4, p. 1-23, out. 2022.
http://dx.doi.org/10.1177/00368504221140766.
LI, Jia; ZHANG, Kaina; ZHANG, Hua. Adsorption of antibiotics on microplastics.
Environmental
Pollution,
v.
237,
p.
460-467,
jun.
2018.
http://dx.doi.org/10.1016/j.envpol.2018.02.050.
LI, Hui; WANG, Fenghe; LI, Jining; DENG, Shaopo; ZHANG, Shengtian. Adsorption of
three pesticides on polyethylene microplastics in aqueous solutions: kinetics, isotherms,
thermodynamics, and molecular dynamics simulation. Chemosphere, v. 264, p.
128556, fev. 2021. http://dx.doi.org/10.1016/j.chemosphere.2020.128556.
LI, Xingyu; SHEN, Xiaojing; JIANG, Weiwei; XI, Yongkai; LI, Song. Comprehensive review
of emerging contaminants: detection technologies, environmental impact, and
management strategies. Ecotoxicology And Environmental Safety, v. 278, p. 116420,
jun. 2024. http://dx.doi.org/10.1016/j.ecoenv.2024.116420.

38
LIMONTA, Giacomo; MANCIA, Annalaura; BENKHALQUI, Assja; BERTOLUCCI, Cristiano;
ABELLI, Luigi; FOSSI, Maria Cristina; PANTI, Cristina. Microplastics induce
transcriptional changes, immune response and behavioral alterations in adult zebrafish.
Scientific
Reports,
v.
9,
n.
1,
p.
1-11,
31
out.
2019.
http://dx.doi.org/10.1038/s41598-019-52292-5.
LIN, Hongxing; JIA, Yongyi; HAN, Fenglu; XIA, Chuyan; ZHAO, Qun; ZHANG, Jiliang;
LI, Erchao. Toxic effects of waterborne benzylparaben on the growth, antioxidant
capacity and lipid metabolism of Nile tilapia (Oreochromis niloticus). Aquatic
Toxicology,
v.
248,
p.
106197,
jul.
2022.
http://dx.doi.org/10.1016/j.aquatox.2022.106197.
LIU, Fei-Fei; LIU, Guang-Zhou; ZHU, Zhi-Lin; WANG, Su-Chun; ZHAO, Fei-Fei.
Interactions between microplastics and phthalate esters as affected by microplastics
characteristics and solution chemistry. Chemosphere, v. 214, p. 688-694, jan. 2019.
http://dx.doi.org/10.1016/j.chemosphere.2018.09.174.
LIU, Wen; SALMOND, Jennifer; ALLEN, Kristiann; TADAKI, Marc. Managing emerging
environmental risks when we do not know enough about them: the case of respirable
mineral dust. Environmental Science & Policy, v. 155, p. 103715, maio 2024.
http://dx.doi.org/10.1016/j.envsci.2024.103715.
LIU, Huanpeng; LI, Huiqi; LIU, Yinai; ZHAO, Haiyang; PENG, Renyi. Toxic effects of
microplastic and nanoplastic on the reproduction of teleost fish in aquatic environments.
Environmental
Science
And
Pollution
Research,
out.
2024.
http://dx.doi.org/10.1007/s11356-024-35434-9.
MALAFAIA, Guilherme; SOUZA, Andreza Martins de; PEREIRA, Aryelle Canedo;
GONÇALVES, Stênio; ARAÚJO, Amanda Pereira da Costa; RIBEIRO, Renan Xavier;
ROCHA, Thiago Lopes. Developmental toxicity in zebrafish exposed to polyethylene
microplastics under static and semi-static aquatic systems. Science Of The Total
Environment,
v.
700,
p.
134867,
jan.
2020.
http://dx.doi.org/10.1016/j.scitotenv.2019.134867.
MARTA FAVA. UNESCO — Ocean Literacy Portal. Ocean plastic pollution, an overview:
data
and
statistics.
2022.
Disponível
em:
https://oceanliteracy.unesco.org/plastic-pollution-ocean/. Acesso em: 06 nov. 2024.
QIANG, Liyuan; CHENG, Jinping. Exposure to polystyrene microplastics impairs gonads
of zebrafish (Danio rerio). Chemosphere, v. 263, p. 128161, jan. 2021.
http://dx.doi.org/10.1016/j.chemosphere.2020.128161.

39
MEROLA, C.; PERUGINI, M.; CONTE, A.; ANGELOZZI, G.; BOZZELLI, M.; AMORENA, M..
Embryotoxicity of methylparaben to zebrafish (Danio rerio) early-life stages.
Comparative Biochemistry And Physiology Part C: Toxicology & Pharmacology, v. 236,
p. 108792, out. 2020. http://dx.doi.org/10.1016/j.cbpc.2020.108792.
NGOUBEYOU, Prisca Stephanie Kandjo; WOLKERSDORFER, Christian; NDIBEWU, Peter
Papoh; AUGUSTYN, Wilma. Toxicity of polychlorinated biphenyls in aquatic
environments – A review. Aquatic Toxicology, v. 251, p. 106284, out. 2022.
http://dx.doi.org/10.1016/j.aquatox.2022.106284.
NOWAK, Karolina; RATAJCZAK–WRONA, Wioletta; GÓRSKA, Maria; JABłOŃSKA, Ewa.
Parabens and their effects on the endocrine system. Molecular And Cellular
Endocrinology,
v.
474,
p.
238-251,
out.
2018.
http://dx.doi.org/10.1016/j.mce.2018.03.014.
PENHA, Larissa Cristine de Carvalho; ROLA, Regina Coimbra; SILVA JUNIOR, Flávio
Manoel da; MARTINS, Camila de Martinez Gaspar. Toxicity and sublethal effects of
methylparaben on zebrafish (Danio rerio) larvae and adults. Environmental Science And
Pollution Research, v. 28, n. 33, p. 45534-45544, 18 abr. 2021.
http://dx.doi.org/10.1007/s11356-021-12800-5.
PEREIRA, Ana Rita; SIMÕES, Manuel; GOMES, Inês B.. Parabens as environmental
contaminants of aquatic systems affecting water quality and microbial dynamics.
Science Of The Total Environment, v. 905, p. 167332, dez. 2023.
http://dx.doi.org/10.1016/j.scitotenv.2023.167332.
PLASTICSEUROPE. An analysis of European plastics production, demand and waste
data. Plastics — The facts. Association of Plastic Manufacturers. Bruxelas, Bélgica, 64
pp, 2020. Disponível em: https://www.plasticseurope.org/en/resources/market-data.
Acesso em: 17 out 2021.
RAFA, Nazifa; AHMED, Bushra; ZOHORA, Fatema; BAKYA, Jannatul; AHMED,
Samiya; AHMED, Shams Forruque; MOFIJUR, M.; CHOWDHURY, Ashfaque Ahmed;
ALMOMANI, Fares. Microplastics as carriers of toxic pollutants: source, transport, and
toxicological effects. Environmental Pollution, v. 343, p. 123190, fev. 2024.
http://dx.doi.org/10.1016/j.envpol.2023.123190.
RIBAS, Laia; PIFERRER, Francesc. The zebrafish (Danio rerio) as a model organism,
with emphasis on applications for finfish aquaculture research. Reviews In

40
Aquaculture, v. 6, n. 4, p. 209-240, 31 jul. 2013. http://dx.doi.org/10.1111/raq.12041.
RAJA, Glancis Luzeena; SUBHASHREE, K. Divya; LITE, Christy; SANTOSH, Winkins;
BARATHI, S.. Transient exposure of methylparaben to zebrafish (Danio rerio) embryos
altered cortisol level, acetylcholinesterase activity and induced anxiety-like behaviour.
General And Comparative Endocrinology, v. 279, p. 53-59, ago. 2019.
http://dx.doi.org/10.1016/j.ygcen.2018.11.001.
REZANIA, Shahabaldin; PARK, Junboum; DIN, Mohd Fadhil Md; TAIB, Shazwin Mat;
TALAIEKHOZANI, Amirreza; YADAV, Krishna Kumar; KAMYAB, Hesam. Microplastics
pollution in different aquatic environments and biota: a review of recent studies. Marine
Pollution
Bulletin,
v.
133,
p.
191-208,
ago.
2018.
http://dx.doi.org/10.1016/j.marpolbul.2018.05.022.
RICHARDSON, Katherine; STEFFEN, Will; LUCHT, Wolfgang; BENDTSEN, Jørgen;
CORNELL, Sarah E.; DONGES, Jonathan F.; DRÜKE, Markus; FETZER, Ingo; BALA,
Govindasamy; VON BLOH, Werner. Earth beyond six of nine planetary boundaries.
Science
Advances,
v.
9,
n.
37,
p.
1-16,
15
set.
2023.
http://dx.doi.org/10.1126/sciadv.adh2458.
ROCKSTRÖM, Johan; STEFFEN, Will; NOONE, Kevin; PERSSON, Åsa; CHAPIN, F. Stuart;
LAMBIN, Eric F.; LENTON, Timothy M.; SCHEFFER, Marten; FOLKE, Carl;
SCHELLNHUBER, Hans Joachim. A safe operating space for humanity. Nature, v. 461,
n. 7263, p. 472-475, set. 2009. http://dx.doi.org/10.1038/461472a.
ROJE, Željka; ILIĆ, Krunoslav; GALIĆ, Emerik; PAVIČIĆ, Ivan; TURČIĆ, Petra;
STANEC, Zdenko; VRČEK, Ivana Vinković. Synergistic effects of parabens and plastic
nanoparticles on proliferation of human breast cancer cells. Archives Of Industrial
Hygiene And Toxicology, v. 70, n. 4, p. 310-314, 1 dez. 2019.
http://dx.doi.org/10.2478/aiht-2019-70-3372.
ROPER, Courtney; TANGUAY, Robert L.. Zebrafish as a Model for Developmental
Biology and Toxicology. Handbook Of Developmental Neurotoxicology, p. 143-151,
2018. http://dx.doi.org/10.1016/b978-0-12-809405-1.00012-2.
ROUTLEDGE, Edwin J.; PARKER, Joanne; ODUM, Jenny; ASHBY, John; SUMPTER,
John P.. Some Alkyl Hydroxy Benzoate Preservatives (Parabens) Are Estrogenic.
Toxicology And Applied Pharmacology, v. 153, n. 1, p. 12-19, nov. 1998.
http://dx.doi.org/10.1006/taap.1998.8544.

41
ROWLAND, F. Sherwood. Earth's changing atmosphere: chlorofluorocarbons and
ozone. Environmental Impact Assessment Review, v. 10, n. 4, p. 359-370, dez.
1990. http://dx.doi.org/10.1016/0195-9255(90)90028-x.
SALLA, Raquel Fernanda; OLIVEIRA, Fagner Neves; JACINTHO, Jaqueline C.;
CIRQUEIRA, Felipe; TSUKADA, Elisabete; VIEIRA, Lucélia Gonçalves; ROCHA,
Thiago Lopes. Microplastics and TiO2 nanoparticles mixture as an emerging threat to
amphibians: a case study on bullfrog embryos. Environmental Pollution, v. 346, p.
123624, abr. 2024. http://dx.doi.org/10.1016/j.envpol.2024.123624.
SANTOS, Mauricius Marques dos; BREHM, Franciane de Almeida; FILIPPE, Tais
Cristina; KNAPIK, Heloise Garcia; AZEVEDO, Júlio César Rodrigues de. Occurrence
and risk assessment of parabens and triclosan in surface waters of southern Brazil: a
problem of emerging compounds in an emerging country. Rbrh, v. 21, n. 3, p. 603-617,
set. 2016. http://dx.doi.org/10.1590/2318-0331.011616018.
SEQUEIRA, Inês F.; PRATA, Joana C.; COSTA, João P. da; DUARTE, Armando C.;
ROCHA-SANTOS, Teresa. Worldwide contamination of fish with microplastics: a brief
global overview. Marine Pollution Bulletin, v. 160, p. 111681, nov. 2020.
http://dx.doi.org/10.1016/j.marpolbul.2020.111681.
SINGH, Nisha; KHANDELWAL, Nitin; TIWARI, Ekta; NASKAR, Nabanita; LAHIRI,
Susanta; LÜTZENKIRCHEN, Johannes; DARBHA, Gopala Krishna. Interaction of metal
oxide nanoparticles with microplastics: impact of weathering under riverine conditions.
Water
Research,
v.
189,
p.
116622,
fev.
2021.
http://dx.doi.org/10.1016/j.watres.2020.116622.
STANTON, Thomas; JOHNSON, Matthew; NATHANAIL, Paul; MACNAUGHTAN,
William; GOMES, Rachel L.. Freshwater microplastic concentrations vary through both
space and time. Environmental Pollution, v. 263, p. 114481, ago. 2020.
http://dx.doi.org/10.1016/j.envpol.2020.114481.
TAN, Xiangling; YU, Xubiao; CAI, Liqi; WANG, Jundong; PENG, Jinping. Microplastics
and associated PAHs in surface water from the Feilaixia Reservoir in the Beijiang River,
China.
Chemosphere,
v.
221,
p.
834-840,
abr.
2019.
http://dx.doi.org/10.1016/j.chemosphere.2019.01.022.
TREVISAN, Rafael; UZOCHUKWU, Daniel; GIULIO, Richard T. di. PAH Sorption to
Nanoplastics and the Trojan Horse Effect as Drivers of Mitochondrial Toxicity and PAH
Localization in Zebrafish. Frontiers In Environmental Science, v. 8, n. 78, p. 1-26, 24
jul. 2020. Frontiers Media SA. http://dx.doi.org/10.3389/fenvs.2020.00078.

42
TRIGUEIRO, Nicholas Silvestre de Souza; CANEDO, Aryelle; BRAGA, Daniel Lôbo de
Siqueira; LUCHIARI, Ana Carolina; ROCHA, Thiago Lopes. Zebrafish as an Emerging
Model System in the Global South: two decades of research in brazil. Zebrafish, v. 17,
n. 6, p. 412-425, 1 dez. 2020. http://dx.doi.org/10.1089/zeb.2020.1930.
United Nations Environmental Programme. Stockholm Convention on Persistent
Organic
Pollutants
(POPS).
2023.
Disponível
em:
https://www.pops.int/Convention/ConventionText/tabid/2232/Default.aspx. Acesso em:
28 set. 2024.
URBANEK, Aneta K.; RYMOWICZ, Waldemar; MIROńCZUK, Aleksandra M..
Degradation of plastics and plastic-degrading bacteria in cold marine habitats. Applied
Microbiology And Biotechnology, v. 102, n. 18, p. 7669-7678, 11 jul. 2018.
http://dx.doi.org/10.1007/s00253-018-9195-y.
WANG, Jun; LI, Yuejiao; LU, Lin; ZHENG, Mingyi; ZHANG, Xiaona; TIAN, Hua; WANG,
Wei; RU, Shaoguo. Polystyrene microplastics cause tissue damages, sex-specific
reproductive disruption and transgenerational effects in marine medaka (Oryzias
melastigma). Environmental Pollution, v. 254, p. 113024, nov. 2019.
http://dx.doi.org/10.1016/j.envpol.2019.113024.
WANG, Ting; HU, Menghong; SONG, Lili; YU, Jun; LIU, Ruijuan; WANG, Shixiu;
WANG, Zhifu; SOKOLOVA, Inna M.; HUANG, Wei; WANG, Youji. Coastal zone use
influences the spatial distribution of microplastics in Hangzhou Bay, China.
Environmental
Pollution,
v.
266,
p.
115137,
nov.
2020.
http://dx.doi.org/10.1016/j.envpol.2020.115137.
WEI, Fang; MORTIMER, Monika; CHENG, Hefa; SANG, Nan; GUO, Liang-Hong.
Parabens as chemicals of emerging concern in the environment and humans: a review.
Science Of The Total Environment, v. 778, p. 146150, jul. 2021.
http://dx.doi.org/10.1016/j.scitotenv.2021.146150.
XUE, Jingchuan; KANNAN, Kurunthachalam. Accumulation profiles of parabens and
their metabolites in fish, black bear, and birds, including bald eagles and albatrosses.
Environment
International,
v.
94,
p.
546553,
set.
2016.
http://dx.doi.org/10.1016/j.envint.2016.06.015.
XUE, Xiaohong; XUE, Jingchuan; LIU, Wenbin; ADAMS, Douglas H.; KANNAN,
Kurunthachalam. Trophic Magnification of Parabens and Their Metabolites in a
Subtropical Marine Food Web. Environmental Science & Technology, v. 51, n. 2, p.
780-789,
5
jan.
2017.
American
Chemical
Society
(ACS).
http://dx.doi.org/10.1021/acs.est.6b05501.

43
YIN, Liyun; CHEN, Bijuan; XIA, Bin; SHI, Xiaotao; QU, Keming. Polystyrene
microplastics alter the behavior, energy reserve and nutritional composition of marine
jacopever (Sebastes schlegelii). Journal Of Hazardous Materials, v. 360, p. 97-105,
out. 2018. http://dx.doi.org/10.1016/j.jhazmat.2018.07.110.
YONG, Cheryl; VALIYAVEETTIL, Suresh; TANG, Bor. Toxicity of Microplastics and
Nanoplastics in Mammalian Systems. International Journal Of Environmental
Research And Public Health, v. 17, n. 5, p. 1509, 26 fev. 2020.
http://dx.doi.org/10.3390/ijerph17051509.
ZHOU, Ranran; LU, Guanghua; YAN, Zhenhua; JIANG, Runren; BAO, Xuhui; LU, Ping.
A review of the influences of microplastics on toxicity and transgenerational effects of
pharmaceutical and personal care products in aquatic environment. Science Of The
Total
Environment,
v.
732,
p.
139222,
ago.
2020.
http://dx.doi.org/10.1016/j.scitotenv.2020.139222.
ZHOU, Qian; ZHANG, Haibo; FU, Chuancheng; ZHOU, Yang; DAI, Zhenfei; LI, Yuan;
TU, Chen; LUO, Yongming. The distribution and morphology of microplastics in coastal
soils adjacent to the Bohai Sea and the Yellow Sea. Geoderma, v. 322, p. 201-208, jul.
2018. http://dx.doi.org/10.1016/j.geoderma.2018.02.015.

44
CAPÍTULO 1
BIOACCUMULATION AND ECOTOXICITY OF PARABENS IN AQUATIC
ORGANISMS: CURRENT STATUS AND TRENDS
Felipe Felix Costa Lima da Silveira

Published in the journal Environmental Pollution on November 4th, 2024.
DOI: https://doi.org/10.1016/j.envpol.2024.125213

Abstract
Parabens are preservatives commonly found in various personal care products,
pharmaceuticals, and foodstuffs and mostly are unregulated chemical compounds.
Given their extensive use, they are ubiquitously detected in different compartments,
including aquatic environments, and classified as emerging pollutants. When parabens
reach aquatic environments, they may compromise animal health. Thus, the current
study aimed to review the data available concerning the bioaccumulation and ecotoxic
effects of parabens on aquatic species. After the review search, which included articles
published online around the world until December 2023, a total of 71 articles were
systematically analyzed and summarized. The first study on paraben ecotoxicity was
published in 2000. Studies were conducted mainly in laboratory conditions (80.28%)
using fish, crustaceans, bivalves, algae and bacteria, among others taxa. Field studies
were conducted at 82 sampling sites across five countries. Paraben bioaccumulation
was detected primarily in fish muscle, liver, brain, gills, and testis. Also, aquatic animals
(i.e., fish and invertebrates) were more susceptible to the effects of parabens than
microorganisms such as periphyton. Parabens can cause lethal and sublethal effects on
aquatic organisms, such as oxidative stress, endocrine disruption, neurotoxicity,
behavioral changes, reproductive impairment, and developmental abnormalities. In
addition, the toxicity of parabens depends on the species, taxon, developmental stage,
exposure period, and concentrations tested. As long as its use remains massive and its
detection ubiquitously, our literature overview denotes that further ecotoxicological
research about the parents' parabens and their metabolites in different taxa of aquatic
organisms is deeply needed.
Keywords: aquatic pollution; ecotoxicology; endocrine
p-Hydroxybenzoic Acid; sentinel species; zebrafish

1.1 Introduction

disruption;

Esters

of

45
Parabens are a family of organic compounds broadly used as preservatives
in pharmaceuticals, personal care products (PCPs), and foodstuffs due to their
antimicrobial (Terasaki et al., 2013) and fungicidal (Murata et al., 2019) activities. Also,
these substances have favorable properties as colorless, odorless, active at various pH
levels, and mix well with other ingredients. Parabens were produced with the intention of
replacing the use of salicylic acid and benzoic acid, which had the disadvantage of
being effective only at a highly acidic pH (Lück & Jager, 1997). Therefore, since the
1920s, parabens have become common additives in cosmetics and pharmaceutical
products (Wei et al., 2021; Nowak et al., 2018). Additionally, the worldwide cosmetics
market, worth about 500 billion EUR in 2018, is expected to increase the use of
parabens and other preservatives in response to consumer demand for longer shelf life
(Nowak et al., 2021).
Parabens are produced in large volumes in Europe, the USA, and Asia, with
production rates as high as 5000 tons reported in the 1990s (Nowak et al., 2018). In the
United States, the estimated average daily total exposure to parabens is as follows: 50
mg from personal care products (PCPs), 25 mg from pharmaceutical products, and 1
mg from food products, resulting in a cumulative exposure of 76 mg/day. For an
individual weighing 70 kg, this corresponds to a daily intake of 1.26 mg kg-1 body weight
(bw) (Vale et al., 2022; Błędzka et al., 2014). In China, monthly production of parabens
is reported to be 500 tons for methylparaben (MeP) and ethylparaben (EtP) and over
10,000 tons for propylparaben (PrP), which can be attributed to the approval in March
2002 of these parabens as food additives, including a maximum allowed concentration
of 0.5 g kg-1 in food (Ministry of Health of the People’s Republic of China, 2015).
These compounds are chemically characterized as alkyl esters of the
p-hydroxybenzoic acid (pHBA) or 4-hydroxybenzoic acid (4-HB) (Fig. 1), such as
methylparaben (MeP), ethylparaben (EtP), n-propylparaben (PrP), isopropylparaben
(iPrP),

n-butylparaben

(BuP),

isobutylparaben

(iBuP),

phenylparaben

(PhP),

benzylparaben (BzP) and heptylparaben (HeP). The MeP and PrP being the most
widely used in generic formulations, isolated or in combination (Haman et al., 2015).
Parabens' physical and chemical properties are determined by their alkyl chain length,
which is also directly related to their biological activity. For example, a longer carbon

46
chain will generally enhance their antimicrobial activity; however, a longer carbon chain
can reduce their solubility in water and can increase their persistence in aquatic
environments (Piao et al., 2014) and bioaccumulation in fatty tissues (Wang & Kannan,
2015). In addition, parabens are usually biodegraded within a few days (2.1 days in
case of MeP and EtP to 4.5 days in case of BuP), especially those with shorter alkyl
chains and under aerobic conditions, giving way to their main metabolite, 4-HB
(González-Mariño et a., 2011; Santos et al., 2016).

Figure 1. Chemical structure of parabens commonly employed in commercial and industrial
formulations and also for p-hydroxybenzoic acid.

The sewage treatment may efficiently remove several parabens from urban
sludge (Błędzka et al., 2014; Li et al., 2015); however, the regular input from multiple
sources often outpaces their removal capacity. These remaining parabens are
sometimes called a “pseudo-persistent” contaminant (Albero et al., 2012). Parabens can

47
react with chlorine, commonly used as a disinfectant in water treatment plants. These
chlorinated derivatives are more persistent than the original molecules; however, little is
known about their toxicity (Haman et al., 2015).
The growing use of PCPs and pharmaceuticals with parabens derived from
their composition has been accompanied by increased environmental concentrations,
specifically in aquatic habitats (Bolujoko et al., 2021; Brausch & Rand, 2011). The
wash-off from human bodies during personal hygiene and the unregulated discharge of
PCPs in sewage systems and regular refuse collection containers are the primary
sources of parabens in urban wastewater and overall residues (Haman et al., 2015;
Bolujoko et al., 2021). However, widespread regulation regarding their discharge, input,
and presence in shallow waters is generally lacking. Parabens are classified as
emerging pollutants (EPs) due to being detected in low concentrations (ng L-1 or µg L-1)
in aquatic environments and not being commonly included in routine monitoring of water
quality (Vale et al., 2022). However, evidence regarding their potential toxic effects on
aquatic organisms and human health raises worldwide concerns (Maia et al., 2023;
Medkova et al., 2023).
Nonetheless, it is worth noting that the widespread use of such substances is
related to their overall presumed safety to humans concerning their acute toxicity based
on concentrations typically used in commercial formulations (Crovetto et al., 2022). As
the market formulations, particularly from the cosmetic industry, and the world
population are continually growing, the concentration of parabens in shallow waters will
likely increase in the following years (Bolujoko et al., 2021). Parabens released into the
aquatic ecosystem can interact with aquatic organisms and induce ecotoxic effects at
different levels of biological organization, such as phytoplankton (Di Poi et al., 2017),
aquatic invertebrates (Lee et al., 2017; Shore et al., 2022), and fish (Lin et al., 2022).
The toxicity of parabens is related to the disruption of the cell membrane and
intracellular proteins, and the inhibition of mitochondrial function (Bolujoko et al., 2021;
Crovetto et al., 2022). In addition, recent in silico, in vitro, and in vivo studies using
zebrafish (Danio rerio) adults and also during the embryo-larval stage have revealed
that parabens exposure results in disruption of the hypothalamic-pituitary-gonad axis
(HPG),

of

the

hypothalamic-pituitary-thyroid

axis

(HPT)

and

also

for

48
hypothalamic-pituitary-adrenal axis (HPA) (Bereketoglu and Pradhan, 2019; Dambal et
al., 2017; Hu et al., 2023; Liang et al., 2022; 2023a, b). Furthermore, another in vivo
study evaluated parabens' effects on a copepod (Tigriopus japonicus) and showed that
chronic exposure to MeP, EtP, and PrP induced estrogenic effects (Kang et al., 2019).
All these data raise concerns about the effects of parabens on aquatic organisms and
ecosystems. However, understanding the mechanism of action (MoA) and ecotoxicity of
parabens in different aquatic organisms remains unclear. In this context, this review has
four main objectives:
i) Gather data on the concentrations of parabens and their metabolites found
in samples from several aquatic species in various locations worldwide;
ii) Build a historical and methodological perspective on the experimental
ecotoxicology studies with parabens, focusing on the evolution of techniques and
biomarkers since the first laboratory studies;
iii) Describe the state-of-the-art knowledge of the potential toxicity of different
types of parabens and their metabolites on aquatic organisms and;
iv) Integrate background and current experimental approach in comparison to
environmental findings;
Furthermore, research gaps and recommendations for future research were
also identified.
1.2 Methodological approach
This review was conducted from January to December 2023. After the
delimitation of the scope, keywords were expanded to capture the maximum number of
articles per search. The keywords “parabens”, in combination with “ecotoxicology”,
“biomarkers”,

“fish”,

“conservation”,

“aquatic

organisms”,

“ecology”,

“aquatic

invertebrates”, “crustaceans”, “algae”, or “mollusks,” both singular and plural forms
were used. The groups of crustaceans and mollusks were individually searched due to
the lack of studies on other invertebrate phyla/subphyla found using the keywords
“aquatic invertebrates”. The databases used were “PubMed”, “ScienceDirect”, “Web of
Science”, and “SCOPUS”. Initially, a total of 5276 articles were found in the databases.

49
After the exclusion criteria (non-English papers, gray literature, letters/short
communications, review articles, book or book chapters, duplicated documents,
articles with chemical/in vitro/in silico analysis only, papers focused on human health or
not aligned with the goal of this article in general), a total of 76 works were maintained
for further analysis (Fig. 2).

Figure 2. Summarized study methodology step by step, from the scoping to the final
presentation.

As one of the goals of our study was to picture a historical retrospective of
this field of study, the year of publication was not used as an exclusion criterion.
Instead, each paper was analyzed according to the following parameters: a) year of
publication; b) field or experimental research; c) species, sex, and developmental
stage of the organisms. Furthermore, regarding the field studies, the following
parameters were included: a) geographical correspondent authors' location; b) number

50
of species collected; c) tissues/organs investigated; d) type and concentration of
parabens and their metabolites found in samples. The bioaccumulation of parabens in
birds and aquatic mammals was not added. Finally, for the experimental studies, the
following aspects were reviewed: a) experimental design (number of samples,
treatments, and replicates); b) type of exposure (acute, chronic or subchronic
exposure; static, semi-static, or flow-through); c) route of exposure (by food, gavage, or
water exposure); d) type and concentration/dose of the parabens; e) analytical control
(if present); f) biomarkers/parameters investigated; g) omics analysis (if present, which
type); h) bioaccumulation analyses (if present); and i) biological effects found.
Graphics reporting the articles published in each country were made
through the MapChat tool (https://www.mapchart.net/). The remaining data were
compiled using Microsoft Excel and the graphics were organized according to the year
of study publication, experimental or field study, organisms' clade, life stage, potential
biomarkers, study location, and type of paraben evaluated, among others, by using the
software GraphPad PRISM®.

1.3 Historical, geographical analysis and background
From 2000 to 2023, 76 studies were published (Fig. 3A). The first published
study concerning the toxicity of parabens addressed their potential estrogenic effect
and discussed its possible outcomes for human health. Pedersen et al. (2000) tested
EtP, PrP, BuP, and their primary metabolite (4-HB) through an in vivo assay using
sexually immature rainbow trout (Oncorhynchus mykiss), this being the first
experimental evidence that exposure to parabens can cause endocrine disruption in
fish.

51

Figure 3. Number of articles published about ecotoxicological studies with parabens on
aquatic organisms and their distribution across the world. A) Absolute (columns) and
cumulative (line) number of articles published up to December 2023. B) Global distribution and
number of papers published by corresponding author's country.

Significant growth in research interest in paraben toxicity was observed from
2016 onwards (Fig. 3A). This trend may be attributed to increased awareness of
emerging pollutants and their potential impact on ecosystems. Notably, the COVID-19
pandemic accelerated this interest, as the heightened use and disposal of personal
care and hygiene products may contain parabens in their composition. For instance,

52
Fig. 3A illustrates a marked rise in publications on the ecotoxicity of parabens to
aquatic organisms in 2022 and 2023, coinciding with the scientific community's
heightened focus on the environmental consequences of pandemic-related chemical
use (Qualhato et al., 2023). This suggests that the pandemic may have acted as a
catalyst for more urgent research into the environmental risks of parabens.
Interestingly, between 2022 and 2023, a series of studies in China
investigated the ecotoxicity of MeP in zebrafish. These studies showed that MeP
exposure affects gene expression in various organs, induces oxidative stress, disrupts
embryonic development, and impairs endocrine functions (Hu et al., 2022a, 2022b;
2023a,b; Liang et al., 2022, 2023a; 2023b). Currently, most research has focused on
areas such as water and air pollution, conservation, and the sustainable use of
environmental resources, all of which have shaped investigations into the toxic
potential of parabens to human and environmental health (Azeredo et al., 2023; Mejías
et al., 2023; Vale et al., 2022).
Fig. 3B depicts the geographical distribution of studies investigating the
ecotoxicity of parabens in aquatic environments across 19 countries. China had the
highest number of authors who have researched this subject (n = 26; 34.21% of the
total), followed by the USA (n = 7; 9.21%), Italy (n = 6; 7.89%), India, and Japan (n = 6;
7.89% each) (Fig. 3B). Therefore, it is essential to note that further research is needed
in regions such as Africa, Oceania, Canada, Russia, South America, and others to
ensure a broader understanding of the concentrations of parabens present in
environmental samples.
Ecotoxicological studies on parabens and aquatic organisms were
predominantly conducted under laboratory conditions (n = 57; 75%) compared to field
studies (n = 19; 25%) (Fig. 4A). Laboratory studies have primarily focused on
assessing multiple biomarker responses and elucidating parabens' MoA and effects in
different levels of organization (i.e., organs, tissues, and cells) of various aquatic
species. In contrast, field studies have primarily accessed paraben bioaccumulation or
metabolites in different tissues or the entire organism. Field studies were conducted at
128 sampling sites across five countries: China (50%), USA (15.79%), Spain (10.59%),
Italy, Colombia, and Vietnam (5.26% each). These data suggest the need for more

53
field studies in countries or regions with high biodiversity, such as Brazil, Colombia,
Mexico, and Vietnam, to evaluate the bioaccumulation in various taxonomic groups,
mainly in the global south countries.

54

Figure 4. Details of ecotoxicological studies focused on parabens in aquatic organisms. A)
Percentage of field and laboratory studies. B) Number of articles published by groups of
organisms in field studies. C) Frequency of the different biological samples from aquatic
organisms (invertebrates, vertebrates, and plants) whose paraben s bioaccumulation was
evaluated in field studies. The 21.9% include the soft tissue of invertebrate species and also
the whole body of some fish species.

1.4 Field studies on the bioaccumulation of parabens in aquatic organisms
Field studies are highly beneficial in identifying potential sources of
contamination that may not be possible in laboratory studies (Graney et al., 2020). This
approach offers the advantage of understanding the effects of realistic paraben
exposure in aquatic ecosystems and organisms (Zhang et al., 2021; Haman et al.,
2015). By studying the responses of organisms exposed to varying substance
concentrations, researchers can assess their potential ecotoxicity and understand the
long-term consequences for individual species and ecosystem health (Bom & Sá,
2021). Also, field studies using aquatic animal models allow researchers to analyze the
interactions between parabens and other environmental conditions and stressors
(Graney et al., 2020). This way, water bodies might be exposed to multiple
contaminants and stressors, which can have greater than additive or antagonistic
effects on the toxicity of parabens (Miranowicz-Dzierżawska et al., 2023). Researchers
can provide a more comprehensive understanding of the environmental impact by
studying paraben exposure with factors such as temperature, pH, and co-occurring
pollutants in this context.
Field studies were conducted with a wide range of aquatic species, mainly
fish (72.97%), followed by mollusks (14.23%), crustaceans (4.39%), aquatic
angiosperms (4.39%), algae (0.34%), porifera (0.34%), tunicates (0.34%), and
zooplanktons (0.3%). Among these groups, the number of articles published with fish is
almost three times higher than the number of studies conducted with mollusks, as seen
in Fig. 4B. Field studies have assessed 92 freshwater (50.27%) and 91 marine species
(49.73%), spanning both vertebrates and invertebrates. These field studies reveal the
complex interactions between parabens and aquatic ecosystems, uncovering the
potential risks of their widespread presence. Therefore, aiming to give more detailed

55
information about paraben bioaccumulation data from field studies of different taxa of
aquatic organisms, the following sections present the most relevant works in the
literature covering this topic.
1.4.1 Green algae and aquatic angiosperms
Up to now, one field study has analyzed the bioaccumulation of parabens in
three species of green algae and three species of aquatic angiosperms. Xue et al.
(2017) investigated the presence of MeP, EtP, PrP, BuP, and BzP in the green alga
Caulerpa prolifera, in marine angiosperms Syringodium filiforme, Halodue wrightii, and
Ruppia martima, and the mangrove angiosperms Avicennia germinans, Rhizophora
mangle, and Laguncularia racemosa, all collected from coastal waters in Florida
(USA). MeP was detected in all mangrove samples at concentrations ranging from 8 to
33.3 ng g-1, and also in all seagrass samples in concentrations ranging from 4.0 to
42.2 ng g-1 (wet weight – ww), being both higher than those found in the green alga
(9.6–16.5 ng g-1 ww). Other parabens, such as EtP, BuP and BzP were found in some
angiosperm samples (4.5%–13.6%), at concentrations up to 7.78 ng g-1 ww for BzP in
a mangrove samples (Xue et al., 2017).
Similarly, a study conducted in the Dongjiang River Basin, China,
investigated the occurrence and bioaccumulation of nine types of parabens and two
metabolites in samples of plants and algae (Lin et al., 2024). The study found
considerable concentrations of these chemicals in all samples, in which MeP (240 ng
g-1 dry wt), i-BuP (18.7 ng g-1 dry wt), and BzP (5.18 ng g-1 dry wt) were detected in
all plant samples, while MeP (48.4 ng g-1 dry wt), EtP (10.8 ng g-1 dry wt), and PrP
(4.03 ng g-1 dry wt) were most abundant in algae. The author also estimated the
bioaccumulation factor (BAF) in plankton samples (including zooplankton), being the
higher log BAF seen for EtP (5.61), followed by PrP (5.54) and MeP (4.74) (Lin et al.,
2024). However, they did not calculate the log BAF for plants. Although these log BAF
values denote a high potential for bioaccumulation in plankton, the authors suggested
that these may have been overestimated, since they used normalized dry weight (dw)
concentrations in such calculation. All these findings provide baseline information
about the occurrence and fate of parabens in aquatic plant biodiversity.

56
Interestingly, parabens are naturally produced by certain terrestrial plants,
where they act as natural preservatives to protect against microbial growth and other
environmental stressors (Ali et al., 1998; Li et al., 2003; Nowak et al., 2018). However,
the levels of parabens found in plants are lower than in PCPs and cosmetics, which
use synthetic parabens as preservatives.
1.4.2 Aquatic invertebrates
Reviewed data showed that ten field studies analyzed the bioaccumulation
of parabens in invertebrate species, mainly mollusks (n = 6) and crustaceans (n = 4).
Furthermore, tunicate and poriferan species were analyzed in only one study. For
these species, parabens were quantified in the whole organism, as presented in Fig.
4C. Aquatic invertebrates and benthic biodiversity research can provide valuable
information about the potential exposure to parabens in aquatic environments (Zhu et
al., 2024). Also, it is worth mentioning that bivalves, such as mussels and oysters, are
commonly used in ecotoxicity testing as they filter organisms and can accumulate
contaminants from their surrounding environment, making them valuable indicators of
environmental contamination. On the other hand, bivalves can also close their shells
and potentially reduce exposure to aquatic contaminants (Bom & Sá, 2021).
In a study conducted in the Chinese Bohai Sea between 2006 and 2015 by
Liao & Kannan (2018), the total concentrations of parabens detected in mollusk tissue
samples from Mactra veneriformis, Mytilus edulis, and Cyclina sinensis ranged from
2.66 to 299 ng g−1 ww, with MeP being the most commonly detected parent paraben. In
addition, a gradual increase in the concentration of parabens in the samples was
observed over the years (Liao & Kannan, 2018). These authors point out that the
increasing paraben concentrations in mollusk samples were related to the growing
production and use of parabens-containing products in China, culminating in the
inevitable increase in the disposal of parabens and their metabolites in aquatic
environments. Similarly, MeP was also the main paraben found in invertebrate samples
from Florida coastal waters, with a detection frequency of 82% (n = 186). The
concentration of MeP varied from less than 2.01 ng g-1 ww for shrimp Farfantepenaeus
aztecus to 337 ng g-1 ww for the bivalve Chione cancellata (Xue et al., 2017).

57
In addition, Chiesa et al. (2018) evaluated two bivalves: a Mediterranean
mussel (Mytilus galloprovincialis) and a clam species (Chamelea gallina) collected
from fish markets in Milan, obtained from farming sites in the Mediterranean Sea. The
authors identified MeP as the main paraben detected in the samples, with
concentrations up to 32 ng g−1 (Chiesa et al., 2018). Likewise, Zhu et al. (2024)
evaluated parabens' bioaccumulation in four crustacean species (n = 24) and two
cephalopod species (n = 11), from the Beirpweihbu Gulf, South China Sea. The
primary contaminants identified in marine organisms were MeP and 4-HB, with
concentrations ranging from 0.18 to 5.03 ng g-1 ww for MeP in crustaceans, as well as
from 1.45 to 2.67 ng g-1 ww for cephalopods. The log BAF of MeP in cephalopods
(2.80) was similar to those found in fish (2.85), but significantly higher than that in
crustaceans (2.37), denoting that fish and cephalopods have a greater capacity to
enrich MeP from seawater (Zhu et al., 2024). Since most species evaluated were
benthic organisms, biota-sediment accumulation factors (log BSAF) were estimated for
MeP (5.51), revealing notable accumulation from sediment in the investigated
environment.
Considering freshwater ecosystems, Lin et al. (2024) evaluated paraben
concentrations in zooplankton samples from Dongjiang River wetland ecosystem in
southern China. MeP was detected in all plankton samples, followed by EtP (95.0%)
and PrP (95.0%). However, i-PrP, BuP, i-BuP, BzP, pentylparaben (PeP), and HeP
were less frequently detected (<40.0%). MeP (7.37–139 ng g-1 dry wt) was the most
abundant paraben detected, followed by PrP (<LOQ – 113 ng g-1 dry wt), and EtP
(<LOQ – 598 ng g-1 dry wt). Similarly, MeP, followed by EtP and PrP, were the most
abundant parabens detected in eight shellfish families (Buccinidae, Veneridae,
Haliotidae, Mytilidae, Pectinidae, Ostreidae, Corbiculidae, and Pteriidae) from
Shenzhen coastal waters, which represents more than 95% of the total the parabens
detected (Lu et al., 2019). All these studies performed in distinct aquatic environments
and countries denote the ubiquitous contamination by parabens, as needed for further
field studies in elucidating the real-world exposure of aquatic organisms to parabens,
which becomes important in assessing human risks in areas with high seafood
consumption.

58

1.4.3 Fishes
A total of 17 field studies analyzed the bioaccumulation of parabens in 208
fish species. Among them, 12 studies (63.15%) were performed with freshwater
species, while 7 studies 36.84%) used marine ones. The tissues or organs most
evaluated in these field studies (Fig. 4C) were the muscles (28.57%), liver (14.29%),
brain and kidney (7.14% each), followed by gills, fat tissue, plasma, and bile (3.57%
each). Among such studies, the parabens were detected only in 16 of them, since
Renz et al. (2013) did not detect paraben concentrations in the brain samples of three
fish

species

(Alosa

pseudoharenga,

Micropterus

dolomieu,

and

Dorosoma

cepedianum) from the Greater Pittsburgh area, USA.
The first study devoted to the bioaccumulation of parabens in aquatic
organisms was carried out by Ramaswamy et al. (2011), in which the determination of
four parent parabens was investigated in muscle samples of 20 fish species (11
demersal and nine pelagic fish) from Manila Bay (Philippines). MeP, PrP, and BuP
were detected in more than 90% of the fish samples analyzed, while EtP was detected
in around 70% of the samples. MeP was detected in higher concentrations, ranging
from <0.05 to 3600 ng g-1, whereas EtP <0.011–840 ng g-1, PrP <0.024–1100 ng g-1,
and BuP <0.003–70 ng g-1 pointing out the contamination of Manila Bay by parabens.
Interestingly, they also observed a positive correlation between parabens concentration
and fish length (Ramaswamy et al., 2011). Xue & Kannan (2016), when studying inland
lakes and rivers of New York State, detected concentrations of MeP ranging from
<2.01 to 690 ng g-1 ww in muscle and liver of the fish species Micropterus dolomieu
and M. salmoides. In addition, MeP was detected at concentrations ranging from <2.01
ng g-1 in the muscle of Mugil cephalus to 735 ng g-1 ww in the brain of Rhizoprionodon
terraenovae from Florida coastal waters (Xue & Kannan, 2016).
In another study conducted with 13 species of fish from four river basins in
Spain, MeP and BzP were found in specimens from all basins, with concentrations
ranging from 3.41 to 84.69 ng g−1 dw for MeP, from 0.63 to 3.46 ng g-1 dw for PrP, and
from 0.35 to 0.54 ng g-1 dw for BzP, with a log BAF of 1447 for MeP and 166 for PrP.
The study also reported that the predominant group of EDCs in the samples consisted

59
mainly of parabens (Pico et al., 2019). Consistent with these results, Wang et al.
(2019) detected parabens in all fish samples from Taihu Lake, China, during an
investigation conducted from 2009 to 2017. MeP, EtP, and PrP were found in all
samples

from

the five fish species (Culter alburnus, Carassius carassius,

Hypophtalmichtys molitrix, H. nolitrix, and Protosalanx hyalocranius). At the same time,
BuP was detected in 47.70% of the samples, and BzP in 79.60%. The concentrations
of MeP ranged from 88.1 to 1200 pg g-1 ww, while EtP concentrations ranged from 33.6
to 450 pg g-1 ww, and for PrP, from 55.3 to 543 pg g-1 ww. The total sum of parabens
ranged from 261 to 1710 pg g-1 ww (Wang et al., 2019). In the Chinese Pearl River,
nine species of fish were studied for the determination of phenolic EDCs and, following
the abovementioned study, MeP was the main paraben found in fish samples (from
2.01 to 201 ng g-1 ww), and the second most frequently detected EDC, with a median
concentration of 40 ng g-1 lipid weight (lw) (Peng et al., 2018). The same study found
that the total concentration of parabens was higher in the liver than in abdominal fat
and dorsal muscles, as well as increasing paraben bioaccumulation in parallel to
increasing fish weights. The highest men log BAF was found for MeP (2.0 in dorsal
muscle, 2.4 in belly fat, and 3.8 in liver), whilst the lowest was for Bup (1.3 in dorsal
muscle, 1.9 in belly fat, and 2.1 in liver) (Peng et al., 2018).
Xue et al. (2017) analyzed 35 fish species from coastal Florida and showed
that the MeP concentrations ranged from 2.01 ng g-1 (Lagodon rhomboides, whole
homogenized fish) to 610 ng g-1 ww (Lutjanus campechanus, liver). The mean
concentration of MeP in the kidney (181 ± 138 ng g-1 ww) was higher than in the liver
(53 ± 61.4 ng g-1 ww) and the whole fish (30.3 ± 19.9 ng g-1 ww), indicating differential
tissue/organ bioaccumulation. The study also detected EtP concentrations in the
kidney ranging from <2.01 in Sphyrna mokarran to 16.8 ng g-1 ww in M. cephalus. As
for the other parabens (i.e., PrP, BuP, BzP, and HeP), concentrations were generally
very low: the highest concentrations were 7.29 ng g-1 ww of PrP in Scomberomorus
maculatus and 7.32 ng g-1 ww of BuP in S. mokarran. Chiesa et al. (2018) evaluated
specimens from the fish market (from farmed or wild by multiple sites) and showed that
the main paraben found was also MeP, being detected in 41% of samples (n = 54),
followed by EtP (3%). The other parabens were not detected in any samples except for

60
EtP, which was found in 3% of samples (Chiesa et al., 2018). These authors also drew
attention to the high variability in the accumulations found, depending on the species,
location of origin, and capture site.
MeP also was the most frequent parent paraben found in muscle samples
from fish collected in Dongjiang River, China, with concentrations ranging from
<LOQ–194 ng g-1 dw); followed by EtP (88.0%) and PrP (81.0%) (Lin et al., 2024).
Other parabens (i-PrP, BuP, i-BuP, BzP, PeP, and HeP were detected with <20%).
These authors also found a higher log BAF for MeP (3.77), followed by PrP (3.77), and
EtP (3.43). Also, Zhu et al. (2024) detected MeP in all fish species from Beibu Gulf,
South China Sea, with concentrations ranging from 0.66 to 13.77 ng g-1 ww, followed
by PrP (71.70%), EtP (49.06%), BuP (25.47%), and BzP (7.55%). These authors also
estimated the log BAF ranging from 2.36 to 3.37 for MeP, and from 1.42 to 2.34 for
PrP, pointing to insignificant bioaccumulation. Furthermore, they found a Trophic
Magnification Factor (TMF) of 2.88 and 1.17 for MeP and PrP, respectively, unveiling
MeP biomagnification throughout the food web (Zhu et al., 2024).
Regarding the accumulation of polar parabens (i.e., short-chain parabens as
MeP and EtP), among the 19 field studies analyzed in this review, we found no
plausible explanation for the differences in the detection of MeP and EtP in different
organs. In these, paraben accumulation was assessed mostly in a single organ/tissue
or the entire animal (Cacua-Ortiz et al., 2020; Jakimska et al., 2013; Liao & Kanan,
2018; Lin et al., 2024; Pico et al., 2019; Renz et al., 2013; Wang et al., 2019; Xue et
al., 2017; Yao et al., 2018a; Yao et al., 2019; Zhu et al., 2024). Only in three studies
(Peng et al., 2018; Xue & Kannan, 2016; Xue et al., 2017) more than one organ was
evaluated, not allowing us to draw up a profile for paraben bioaccumulation in different
organs in aquatic organisms. Although MeP and EtP have low bioaccumulation
potential (Lu et al., 2019), given the low lipophilicity for MeP (Log octanol-water
partition coefficient – log KOW = 1.96) and moderate lipophilicity for EtP (log KOW =
2.47), together with PrP (log KOW = 3.04) are the most commonly detected parabens
in aquatic organisms (Haman et al., 2015; Lu et al., 2019; Tran-Lam et al., 2023; Xue
et al., 2017; Yao et al., 2018b). This situation can be explained by the extensive use of
products containing these parabens by the world population (Bledzka et al., 2014), and

61
also due to MeP's higher solubility in water (Ramaswamy et al., 2011), since the
solubility of parabens decreases according to the length of the ester chain (Haman et
al., 2015).
1.4.4 Detection of paraben metabolites
Only eight (42.10%) field studies aimed at detecting parabens in aquatic
organisms have evaluated their metabolites, even though their concentrations,
especially of 4-HB, were usually much higher than those of the parent parabens. Xue
et al. (2017) detected 4-HB in 99% of the biotic samples collected for the study, with
concentrations up to 68,100 ng g-1 ww for the marine bivalve C. cancellata. Also, other
metabolites detected in the same study were methyl-protocatechuate (OH-MeP),
ethyl-protocatechuate (OH-EtP), 3,4-dihydroxybenzoic acid (3,4-DHB), and 4-HB, in all
marine plant samples, at concentrations up to 13,500 ng g-1 ww for mangrove samples.
Interestingly, the same study found higher concentrations of 4-HB in fish kidneys (4110
± 5354 ng g-1 ww) in comparison with whole fish (1620 ± 1640 ng g-1) and liver (738 ±
983 ng g-1) (Xue et al., 2017). Consistent with these data, 4-HB was the predominant
metabolite found in mollusk (gastropods and bivalves) samples in the Bohai Sea,
China, between 2006 and 2015, with concentrations up to 161,000 ng g-1 dw in marine
bivalve Mytilus edulis (Liao & Kannan, 2018). The metabolite 3,4-DHB was also
frequently found (91.9%) at concentrations up to 4960 ng g-1 dw. In addition, this study
detected a temporal increase in the concentrations of MeP and 4-HB in the mollusk
samples collected between 2006 and 2012, suggesting a common source of parabens
in these mollusks (Liao & Kannan, 2018).
Lin et al. (2024) detected 4-HB in all plankton samples (1180–24800 ng g-1
dw in algae, and 336–10300 ng g-1 dw in zooplankton), in all plant samples
(1310–11800 ng g-1 dw), and fish muscle samples (7.34–814 ng g-1 dw) from Dongjiang
River Basin, China. On the other hand, 3,4-DHB was more abundant in plants
(<LOQ–27200 ng g-1 dw) in comparison with plankton (1.21–185 ng g-1 dw in algae and
<LOQ–75.5) and fish muscle (<LOQ–14.1 ng g-1 dw). Also, 4-HB showed the highest
log BAF (5.79) in plankton samples than in fish muscle (4.44) (Lin et al., 2024). In
another study conducted in fish and bivalve samples from a fish market in Milan, Italy,

62
4-HB was detected in 75% of the samples, with concentrations up to 35.660 ng g-1
Chiesa et al. (2018). The highest amounts found in bivalve samples were attributed to
the fact that they are benthic filter-feeding organisms and capable of bioaccumulating
certain contaminants Chiesa et al. (2018).
Recently, Zhu et al. (2024) provided important data on the fate and transfer
of 4-HB in a subtropical marine ecosystem. They investigated the parent parabens and
4-HB in a marine food web from the Beibu Gulf, South China Sea, including 4
crustacean species, 2 cephalopod species, and 19 fish species. 4-HB was one of the
predominant contaminants in the marine organisms analyzed, with concentrations
ranging from 13.48 to 222.24 ng g-1 ww, but had a low log BAF in fishes (2.73–3.24 ng
g-1 ww), in crustaceans (2.86–3.45), and cephalopods (2.76–2.79 ng g-1 ww). Such a
study also indicated that 4-HP has an estimated high BSAF, but has a low TMF, which
suggests trophic dilution throughout a marine food web (Zhu et al., 2024). All these
data highlight the importance of evaluating the bioaccumulation of parent parabens
and their metabolites in aquatic organisms using species from different trophic levels
and sampling periods.
1.5 Laboratory studies
Up to now, 57 laboratory studies have investigated the ecotoxicity of
parabens in aquatic organisms, using 78 freshwater species (64.46 %) and 43 marine
species (35.54 %). Experimental studies are considered essential techniques,
especially for ecotoxicology researchers, to investigate the effects of certain
substances on living organisms. These studies are carefully designed and controlled
environments in which researchers can manipulate one or more variables to determine
their effects on the subject under investigation (Escher et al., 2018). Experimental
approaches also allow researchers to establish cause-and-effect relationships between
variables. This control allows researchers to better understand the specific effects of
the factor under study, and to make more accurate predictions about how it will affect
the individual (Li & Xia, 2019).
In the context of parabens and their potential ecotoxicity to aquatic
organisms, experimental studies can provide valuable information on the specific

63
mechanisms by which parabens affect these organisms, the concentration and or
dose-response relationship between paraben exposure and ecotoxicity, and the
potential long-term effects of chronic exposure. In this sense, the works applying
different model organisms can assess their paraben-toxicity level depending on the
method and strategy employed. They can also develop a comprehensive view of the
relationship between paraben contamination on the environment and organisms, which
foment further studies in paraben toxicity in humans.
Shore et al. (2022) evaluated MeP toxicity in an echinoderm species
(Strongylocentrotus purpura) and showed adverse effects on general animal
development, such as mortality and reduced body growth. This is an interesting finding
since these marine organisms can indicate environmental quality and potential
contamination. In addition, other studies applying in vitro assay and the marine fish
species (Oryzias latipes) reported severe impacts on endocrine and reproductive
systems due to exposure to MeP and PrP paraben derivatives (Kawashima et al.,
2022), indicating that experimental and controlled assay methods are crucial to verify
the changes related to endocrine signaling.
Aiming to organize and provide detailed information about the several
bioassays that have investigated the toxicity of parabens within aquatic species, a
summary is presented in Table 2. It describes their experimental details, biomarkers
evaluated and the main findings, from studies performed with aquatic microorganisms,
invertebrates, and vertebrate species.
1.5.1 Model species
Laboratory studies about the ecotoxicity of parabens were conducted mainly
with fish (n = 40; 70.17%), followed by crustaceans and or microcrustaceans (n = 9;
15.78%), algae (n = 4; 7%), platyhelminth, algae and bacteria (n = 4; 7% each),
bivalves (n = 2; 3.5%), reptile (turtles) (n = 2; 3.5%), echinoderms (n = 1; 2%), and
amphibians (n = 1; 2%), as seen in Fig. 5A. Fish are the most biodiverse group within
vertebrates (Volff, 2005), and have been so far the most studied group among them to
investigate the ecotoxicity of parabens (Fig. 5A). Furthermore, there is a
disproportionate number of studies using fish models in comparison with other groups

64
of vertebrates, invertebrates, and microorganisms (Fig. 5A). Six fish species were used
in all the papers screened (Fig. 5B), which might be skewing the current evidence
towards the sensitivity of such group to parabens, even though it may vary between
species (González-Doncel et al., 2014). Among the fish model species used (Fig. 5B),
the zebrafish was by far the most commonly utilized (n = 26; 65%) in such studies,
followed by the medaka (O. latipes, n = 5; 12.5%), and the rainbow trout (O. mykiss, n
= 3; 7.5%). Similarly, other reviews have also shown that zebrafish is the most
commonly used fish species as a model system to assess the ecotoxicity of traditional
and emerging pollutants (Bambino and Chu, 2017; Canedo et al., 2021; Meyers, 2018;
Porto et al., 2023).

65

Figure 5. An overview of laboratory studies related to the ecotoxicity of parabens in aquatic
organisms. A) Number of published articles per taxon. B) Number of articles published per fish
species adopted in each study. C) Number of articles published by fish life cycle stage
assessed in each study. D) Frequency of organs and tissues investigated in laboratory studies.

66
E) Number of articles by type of paraben tested in each study. F) Biomarkers and other
parameters used in laboratory studies to investigate the ecotoxicological effects of parabens.

Considering the experimental studies carried out with fish, the most
commonly used life stages cycle were the early developmental stages (embryo-larval
stage) (Fig. 5C). During this stage, fish are more sensitive to toxicants, allowing for the
observation of morphological abnormalities and facilitating the development of
large-scale studies (Porto et al., 2023). Furthermore, laboratory analyses have
examined various organs and tissues from adult fish species to elucidate the effects
and MoA of parabens (Fig. 5D). Among the samples evaluated, the liver (9.43%), brain
(5.66%), and intestine (5.66%) were the most extensively investigated organs, followed
by gills, testis, and blood (3.77% each) (Fig. 5D). Bioaccumulation was also
investigated in a few laboratory studies, as detailed in section 5.3.
In addition to fish, other vertebrates such as turtles and amphibians have
recently been used to evaluate paraben ecotoxicity. As seen for the Chinese
striped-neck turtle Mauremys sinensis (n = 2; 3.5%) (Ding et al., 2023) and for the
African clawed frog Xenopus laevis (n = 1; 2%) (Medkova et al., 2023). Chelonians are
excellent bioindicator species and models for assessing the impacts of anthropogenic
pollution, due to their sedentary lifestyle, long life cycle, multiple feeding habits, and
capacity for bioaccumulation and biomagnification of contaminants (Adams et al.,
2016; Ding et al., 2023). Similarly, amphibians, given their dual aquatic-terrestrial life
stages, sensitivity to both water and soil contaminants, and vulnerability to
environmental stressors (Do Amaral et al., 2019; He et al., 2014).
Microorganisms and invertebrates were also utilized to assess the
ecotoxicological impact of parabens. Specifically, the green algae Pseudokirchneriella
subcapitata (n = 2; 3.8%), the bacterium V. fischeri (n = 2; 3.8%), and the freshwater
microcrustacean D. magna (n = 6; 11%). Other aquatic arthropods, such as Artemia
franciscana, C. dubia, and Tigriopus japonicus, appear in various studies. Beyond
crustaceans,

invertebrate

model

organisms

used

included

the sea urchins

Strongylocentrotus purpuratus and Paracentrotus lividus, the planarian Dugesia
japonica, and the oyster Crassostrea gigas (n = 1; 1.8% each).

67
In this context, a few studies (n = 8; 15%) used microorganisms to assess
paraben

ecotoxicology.

approximately

Similarly,

studies

involving

invertebrates

represented

20% of all the research analyzed in this review. Including

microorganisms in future studies is particularly interesting due to their rapid
reproduction rates and simple cellular structures, which allow for high-throughput
screening and detailed studies. Additionally, microorganisms often serve as primary
producers or decomposers in aquatic ecosystems, making them crucial indicators of
ecological

health

(Sivapragasam

et

al.,

2020).

Notably,

the

sensitivity

of

microorganisms and invertebrates to parabens may surpass that of some vertebrates
under similar conditions (Yamamoto et al., 2011). Therefore, it is pertinent to consider
including these species in future ecotoxicological assessments of parabens, as they
can provide early warning signs of environmental contamination and contribute to a
more comprehensive understanding of the ecological impacts of these chemicals.
1.5.2 Types of parabens tested
Most published studies only investigated a single paraben (n = 37; 86.04 %)
(Fig. 5E). Still, some studies have tested more than one paraben individually (n = 18;
41.86 %), while few studies (n = 2; 4.65 %) have used mixtures of parabens as the
focus of research (Fig. 5E). The paraben type most commonly employed in
experimental studies was MeP (57.74% of 71 studies), followed by PrP (42.25%) (Fig.
5E). Also, BuP was the third most employed paraben in the revised experimental
studies (Fig. 5E), exhibiting higher ecotoxicity than parabens of shorter alkyl chains
(Merola et al., 2020b). MeP and PrP are the most frequently investigated and detected
parabens in surface water and sediments, as well as in wastewater, effluent
wastewater treatment plants, and sewage sludge (Haman et al., 2015), due to the
continuous consumption of products containing parabens and their final destination in
aquatic environments. Given this, researchers have focused their work mainly on these
two parabens.
Regarding their metabolites, Pedersen et al. (2000) failed to find any toxicity
related to 4-HB in O. mykiss exposed via intraperitoneal injection. Since this was the
earliest paraben-related toxicity study made with an aquatic organism, it may explain

68
why only a single study employed paraben metabolites after that, even though there
are some other than 4-HB that might need to be investigated. Some of those were
detected at high concentrations in fish samples, such as OH-MeP, OH-EtP, and 4-DHB
(Xue et al., 2017). Additionally, studies conducted over the last decade have detected
4-HB levels ranging from 4.58 to 2380 ng L-1 in surface waters from China (Li et al.,
2015; Ma et al., 2018; Zhao et al., 2019), and from 1380 to 31,400 ng L-1 in wastewater
from India (Karthikraj et al., 2017). These findings underscore the need for further
research on the ecotoxicity of paraben metabolites across different groups of
organisms. Thus, understanding these contaminants' ecological and health impacts is
essential, given their widespread presence in aquatic environments.
1.5.3 Bioaccumulation in model species
Among the laboratory studies analyzed in this review, only four evaluated
the paraben bioaccumulation in aquatic organisms. In O. mykiss exposed to water
containing 225 μg L−1 for 12 days, PrP accumulation was investigated and detected in
the liver (6700 μg kg-1) and muscle (870 μg kg-1) (Bjerregaard et al., 2003), denoting
that that bioaccumulation in the liver was almost eight times greater than in the
muscles. However, in this same study, when specimens were orally exposed to
7.2–1830 mg kg-1 PrP by food for 10 days, the PrP bioaccumulation was detected only
in the liver (37 mg g-1 ww) from specimens exposed to the highest dose. In addition,
although muscle samples were also collected after the oral exposure experiment, in
such a study the authors could not detect PrP in this tissue (Bjerregaard et al., 2003),
confirming that the bioaccumulation of parabens in fish also depends on the route of
exposure.
Alslev et al. (2005) also conducted a study with O. mykiss, in which juvenile
specimens were orally exposed to 4–74 mg kg-1 of BuP by food for 10 days. BuP
accumulation was not detected in the muscle tissue sampled, whilst on the liver range
of 0.6–5 μg g-1 ww, presenting a discrete dose-dependent increase in PrP
concentrations in this organ (Alslev et al., 2005). In a second experimental condition,
BuP concentrations of 9 and 183 μg g-1 were detected in the plasma samples of fish
exposed to water containing the nominal concentrations of 50 and 250 μg L-1,

69
respectively, over 12 days (Alslev et al., 2005). Although liver and muscle samples
from O. mykiss were also collected in the second experiment by water exposure, the
authors could not assess BuP concentrations in them due to technical problems. In
both studies conducted with O. mykiss and paraben exposure by food, PrP and BuP
accumulation represented less than 1% of administered doses, suggesting that when
administered orally to fish, these two parabens are rapidly metabolized or poorly
absorbed. In mammals, parabens are rapidly absorbed by the gastrointestinal tract and
metabolized by esterases, producing p-hydroxybenzoic and other metabolites. This
metabolization may occur in the skin, subcutaneous adipose tissue, and in the
digestive tract (Bledzka et al., 2014; Soni et al., 2005). However, a study with five fish
species detected esterase activity in the intestine, liver, and bile. Nevertheless, O.
mykiss presented the lowest activity among the species investigated (Li & Fan, 1997).
In the study conducted by Barse et al. (2010), adult males of Cyprinus
carpio were exposed to water containing 0.84, 1.68 and 4.20 mg/L of MeP - which is
equivalent to 1/143, 1/71st and 1/29 of LC50 of MeP calculated for this species,
respectively, for 28 days. MeP bioaccumulation was investigated and detected in gill
(∼0.48 and ∼0.84 mg kg-1), brain (∼0.9 and ∼2.3 mg kg-1), liver (∼1.0 and ∼1.7 mg kg-1),
muscle (∼0,48 and ∼0,84 mg kg-1) and testes (∼0.7 and ∼1.4 mg kg-1), presenting a
concentration-dependent bioaccumulation in the organs of fish exposed to the two
lowest concentrations (0.84 and 1.68 mg L-1), respectively. Such data corroborated the
findings in O. mykiss on the bioaccumulation of parabens in fish liver and muscle after
oral exposure, as well as being the first published study to indicate that parabens can
bioaccumulate in multiple organs of aquatic organisms and, consequently, cause some
disturbance in their activity.
In addition to the three studies mentioned, only another study evaluated the
bioaccumulation of parabens in a different group of aquatic organisms. In it, adult
females of the turtle species M. sinens were exposed to different concentrations of
BuP (5, 50, and 500 μg L-1) for 20 weeks, and the bioaccumulation of BuP only in the
intestine

was

assessed

(Ding

et

al.,

2023).

A

concentration-dependent

bioaccumulation was observed for this organ (∼4.1, ∼60, and ∼1000 ng g-1 of BuP),
respectively (Ding et al., 2023). In such study, turtles were exposed to BuP through

70
water ingestion and, although the authors stated that BuP bioaccumulation in the
intestine was low, this was sufficient to induce dysbiosis as well as changes in the
intestinal physiology and structure of M. sinens. Among the articles analyzed in this
review, we did not find any laboratory studies that evaluated the bioaccumulation of
paraben metabolites, reinforcing that the bioaccumulation of parent parabens and their
metabolites should be investigated in further studies.
1.5.4 Ecotoxicological effects on aquatic organisms
Unlike articles based on field studies, in which only the presence and
amounts of parabens in biota samples were investigated, in the laboratory articles
reviewed here, several biological effects caused by paraben contamination were found,
as shown in Fig. 5. These effects were evaluated using several biomarkers or
parameters presented in Fig. 5F and also detailed in Table 2, which were chosen
according to the type of aquatic organism used, the stage of the life cycle evaluated,
and the objectives of each investigation.
In ecotoxicology, biomarkers are carefully chosen based on the expected
responses resulting from exposure, thereby providing valuable data for conducting
biomarker-based ecological monitoring projects (Porto et al., 2023). Specifically,
studies utilizing aquatic animals as models have identified a diverse range of
biomarkers that indicate exposure to parabens or their effects (Fig. 5F and Table 2),
which encompass molecular to physiological endpoints (Pedersen et al., 2000;
Thakkar et al., 2022; Ates et al., 2018; Torres et al., 2016). As demonstrated in Fig. 5C,
endocrine disruption, oxidative stress, developmental toxicity, and reproductive
impairment are the major effects promoted by parabens in aquatic organisms in the
articles analyzed, which will be described in more detail below.
1.5.4.1 Effects on development and growth
The earliest study conducted on fish during the embryonic phase was
performed by Dobbins et al. (2009) using the freshwater fish Pimephales promelas to
explore the growth rate and seven paraben compounds (MeP, EtP, PrP, isoPrP, Bup,
iso-BuP, and BzP). Using 24 h post-hatch P. promelas and a static exposure for 48 h,

71
the LC50 values found ranged from 3.0 to 160 mg L-1, which correspond to the LC50
values calculated for MeP and BzP, respectively. Also, exposure to parabens
containing longer alkyl chains was associated with reduced growth in fish (Dobbins et
al., 2009). Similarly, the developmental toxicity induced by parabens was reported in
the Japanese medaka (O. latipes) (González-Doncel et al., 2014), and zebrafish
(Bereketoglu & Pradhan, 2019; Torres et al., 2016). Furthermore, O. latipes embryos
and larvae exposed to PrP (40, 400, 1000, and 4000 mg L-1) showed decreased
survival rate, developmental delay, increased morphological changes, and histological
abnormalities (González-Doncel et al., 2014). Besides, zebrafish embryos and larvae
exposed to MeP (100–1000 µM) for 96 h showed several malformations, and a decline
in heart rate (cardiotoxicity) and hatching rate (Dambal et al., 2017). Bereketoglu &
Pradhan (2019), using a semi-static exposure, also observed that MeP (≥100 µM) and
PrP (≥10 µM) exposure promotes concentration-dependent toxicity in zebrafish
embryos, in parallel to the induction of morphological abnormalities. They also found
that PrP is more toxic to zebrafish embryos than MeP. In addition to these endpoints,
the study conducted by Merola et al. (2020a) demonstrated the occurrence of blood
stasis, reduction in blood circulation, and also reduced heartbeat in zebrafish larvae
exposed to MeP (1–80 mg L-1) during 96 h, drawing attention to the potential
ecotoxicological impact of parabens during the fish early developmental stages.
Similar to fish, Xenopus laevis embryos exposed to MeP, PrP, and BtP
showed 100% mortality for the higher concentrations tested (5000 and 100,000 μg L-1),
hatching delay, developmental abnormalities, and gene expression downregulation
(Medkova et al., 2023). To our knowledge, there are no studies addressing the effects
of parabens on the development of other aquatic vertebrates, such as turtles, different
classes of amphibians, as well as in neotropical fish species.
Paraben exposure, including MeP, EtP, and PrP, has been associated with
developmental alterations in aquatic invertebrates, such as developmental delays in
the microcrustacean Tigriopus japonicus. The study involved acute exposure to
concentrations raning from 5000 to 20,000 μg L-1 (Kang et al., 2019). In another study
with the sea urchin Paracentrotus lividus, PrP exposure decreased larval length and
induced several morphological changes. The experiment utilized acute exposure with

72
concentrations ranging from 10 to 10,000 μg L-1 (Torres et al., 2016). Therefore, all
these data suggest that paraben exposure, particularly to those with longer alkyl chains
and higher concentrations, may compromise the development of aquatic animals.
1.5.4.2 Effects on endocrine system
The initial biomarkers used in experimental studies involving fish were
primarily focused on endocrine disruption, particularly related to sex hormones and
induction of vitellogenin (vtg) secretion. Pedersen et al. (2000) performed a bioassay
using EtP, PrP, BuP, and 4-HB in juveniles of O. mykiss through an intraperitoneal
injection dose of 1 mg kg-1, as the exposure method. Blood samples were obtained on
days 0, 6 and 12 after the injection. The results indicated that all the tested parabens
presented estrogenicity (increased plasma levels of vtg) in doses between 100 and
300 mg kg-1 (Pedersen et al., 2000). However, according to these authors, PrP and
BuP were approximately six times more potent than EtP, having an estrogenic potential
comparable to that previously reported for the plastic additive bisphenol A. It is
important to mention that in this study, neither the route of exposure nor the
concentrations used were meant to be environmentally realistic.
In the following studies, using O. mykiss as a model organism, Bjerregaard
et al. (2003) and Alslev et al. (2005) tested the potential estrogenicity of PrP and BuP,
respectively, as depicted in section 5.3. Both studies used the oral administration
protocol of the paraben by food gavage. These studies found that the vtg synthesis in
O. mykiss was a more sensitive biomarker of estrogenic activity, which may have
opened the path for more experimental studies using this biomarker in other fish
species. It was also found that orally administered BuP is quickly metabolized by the
liver and that water exposure also leads to detectable concentrations of BuP in fish
bloodstream. In addition, the authors pointed out that absorption of parabens through
water exposure might be more damaging to the HPG axis due to the lack of
first-passage effect involved in oral administration (Alslev et al., 2005).
Subsequently, Barse et al. (2010) also detected an increased vtg production
in adult male common carp (C. carpio) exposed to three MeP concentrations (0.84,
1.68, and 4.20 mg L-1). The first study where vtg induction was assessed by gene

73
expression was developed by Yamamoto et al. (2011) in O. latipes. As from liver
microarray analyses, they detected upregulation of vtg and choriogenin (chg) genes
after exposure to 10 μg L-1 of MeP. Additionally, using the medaka vitellogenin test,
they found increased vtg plasma levels in fish exposed to 630 μg L-1.
Since then, endocrine disruption has been reported in fish in many studies.
It has been observed in zebrafish larvae that MeP, EtP, PrP, and BuP (100, 200, 400,
800 and 1000 µM) exposure during 96 h (Dambal et al., 2017) or to different
concentrations (20 and 100 µM of MeP, 20 and 50 µM of EtP, 2 µM of PrP and 1 and 2
µM of BuP) during to 120 h (Liang et al., 2023a), promotes vtg upregulation,
demonstrating its estrogenic effect. Nonetheless, an opposite effect was observed in
adult zebrafish, in which they detected a reduction in vtg levels in males and females
exposed to MeP (1, 3, and 10 μg L-1) for 28 days (Hu et al., 2023a,b). Additionally,
such exposures also alter the gene expression of key factors related to the
hypothalamic-pituitary-gonadal axis in zebrafish larvae and adults, disrupting the
secretion of steroid hormones (Hu et al., 2023b; Liang et al., 2023a), which can have
negative consequences for the sexual differentiation, gonad differentiation and also in
the reproductive success of adult animals.
Furthermore, recent studies evidenced that paraben exposure can also
disrupt other endocrine axes. Using embryo-larval stages of zebrafish exposed to MeP
(20–200 µM), EtP (20–100 µM), PrP (5–20 µM), and BuP (2 µM) for 120h, a reduction
in thyroid hormone levels (T3 and T4) was detected (Liang et al., 2022). The
transcription levels of several target genes along the hypothalamic-pituitary-thyroid
(HPT) axis (Liang et al., 2022) also was disturbed. Using molecular docking, the
authors also showed that all tested parabens acted as thyroid receptor agonists (Liang
et al., 2022).
Parabens can also disrupt the HPA axis. In adults of zebrafish exposed to
environmentally realistic concentrations of MeP (1, 3, and 10 μg L-1) for 28 days, both
sexes exhibited alterations in the HPA axis activity, including reduced transcription
levels of corticotropin-releasing hormone (CRH) and its binding protein, as well as
decreased blood cortisol concentrations (Hu et al., 2023a). Similarly, the exposure of
zebrafish embryos and larvae to MeP, EtP, PrP, and BuP during 120 h increased

74
adrenocorticotropic hormone (ACTH) levels and reduced cortisol levels, as well as
upregulates several genes related to stress response signaling (Liang et al., 2023b).
Also, the activation of zebrafish glucocorticoid receptors (Gr) by parabens was
demonstrated in silico and in vitro (Liang et al., 2023b). Brown et al. (2018) also
investigated the effects of BuP exposure on zebrafish development, focusing on the
development of the endocrine pancreas, using insulin: GFP transgenic zebrafish
embryos exposed to BuP (250, 500, 1,000, and 3000 nM), over 7 days. After 96 h of
exposure, they detected an increase in pancreatic islet area at the lowest BuP
concentration, with 70% of islets presenting variant morphology, as fragmented islet
clusters and ectopic beta cells. Also, they found alterations in GSH content and the
transcripts of GSH-related genes (Brown et al. (2018), evidencing that BuP exposure
affects the development of pancreatic islets and disrupts the redox balance, resulting in
several developmental abnormalities. Such findings indicate that paraben exposure
may affect multiple endocrine axes at different stages of the fish life cycle and act as
an EDC. Notwithstanding, it is important that future studies also evaluate the protein
levels of the various hormones and receptors involved, as well as other proteins
related to the several endocrine axes, which can also be linked to functional studies, to
fully clarify the mechanisms of endocrine disruption triggered by parabens.
Furthermore, the extent to which parabens may impact the endocrine axes in
neotropical fish species or even in other groups of aquatic organisms remains unclear.
1.5.4.3 Effects on reproductive system and reproduction
Using a histopathological approach in zebrafish testes, Hassanzadeh (2017)
revealed that chronic exposure (21 days) to MeP (0.001–10 mg L-1) decreased the
gonadosomatic index (GSI), leading to testicular atrophy, germ cell impairments,
proliferation in spermatogonia, and decrease in the proportion of the spermatozoa,
Leydig cell hyperplasia, interstitial fibrosis, and apoptosis of Sertoli cells. However, no
hormone analysis was carried out. Such results are in agreement with those found by
Barse et al. (2010) in males of C. carpio, in which all MeP concentrations tested (0.84,
1.68, and 4.2 mg L-1) reduced the GSI, increased the occurrence of areas with

75
inflammatory infiltrate and fibrosis, and reduced the interstitial compartment and the
number of sperm in the lumen of the seminiferous tubules.
In another study conducted by Hu et al. (2023b), both males and females of
zebrafish exposed to MeP for 28 days exhibited higher GSI only at the highest tested
concentration (10 μg L-1), while all treatments (1, 3, and 10 μg L-1) resulted in a
blockade of gametogenesis, accompanied by an imbalance in sex hormone (low levels
for T, 11-keto testosterone and estradiol). By evaluating the expression of several
genes related to the HPG axis, a disturbance in steroidogenesis and feedback
regulation mechanisms was demonstrated, evidencing an antiestrogenic activity for
MeP (Hu et al., 2023b). In males, all concentrations inhibited the spermatogenesis
process, increasing the percentage of spermatogonia and spermatocytes, but reducing
spermatozoa. In females, the low estradiol levels downregulated the hepatic
expression of VTG, resulting in a deficiency in production and, consequently, affecting
the vitellogenesis process in the ovary (Hu et al., 2023b). Interestingly, this did not
reflect on egg production, egg weight and protein, but induced a high mortality rate in
larvae and also precocious hatching of offspring larvae derived from females exposed
to 3 and 10 μg L-1 groups (Hu et al., 2023b). These data provided important insights
about the effects of MeP on fish reproductive systems. However, there are still no
studies on the effects of parabens on the quality of gametes in fish and other aquatic
organisms. Also, Hu et al. (2022c) suggested the occurrence of maternal transfer of
parabens to the offspring, which may interfere with the survival and development of the
offspring. To our knowledge, the bioaccumulation of parabens has already been
evidenced only in fish ovaries (Peng et al., 2018) and testes (Barse et al., 2010).
1.5.4.4 Effects on nervous system and behavior
Zebrafish exposed to MeP (0.1, 1, 10, and 100 ppb) during the
embryo-larval stages (over 144 h) exhibited inhibited acetylcholinesterase (AChE)
activity in all treatments, as well elevated cortisol levels and induced anxiety-like
behavior in the two MeP lowest concentrations (Luzeena-Raja et al., 2019). In addition
to these results, Merola et al. (2020a) also reported that embryos exposed to 10 and
30 mg L-1 presented a low number of spontaneous contractions, which is used as a

76
biomarker of neurotoxicity in zebrafish embryos. In adult zebrafish exposed to MeP (1,
11 and 110 ppb of the LC50, over 30 days), Thakkar et al. (2022) found a
concentration-dependent reduction in AChE activity in the brain for both sexes
following chronic sub-lethal methylparaben exposure and altered the expression of two
genes involved in neuronal differentiation (ntrk2a and pax6b). Nonetheless,
sex-specific responses in the brain have also been observed in adult zebrafish.
Females showed an increase in serotonin levels, while males showed a decrease
(Thakkar et al., 2022). Similarly, Hu et al., 2023a reported enhanced glutamatergic
neural signaling in the male zebrafish brain exposed to MeP (1, 3, and 10 mg L-1,
during 28 days), while blockage of synaptic neurotransmission was observed in
females.
O. niloticus (with no information regarding the sex or life cycle stage of the
specimens used) exposed to BuP (5, 50, 500, and 5000 ng -1, over 56 days)
demonstrated reduced dopamine and γ-aminobutyric acid content in the brain, along
with the induction of pathways related to skin pigmentation (Liu et al., 2023a). To date,
no studies have evaluated whether exposure to parabens can affect the mechanisms
of cell proliferation and death (apoptosis) in the nervous system of aquatic organisms,
nor have they investigated specific brain regions.
1.5.4.5 Effects on the digestive system
The effects of MeP have already been investigated in adult zebrafish,
focusing on morphological and physiological biomarkers in the gut and its microbiome.
De Carvalho Penha et al. (2021) conducted an innovative study introducing an
intestinal microbiota analysis through the evaluation of the use of carbon sources by
the microbial community. After the exposure of adult male zebrafish to 30 and 50 mg
L-1 of MeP for 96 h, no significant changes were observed in the diversity or
abundance of gut microbiota, despite concerns about MeP's antimicrobial properties.
However, the animals had an increase in carbon source utilization, suggesting a
potential metabolic adaptation to MeP exposure (De Carvalho Penha et al., 2021).
Furthermore, Hu et al. (2022b) exposed zebrafish to environmentally
relevant concentrations of MeP (1, 3, and 10 μg L-1) for 28 days to assess its impact on

77
gut microbiota and overall health. High-throughput amplicon sequencing revealed that
subchronic MeP exposure significantly disrupted the composition and diversity of the
gut microbial community. Interestingly, MeP caused sex-specific intestinal effects:
males exhibited increased goblet cell density, elevated tight junction protein (Tjp2)
expression, and higher serotonin levels, while females showed reduced goblet cell
density, reduced Tjp2 expression, and decreased serotonin levels, alongside
up-regulated pro-inflammatory cytokines transcription. Additionally, intestinal catalase
(CAT) activity was elevated under MeP stress, contributing to oxidative stress
mitigation (Hu et al., 2022b). These findings showed how MeP exposure may impair
gut barrier function and intestinal health, emphasizing the need for risk assessments of
this contaminant. Another study using the same MeP concentrations and exposure
time in adult zebrafish conducted by Hu et al. (2022a) demonstrated that MeP
subchronic exposure induced hepatotoxic effects, such as hepatocellular vacuolization,
changes in antioxidant system and lipid metabolism and, interesting, also promotes
increased cortisol levels in the liver and as well as inhibiting the synthesis and
conjugation of primary bile acid. The authors also performed a metabolomic analysis,
which showed that MeP mainly alters the composition of fatty acids, retinoids, and
steroids (Hu et al., 2022a). Similarly, Barse et al. (2010) also detected an increase in
vacuoles and areas of focal necrosis in the liver of C. carpio exposed to the higher
MeP concentrations (1.68 and 4.2 mg L-1).
Ding et al. (2023) investigated the effects in the gut of the turtle M. sinensis
exposed to several BuP concentrations (5, 50, and 500 μg L-1) over 20 weeks. The
results revealed changes in gut microbiome composition, with the genus Edwardsiella
emerging uniquely in BuP-exposed turtles, particularly absent in the control group.
Structural changes in the intestines included shortened villi, a thinner muscular layer,
and a marked reduction in goblet cell numbers. Immune responses were notably
affected, with increased numbers of neutrophils and natural killer cells in the lamina
propria of the intestinal mucosa, especially at higher BuP concentrations (500 μg L-1).
This was accompanied by a marked upregulation of pro-inflammatory cytokines,
particularly IL-1β, which was strongly correlated with the abundance of Edwardsiella.
They also showed that the presence of Edwardsiella was inversely related to goblet

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cell counts, further indicating compromised gut barrier function (Ding et al., 2023).
These findings highlight the impacts of BuP, a long alkyl chain paraben, on intestinal
homeostasis in aquatic turtles, as it induces gut dysbiosis, triggers inflammatory
responses, and weakens the physical gut barrier (Ding et al., 2023), demonstrating that
parabens also may promote toxicity to the digestive tract and affect the organism
systemically.
1.5.4.6 Effects on metabolism
Oxidative stress, primarily detected in the liver due to its involvement in the
degradation of xenobiotics, has been extensively investigated in response to exposure
to various parabens. Silva et al. (2018) observed an adaptive response in O. niloticus
exposed to 4 mg L-1 of MeP, EtP, PrP, and BuP over 6 and 12 days. CAT activity only
was increased in the liver of animals exposed to MeP and EtP for six days. After 6
days of exposure, decreased glutathione (GSH) levels in the liver and gills were seen.
However, after 12 days, the levels increased compared to controls, suggesting an
adaptation of the antioxidant response in animals exposed to sublethal concentrations
of parabens. The malondialdehyde (MDA) content was only increased in animals
exposed to EtP and BuP for 12 days (Silva et al., 2018). In a second study in O.
niloticus, but in juveniles exposed to BzP (5, 50, 500, and 5000 ng L-1) for eight
weeks, Lin et al. (2022) detected metabolic disorders in hepatic glycerol phospholipids,
glycerolipids, and sphingomyelins, as well as increased crude fat content, oxidative
stress, and liver tissue inflammation. In the same way, Lite et al. (2022) focused their
study on oxidative stress and found lower levels of antioxidant enzymes and GSH in
zebrafish exposed to PrP and BuP (0.1, 1, and 10 ppb) over 96 h.
In adult zebrafish exposed to 30 and 50 mg L-1 of MeP over 96 h, an
increase of MDA levels and for ethoxyresorufin O-deethylase activity (EROD) activity
were detected in gills at the higher treatment, but no changes in the liver, as well as no
lipid peroxidation, were detected in larvae exposed to 30 and 60 mg L-1 over 168 h
(Carvalho Penha et al., 2021). However, in a second study in zebrafish adults, but in a
subchronic exposure over 28 days, Hu et al. (2022a) detected elevated hepatic ROS
levels and GPX activity in females exposed to 10 μg L-1 of MeP, in parallel to reduced

79
activity for CAT and for GSH content. Also, an increase in HSI was seen in females
exposed to 1 and 10 μg L-1 of MeP, but not for the intermediate concentration of 3 μg
L-1 (Hu et al., 2022a). Furthermore, the effects of parabens exposure have also been
detected along the embryo larval development in zebrafish. After exposure to 50 mg L-1
of MeP over 72 h, Ates et al. (2018) detected a reduction in GST and NO levels, but a
slight increase in lipid peroxidation was observed. However, De Carvalho Penha et al.
(2021) did not detect changes in lipid peroxidation in zebrafish larvae exposed to
higher MeP concentrations (30 and 60 mg L-1) and a longer exposure time (168 h).
Interestingly, at the transcriptional level, Bereketoglu & Pradhan (2019) found that,
depending on paraben tested (MeP or PrP) and its concentration, they can alter
several genes related to the oxidative stress (nrf2, keap1, gst, mgst, sod1, hsp70, mt1)
and fatty acid metabolism (apoab, apoeb, apoa4, fasn, ldlr, lpl, lipc), as well to
apoptosis (bax, bcl2, casp3a, dap3), cell proliferation (p21, p38), to DNA damage
pathways (gadd45a, rad51, apex1 and xpc), to endocrine (ar, esr2a, thraa and thrb)
and immune function (tnfa and il8) (Bereketoglu & Pradhan (2019), drawing attention to
the fact that exposure to parabens can systemically impact animal physiology,
negatively affecting the zebrafish development.
The alanine aminotransferase (ALT) was also evaluated in zebrafish blood
samples by Hu et al. (2022a), since this enzyme is released by the liver after injury to
hepatocytes (Liu et al., 2014). An increased ALT activity was observed in males and
females exposed to the intermediate concentration of MeP tested (3 μg L-1). In
addition, Barse et al. (2010) also evaluated the activity of hepatic enzymes in muscle
tissue of C. carpio exposed to MeP (0.84, 1.68, and 4.2 mg L-1, over 28 days), being
found an increase in alkaline phosphatase (ALK) and alanine aminotransferase (ALN),
but a reduction on acid phosphatase (ACP) and aspartate aminotransferase (AST)
activity, which denotes a change in muscle tissue metabolism. Furthermore, Yin et al.
(2023) investigated the effects of BuP (5, 50, and 500 μg L-1, over a 20-week) on the
liver of the turtle M. sinensis. They detected a decreased antioxidant enzyme activity
(SOD, CAT, GSH-Px) in the liver, as well as increased levels of MDA at the highest
concentrations tested, signaling compromised oxidative defense and oxidative
damages. The Nrf2-Keap1 pathway-related genes, initially upregulated, declined at

80
higher concentrations, indicating oxidative stress overload. Heat shock proteins
(HSP70, HSP90) and inflammatory markers (BAFF, IL-6, P50, P65) were elevated,
suggesting cellular stress and inflammation. BuP exposure also induced apoptosis,
with increased pro-apoptotic gene expression (BAX, CytC, Caspase3 and Caspase9)
and decreased anti-apoptotic Bcl2 levels (Yin et al., 2023). Collectively, such data
indicate that exposure to parabens may also interfere with the metabolism of aquatic
organisms, as well such biomarkers are important to understand the ecotoxicological
effects caused by them.

1.5.5 Types, exposure pathways, and analytical monitoring
The exposure methods employed in laboratory studies are crucial as they
can significantly impact the outcomes and interpretations of the research (Zhang et al.,
2019). These methods can be broadly categorized into acute and chronic exposures,
each serving different purposes and offering unique insights depending on the study's
aims. In this review, most of the studies used acute exposure methods. The prevalence
of acute studies can be attributed to several factors, such as faster results and
cost-effectiveness, as well as the fact that many of the experimental studies employ
the initial stages of zebrafish development, as validated by Guide no. 239, Fish Acute
Embryo Toxicity (FET) Test by OECD (2013). Acute exposure studies, which are
shorter in duration, focus on immediate toxic effects and help identify lethal
concentrations and immediate responses. However, acute exposure studies have
limitations. The short duration may not capture delayed effects or the cumulative
impact of a substance, such as those investigated throughout male and female
gametogenesis, potentially overlooking long-term consequences (Erhirhie et al., 2018).
In contrast, chronic exposure involves prolonged exposure to a substance or condition,
often at lower levels, to simulate real-world scenarios more closely (Silva et al., 2020).
Chronic exposure studies are essential for understanding the long-term effects and
potential continuous or repeated exposure risks (Zhang et al., 2019). The choice
between acute and chronic exposure methods depends on each study's specific
objectives and experimental models.

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Considering the routes of exposure, interestingly, the initial studies
employed exposure to parabens exposure through intraperitoneal injection or by food
using the gavage method (Pedersen et al., 2000; Bjerregaard et al., 2003; Alslev et al.,
2005). However, these substances can be metabolized by the liver when administered
orally, which probably interferes with the absorption rates and generates higher
sensitivity than in experiments with mice (Bjerregaard et al., 2003). From this, most
studies have been carried out with vertebrates via exposure to water. Regarding this,
most studies have employed the exposure using a static system (n = 41; 71.9 %),
rather than semi-static (with renewal, n = 14; 24.5%) or flow-through system (n = 2;
3.50%). Since paraben's half-life ranges from minutes to days varying according to the
physicochemical conditions considered (Bledzka et al., 2014; González-Mariño et al.,
2011b; Wu et al., 2017), depending on the model species adopted and the duration of
the test, a semi-static system should preferably be adopted, where water changes
occur every 24–48 h during the exposure time, since in this type of bioassay the
concentration of the chemical under study is expected to remain within ±20% of the
nominal values, ensuring the experimental results' validity (Erhirhie et al., 2018). In
addition, a few experimental studies (n = 10; 17.5%) measured the paraben
concentration in their treatments, denoting that further experimental studies related to
paraben ecotoxicity should be concerned with this type of analytical validation, as
recommended by OECD (2013) and also by the leading journals in the field of
ecotoxicology.
In addition to the traditional exposure methods, recent advances in
analytical techniques, such as magnetic solid-phase extraction (MSPE), have
enhanced the detection and quantification of personal care products, including
parabens, in environmental samples. A novel porphyrin-based magnetic covalent
organic framework (PCOF) has demonstrated high efficiency in extracting a wide range
of analytes, with log Kow values ranging from 1.96 to 7.60. This method, which offers
enhanced extraction efficiency due to the COF's functional groups, has shown great
promise in identifying parabens in aquatic environments with low detection limits
(0.4–0.9 ng mL-1) (Ning et al., 2023).

82
1.5.6 Environmental relevance of experimental studies and confounding factors
Although this review does not aim to compare paraben concentrations
previously detected in distinct aquatic environments across the world with those used
in the several laboratory studies reviews, it is important for future research to address
this gap. Such comparisons would enhance the realism of environmental risk
assessments. Paraben concentrations in bioassays published until now vary widely,
from a few nanograms per liter (ng L-1) (Lin et al., 2022) to several milligrams per liter
(mg L-1) (Fan et al., 2022). This variability likely reflects differences in study objectives,
as human exposure to parabens tends to be higher than that of aquatic organisms
(Bledzka et al., 2014). Moreover, inconsistencies in measurement units in different
studies, such as molar concentrations (Dambal et al., 2017; Bereketoglu and Pradhan,
2019) and parts per billion (ppb) (Luzeena Raja et al., 2019; Thakkar et al., 2022),
complicate direct comparisons of tested concentrations. The diversity of experimental
protocols, the lack of standardization in exposure durations, and the imprecision in
measuring paraben concentrations during in vivo studies further complicates the
interpretation of these studies.
1.6 Conclusions and future perspectives
This is the first study to review and systematize the current knowledge
about the bioaccumulation and ecotoxicity of parabens in aquatic organisms. Our
review indicates that parabens, particularly MeP, PrP, BuP, and their derivatives, can
adversely affect aquatic organisms. Parabens can induce endocrine disruption,
oxidative stress, and developmental and reproductive impairments. The ecotoxicity of
parabens varies based on species, life cycle stage, and experimental design, such as
exposure period, concentrations tested, and type of exposure. Some studies suggest
that certain organisms, such as fish and crustaceans, exhibit greater susceptibility to
paraben effects than others. Additionally, the persistence of parabens in the
environment and their potential for bioaccumulation in organisms raise significant
concerns regarding their long-term impact on aquatic ecosystems, mainly due to the
potential growth of the world population and the use of paraben-based products. While
further research is necessary to comprehensively understand paraben ecotoxicity in

83
aquatic organisms using environmentally relevant concentrations of parabens, the
current evidence highlights the need for more stringent monitoring and regulation of
these chemicals. Given their status as chemicals of emerging pollutants, enhancing
regulatory frameworks to mitigate parabens' risks to aquatic ecosystems is imperative.
For a more comprehensive understanding of bioaccumulation and ecotoxicity of
parabens in aquatic organisms, future research should focus on the key aspects listed
below.
a)​

Standardizing experimental protocols related to paraben bioassays

among different research groups;
b)​

Perform

laboratory

assays

using

environmentally

relevant

concentrations of parabens, their metabolites, and chlorinated derivatives;
c) Evaluate paraben ecotoxicity in representative aquatic species from
different taxa;
d) Evaluate paraben bioaccumulation in samples of aquatic organisms from
different aquatic environments, and also from distinct geographic regions or
countries;
e) Combine multiple biomarkers;
f) In addition to gene expression analyses, also adopt the quantification of
proteins and other biomolecules or their metabolites in experimental studies
(Multi-omics approaches);
g) Conduct multi- and transgenerational studies;
h) Measure parents' parabens and their metabolites in tissue samples from
field and laboratory studies;
i) Develop analytical methods to measure paraben concentrations in
biological samples with small volumes, such as for zebrafish organs;
j) Measure the real paraben concentration in water throughout the bioassay;
k) Assessment of the interactive effects of parabens with traditional and
emerging pollutants;
l) Analyzing the effects of parabens on microbiota from aquatic organisms.

84
Such efforts will improve the relevance and applicability of experimental
findings to real-world scenarios and increase our understanding of the ecotoxicity of
parabens in aquatic organisms.

CRediT authorship contribution statement
Felipe Felix Costa Lima da Silveira: Writing – review & editing, Writing – original
draft, Visualization, Methodology, Investigation, Formal analysis, Data curation,
Conceptualization. Viviane Amaral Porto: Writing – review & editing, Writing – original
draft, Visualization, Methodology, Investigation, Formal analysis, Data curation. Bianca
Leite Carnib de Sousa: Writing – original draft, Visualization, Software, Formal
analysis, Data curation. Emilly Valentim de Souza: Writing – original draft, Software,
Formal analysis, Data curation. Fabiana Laura Lo Nostro: Writing – review & editing,
Writing – original draft, Investigation, Formal analysis. Thiago Lopes Rocha: Writing –
review & editing, Writing – original draft, Methodology, Formal analysis. Lázaro
Wender Oliveira de Jesus: Writing – review & editing, Writing – original draft,
Visualization, Supervision, Resources, Project administration, Methodology, Funding
acquisition, Data curation, Conceptualization.
Declaration of competing interest
The authors declare that they have no conflict of interest.
Data availability
All data used for the research are described or cited in the article.
Acknowledgment
The authors would like to thank the Alagoas Research Foundation (FAPEAL) for
funding granted to Dr. Lázaro Wender Oliveira de Jesus (E:60030.0000000161/2022).
Also, to the Coordination for the Improvement of Higher Education Personnel (CAPES)
for the Master's Degree scholarship granted to Felipe Félix Costa Lima da Silveira. Dr.

85
Thiago Lopes Rocha was granted a productivity fellowship (n. 306329/2020-4) from
National Council for Scientific and Technological Development (CNPq).
REFERENCES
Albero, B., Pérez, R. A., Sánchez-Brunete, C., & Tadeo, J. L. (2012). Occurrence and
analysis of parabens in municipal sewage sludge from wastewater treatment plants in
Madrid
(Spain).
Journal
of
Hazardous
Materials,
239-240,
48–55.
https://doi.org/10.1016/j.jhazmat.2012.05.017
Ali, Z., Ahmad, V. U., Zahid, M., & Tareen, R. B. (1998). Benzoic acid derivatives from
Stocksia
brahuica.
Phytochemistry,
48(7),
1271-1273.
https://doi.org/10.1016/S0031-9422(98)00124-1
Alslev, B., Korsgaard, B., & Bjerregaard, P. (2005). Estrogenicity of butylparaben in
rainbow trout Oncorhynchus mykiss exposed via food and water. Aquatic Toxicology,
72(4), 295–304. https://doi.org/10.1016/j.aquatox.2005.01.005
Andersen, F. A. (2008). Final amended report on the safety assessment of
methylparaben, ethylparaben, propylparaben, isopropylparaben, butylparaben,
isobutylparaben, and benzylparaben as used in cosmetic products. International
Journal of Toxicology, 27(Suppl 4), 1-82.
Ankley, G. T., Bennett, R. S., Erickson, R. J., Hoff, D. J., Hornung, M. W., Johnson, R.
D., Mount, D. R., Nickols, J. W., Russom, C. L., Schmieder, P. K., Serrano, J. A.,
Tietge, J. E., Villeneuve, D. L. (2009). Adverse outcome pathways: A conceptual
framework to support ecotoxicology research and risk assessment. Environmental
Toxicology and Chemistry, 29(3), 730-741. https://doi.org/10.1002/etc.34
Ateş, P. S., Ünal, İ., Üstündağ, Ü. V., Alturfan, A. A., Yiğitbaşı, T., & Emekli-Alturfan, E.
(2018). Methylparaben induces malformations and alterations on apoptosis,
oxidant-antioxidant status, ccnd1 and myca expressions in zebrafish embryos. Journal
of
Biochemical
and
Molecular
Toxicology,
32(3),
e22036.
https://doi.org/10.1002/jbt.22036
Bambino, K., & Chu, J. (2017). Zebrafish in toxicology and environmental health.
Current
topics
in
developmental
biology,
124,
331-367.
https://doi.org/10.1016/bs.ctdb.2016.10.007
Barse, A. V., Chakrabarti, T., Ghosh, T. K., Pal, A. K., Kumar, N., Raman, R. P., &
Jadhao, S. B. (2010). Vitellogenin induction and histo-metabolic changes following

86
exposure of Cyprinus carpio to methyl paraben. Asian-Australasian Journal of Animal
Sciences, 23(12), 1557–1565. https://doi.org/10.5713/ajas.2010.10118
Bereketoglu, C., & Pradhan, A. (2019). Comparative transcriptional analysis of
methylparaben and propylparaben in zebrafish. Science of the Total Environment, 671,
129–139. https://doi.org/10.1016/j.scitotenv.2019.03.358
Bjerregaard, P., Andersen, D. N., Pedersen, K. L., Pedersen, S. N., & Korsgaard, B.
(2003). Estrogenic effect of propylparaben (propylhydroxybenzoate) in rainbow trout
Oncorhynchus mykiss after exposure via food and water. Comparative Biochemistry
and Physiology Part C: Toxicology & Pharmacology, 136(4), 309–317.
https://doi.org/10.1016/j.cca.2003.10.004
Błędzka, D., Gromadzińska, J. & Wąsowicz, W. (2014) Parabens: From environmental
studies
to
human
health.
Environment
International,
67,
27-42.
https://doi.org/10.1016/j.envint.2014.02.007.
Bolujoko, N. B., Unuabonah, E. I., Alfred, M. O., Ogunlaja, A., Ogunlaja, O. O.,
Omorogie, M. O., & Olukanni, O. D. (2021). Toxicity and removal of parabens from
water: A critical review. Science of the Total Environment, 792, 148092.
https://doi.org/10.1016/j.scitotenv.2021.148092
Bom, F. C., & Sá, F. (2021). Concentration of microplastics in bivalves of the
environment: a systematic review. Environmental Monitoring and Assessment, 193,
1-30. https://doi.org/10.1007/s10661-021-09639-1
Brausch, J. M., & Rand, G. M. (2011). A review of personal care products in the
aquatic environment: environmental concentrations and toxicity. Chemosphere, 82(11),
1518-1532. https://doi.org/10.1016/j.chemosphere.2010.11.018
Brown, S. E., Sant, K. E., Fleischman, S. M., Venezia, O., Roy, M. A., Zhao, L., &
Timme-Laragy, A. R. (2018). Pancreatic beta cells are a sensitive target of embryonic
exposure to butylparaben in zebrafish (Danio rerio). Birth Defects Research, 110(11),
933–948. https://doi.org/10.1002/bdr2.1215
Cacua-Ortiz, S. M., Aguirre, N. J., & Peñuela, G. A. (2020). Methyl paraben and
carbamazepine in water and striped catfish (Pseudoplatystoma magdaleniatum) in the
cauca and magdalena rivers. Bulletin of Environmental Contamination and Toxicology,
105(6), 819–826. https://doi.org/10.1007/s00128-020-03028-z
Canedo, A., de Jesus, L. W. O., Bailão, E. F. L. C., & Rocha, T. L. (2021). Micronucleus
test and nuclear abnormality assay in zebrafish (Danio rerio): Past, present, and future

87
trends. Environmental pollution (Barking,
https://doi.org/10.1016/j.envpol.2021.118019

Essex

:

1987),

290,

118019.

Chiesa, L. M., Pavlovic, R., Panseri, S., & Arioli, F. (2018). Evaluation of parabens and
their metabolites in fish and fish products: a comprehensive analytical approach using
LC-HRMS. Food Additives & Contaminants: Part A, 35(12), 2400-2413.
https://doi.org/10.1080/19440049.2018.1544721
Comeche, A., Martín-Villamil, M., Picó, Y., & Varó, I. (2017). Effect of methylparaben in
Artemia franciscana. Comparative Biochemistry and Physiology Part C: Toxicology &
Pharmacology, 199, 98–105. https://doi.org/10.1016/j.cbpc.2017.04.004
Crovetto, S. I., Moreno, E., Dib, A. L., Espigares, M., & Espigares, E. (2017). Bacterial
toxicity testing and antibacterial activity of parabens. Toxicological & Environmental
Chemistry, 99(5-6), 858–868. https://doi.org/10.1080/02772248.2017.1300905
Dailianis, S., Vlastos, D., Zoppou, C., Moschopoulou, A., & Antonopoulou, M. (2023).
Different isoforms of parabens into marine environment: Biological effects on the
bacterium Aliivibrio fischeri and the marine mussel Mytilus galloprovincialis. Science of
The Total Environment, 900, 165902. https://doi.org/10.1016/j.scitotenv.2023.165902
Dambal, V. Y., Selvan, K. P., Lite, C., Barathi, S., & Santosh, W. (2017). Developmental
toxicity and induction of vitellogenin in embryo-larval stages of zebrafish (Danio rerio)
exposed to methyl Paraben. Ecotoxicology and Environmental Safety, 141, 113–118.
https://doi.org/10.1016/j.ecoenv.2017.02.048
De Carvalho Penha, L. C., Coimbra Rola, R., da Silva Junior, F. M., & de Martinez
Gaspar Martins, C. (2021). Toxicity and sublethal effects of methylparaben on
zebrafish (Danio rerio) larvae and adults. Environmental Science and Pollution
Research, 28(33), 45534–45544. https://doi.org/10.1007/s11356-021-12800-5
De Felice, F., & Polimeni, A. (2020). Coronavirus disease (COVID-19): a machine
learning
bibliometric
analysis.
in
vivo,
34(3
suppl),
1613-1617.
https://doi.org/10.21873/invivo.11951
Di Poi, C., Costil, K., Bouchart, V., & Halm-Lemeille, M.-P. (2017). Toxicity assessment
of five emerging pollutants, alone and in binary or ternary mixtures, towards three
aquatic organisms. Environmental Science and Pollution Research, 25(7), 6122–6134.
https://doi.org/10.1007/s11356-017-9306-9
Ding, L., Wu, X., Lin, J., Zhang, J., Shi, H., Hong, M., & Fang, Z. (2023). Butylparaben
disordered intestinal homeostasis in Chinese striped-necked turtles (Mauremys

88
sinensis).
Ecotoxicology
and
Environmental
https://doi.org/10.1016/j.ecoenv.2023.115193

Safety,

262,

115193.

Dobbins, L. L., Usenko, S., Brain, R. A., & Brooks, B. W. (2009). Probabilistic
ecological hazard assessment of parabens using Daphnia magna and Pimephales
promelas.
Environmental
Toxicology
and
Chemistry,
28(12),
2744.
https://doi.org/10.1897/08-523.1
Eghan, K., Lee, S., & Kim, W. K. (2023). Cardio-and neuro-toxic effects of four
parabens on Daphnia magna. Ecotoxicology and Environmental Safety, 268, 115670.
https://doi.org/10.1016/j.ecoenv.2023.115670
Erhirhie, E. O., Ihekwereme, C. P., & Ilodigwe, E. E. (2018). Advances in acute toxicity
testing: strengths, weaknesses and regulatory acceptance. Interdisciplinary toxicology,
11(1), 5-12. https://doi.org/10.2478/intox-2018-0001
Escher, B. I., Neale, P. A., & Villeneuve, D. L. (2018). The advantages of linear
concentration–response curves for in vitro bioassays with environmental samples.
Environmental
toxicology
and
chemistry,
37(9),
2273-2280.
https://doi.org/10.1002/etc.4178
Fan, Z., Yang, Y., Hu, P., Huang, Y., He, L., Hu, R., Zhao, K., Zhang, H., & Liu, C.
(2022). Molecular mechanism of ethylparaben on zebrafish embryo cardiotoxicity
based on transcriptome analyses. Science of the Total Environment, 156785.
https://doi.org/10.1016/j.scitotenv.2022.156785
Graney, R. L., Giesy, J. P., & Clark, J. R. (2020). Field studies. In Fundamentals of
aquatic toxicology (pp. 257-305). CRC Press. https://doi.org/10.1201/9781003075363
González-Doncel, M., García-Mauriño, J. E., San Segundo, L., Beltrán, E. M., Sastre,
S., & Fernández Torija, C. (2014). Embryonic exposure of medaka (Oryzias latipes) to
propylparaben: Effects on early development and post-hatching growth. Environmental
Pollution, 184, 360–369. https://doi.org/10.1016/j.envpol.2013.09.022
González-Mariño, I., Quintana, J. B., Rodríguez, I., Schrader, S., & Moeder, M. (2011).
Fully automated determination of parabens, triclosan and methyl triclosan in
wastewater by microextraction by packed sorbents and gas chromatography-mass
spectrometry. Analytica
chimica
acta, 684(1-2),
50–57.
https://doi.org/10.1016/j.aca.2010.10.049
Haman, C., Dauchy, X., Rosin, C., & Munoz, J.-F. (2015). Occurrence, fate and
behavior of parabens in aquatic environments: A review. Water Research, 68, 1–11.
https://doi.org/10.1016/j.watres.2014.09.030

89
Hassanzadeh, N. (2017). Histopathological evaluation of the Zebrafish (Danio rerio)
testis following exposure to methylparaben. International Journal of Aquatic Biology,
5(2), 71-78. https://doi.org/10.22034/ijab.v5i2.245
Hu, C., Bai, Y., Sun, B., Tang, L., & Chen, L. (2022a). Significant impairment of
intestinal health in zebrafish after subchronic exposure to methylparaben. Science of
the Total Environment, 156389. https://doi.org/10.1016/j.scitotenv.2022.156389
Hu, C., Bai, Y., Sun, B., Zhou, X., & Chen, L. (2023). Exposure to methylparaben at
environmentally realistic concentrations significantly impairs neuronal health in adult
zebrafish. Journal of Environmental Sciences. https://doi.org/10.1016/j.jes.2022.07.012
Hu, C., Bai, Y., Li, J., Sun, B. & Chen, L (2023). Endocrine disruption and reproductive
impairment of methylparaben in adult zebrafish. Food And Chemical Toxicology, 171,
113545. http://dx.doi.org/10.1016/j.fct.2022.113545.
Haman, C., Dauchy, X., Rosin, C., & Munoz, J. F. (2015). Occurrence, fate and
behavior of parabens in aquatic environments: a review. Water Research, 68, 1-11.
https://doi.org/10.1016/j.watres.2014.09.030
Inui, M. (2003). Effect of UV screens and preservatives on vitellogenin and choriogenin
production in male medaka (Oryzias latipes). Toxicology, 194(1-2), 43–50.
https://doi.org/10.1016/s0300-483x(03)00340-8
Kang, H.-M., Kim, M.-S., Hwang, U.-K., Jeong, C.-B., & Lee, J.-S. (2019). Effects of
methylparaben, ethylparaben, and propylparaben on life parameters and sex ratio in
the marine copepod Tigriopus japonicus. Chemosphere, 226, 388–394.
https://doi.org/10.1016/j.chemosphere.2019.03.151
Kawashima, Y., Onishi, Y., Tatarazako, N., Yamamoto, H., Koshio, M., Oka, T., Horie,
Y., Watanabe, H., Nakamoto, T., Yamamoto, J., Ishikawa, H., Sato, T., Yamazaki, K., &
Iguchi, T. (2021). Summary of 17 chemicals evaluated by OECD TG229 using
Japanese Medaka, Oryzias latipes in EXTEND 2016. Journal of Applied Toxicology,
42(5), 750–777. https://doi.org/10.1002/jat.4255
Karthikraj, R., Vasu, A. K., Balakrishna, K., Sinha, R. K., & Kannan, K. (2017).
Occurrence and fate of parabens and their metabolites in five sewage treatment plants
in India. The Science of the total environment, 593-594, 592–598.
https://doi.org/10.1016/j.scitotenv.2017.03.173
Kusk, K. O., Krüger, T., Long, M., Taxvig, C., Lykkesfeldt, A. E., Frederiksen, H.,
Andersson, A.-M., Andersen, H. R., Hansen, K. M. S., Nellemann, C., &

90
Bonefeld-Jørgensen, E. C. (2010). Endocrine potency of wastewater: Contents of
endocrine disrupting chemicals and effects measured by in vivo and in vitro assays.
Environmental
Toxicology
and
Chemistry,
30(2),
413–426.
https://doi.org/10.1002/etc.385
Lee, J., Park, N., Kho, Y., Lee, K., & Ji, K. (2017). Phototoxicity and chronic toxicity of
methyl paraben and 1,2-hexanediol in Daphnia magna. Ecotoxicology, 26(1), 81–89.
Li, W., Shi, Y., Gao, L., Liu, J., & Cai, Y. (2015). Occurrence, fate and risk assessment
of parabens and their chlorinated derivatives in an advanced wastewater treatment
plant. Journal
of
hazardous
materials, 300,
29–38.
https://doi.org/10.1016/j.jhazmat.2015.06.060
Li, W., Sun, Y., Joseph, J., Fitzloff, J. F., Fong, H. H., & van Breemen, R. B. (2003).
p-Hydroxybenzoic acid alkyl esters in Andrographis paniculata herbs, commercial
extracts, and formulated products. Journal of agricultural and food chemistry, 51(2),
524-529. https://doi.org/10.1021/jf0258712
Li, W., Gao, L., Shi, Y., Wang, Y., Liu, J., & Cai, Y. (2016). Spatial distribution, temporal
variation and risks of parabens and their chlorinated derivatives in urban surface water
in Beijing, China. The Science of the total environment, 539, 262–270.
https://doi.org/10.1016/j.scitotenv.2015.08.150
Li, S., & Xia, M. (2019). Review of high-content screening applications in toxicology.
Archives of toxicology, 93, 3387-3396. https://doi.org/10.1007/s00204-019-02593-5
Li, M.H. (2020). Planarian mobility assay as a simple alternative for rapid screening
aquatic acute toxicity of personal care products. Fresenius Environmental Bulletin,
29(05),3362–3369. https://doi.org/10.1007/s10646-016-1743-6
Liang, J.; Liu, Q. S., Ren, Z., Min, K. Yang, X., Hao, F., Zhang, Q. Liu, Q., Zhou, Q. &
Jiang, G. (2023a). Studying paraben-induced estrogen receptor- and steroid
hormone-related endocrine disruption effects via multi-level approaches. Science Of
the Total Environment, 869, 161793. http://dx.doi.org/10.1016/j.scitotenv.2023.161793.
Liang, J., Yang, X., Liu, Q. S., Sun, Z., Ren, Z., Wang, X., Zhang, Q., Ren, X., Liu, X.,
Zhou, Q., & Jiang, G. (2022). Assessment of thyroid endocrine disruption effects of
parabens using in vivo, in vitro, and in silico approaches. Environmental Science &
Technology, 56(1), 460–469. https://doi.org/10.1021/acs.est.1c06562
Liang, J., Yang, X., Xiang, T., Chen, X., Ren, Z., Wang, X., Su, J., Zhang, Y., Liu, Q. S.,
Qu, G., Zhou, Q., & Jiang, G. (2023b). The perturbation of parabens on the

91
neuroendocrine system in zebrafish larvae. The Science of the total environment, 882,
163593. https://doi.org/10.1016/j.scitotenv.2023.163593
Liao, C., & Kannan, K. (2018). Temporal trends of parabens and their metabolites in
mollusks from the chinese bohai sea during 2006–2015: Species-specific accumulation
and implications for human exposure. Environmental Science & Technology, 52(16),
9045–9055. https://doi.org/10.1021/acs.est.8b02750
Lin, H., Jia, Y., Han, F., Xia, C., Zhao, Q., Zhang, J., & Li, E. (2022). Toxic effects of
waterborne benzylparaben on the growth, antioxidant capacity and lipid metabolism of
Nile tilapia (Oreochromis niloticus). Aquatic Toxicology, 248, 106197.
http://dx.doi.org/10.1016/j.aquatox.2022.106197.
Lin, Y., Zhao, Y., Liu, Y., Lan, Y., Zhu, J., Cai, Y., Guo, F., Li, F., Zhang, Y., Xu, Z. & Xue,
J. (2024). Occurrence and bioaccumulation of parabens and their metabolites in the
biota from a subtropical freshwater river ecosystem: Implications for human exposure.
Environmental Research, 240, 117530. https://doi.org/10.1016/j.envres.2023.117530
Lite, C., Guru, A., Juliet, M., & Arockiaraj, J. (2022). Embryonic exposure to
butylparaben and propylparaben induced developmental toxicity and triggered
anxiety‐like neurobehavioral response associated with oxidative stress and apoptosis
in
the
head
of
zebrafish
larvae.
Environmental
Toxicology.
https://doi.org/10.1002/tox.23545
Liu, S., Zhang, N., Liang, Z., Li, E., Wang, Y., Zhang, S. & Zhang, J. (2023a).
Butylparaben Exposure Induced Darker Skin Pigmentation in Nile Tilapia (Oreochromis
niloticus). Toxics, 11( 2), 119, 25 jan. 2023. https://doi.org/10.3390/toxics11020119
Liu, S., Wang, P., Wang, C., Chen, J., Wang, X., Hu, B., & Shan, X. (2023b). Disparate
toxicity mechanisms of parabens with different alkyl chain length in freshwater biofilms:
Ecological hazards associated with antibiotic resistome. Science of The Total
Environment, 881, 163168.https://doi.org/10.1016/j.scitotenv.2023.163168
Lück E, Jager M (1997) Esters of p-Hydroxybenzoic Acid. In: Lück E, Jager M (eds)
Antimicrobial food additives – characteristics, uses, effects. Springer-Verlag, Berlin, pp
183–189
Luzeena Raja, G., Divya Subhashree, K., Lite, C., Santosh, W., & Barathi, S. (2019).
Transient exposure of methylparaben to zebrafish (Danio rerio) embryos altered
cortisol level, acetylcholinesterase activity and induced anxiety-like behaviour. General
and
Comparative
Endocrinology,
279,
53–59.
https://doi.org/10.1016/j.ygcen.2018.11.001

92
Ma, X., Wan, Y., Wu, M., Xu, Y., Xu, Q., He, Z., & Xia, W. (2018). Occurrence of
benzophenones, parabens and triclosan in the Yangtze River of China, and the
implications
for
human
exposure.
Chemosphere,
213,
517–525.
https://doi.org/10.1016/j.chemosphere.2018.09.084
Maia, C., Sousa, C. A., Sousa, H., Vale, F., & Simoes, M. (2023). Parabens removal
from wastewaters by microalgae–Ecotoxicity, metabolism and pathways. Chemical
Engineering Journal, 453, 139631. https://doi.org/10.1016/j.cej.2022.139631
Matten, S., Fallacara, D., Kamel, A., Lynn, S. G., Fort, D. J., Wolf, J. C., Leak, T., Holst,
H., Bever, R.J., Thomas, A. & Brown, V. J. (2023). Evaluation of multigenerational
effects of 2‐ethylhexyl 4‐hydroxybenzoate in Japanese medaka. Journal of Applied
Toxicology, 43(11), 1645-1666. https://doi.org/10.1002/jat.4502
McCarthy, J. F., & Shugart, L. R. (2018). Biological markers of environmental
contamination. In Biomarkers of environmental contamination (pp. 3-14). CRC Press.
Mejías, C., Martín, J., Santos, J. L., Aparicio, I., & Alonso, E. (2023). Role of polyamide
microplastics as vector of parabens in the environment: An adsorption study.
Environmental
Technology
&
Innovation,
32,
103276.
https://doi.org/10.1016/j.eti.2023.103276
Medkova, D., Hollerova, A., Riesova, B., Blahova, J., Hodkovicova, N., Marsalek, P.,
Doubkova, V., Weiserova, Z., Mares, J., Faldyna, M., Tichy, F., Svobodova, Z. &
Lakdawala, P. (2023). Pesticides and Parabens Contaminating Aquatic Environment:
Acute and Sub-Chronic Toxicity towards Early-Life Stages of Freshwater Fish and
Amphibians. Toxics, 11(4), 333. https://doi.org/10.3390/toxics11040333
Merola, C., Lai, O., Conte, A., Crescenzo, G., Torelli, T., Alloro, M., & Perugini, M.
(2020). Toxicological assessment and developmental abnormalities induced by
butylparaben and ethylparaben exposure in zebrafish early-life stages. Environmental
Toxicology and Pharmacology, 80, 103504. https://doi.org/10.1016/j.etap.2020.103504
Merola, C., Lucon‐Xiccato, T., Bertolucci, C., & Perugini, M. (2021). Behavioural
effects of early‐life exposure to parabens in zebrafish larvae. Journal of Applied
Toxicology. https://doi.org/10.1002/jat.4171
Merola, C., Vremere, A., Fanti, F., Iannetta, A., Caioni, G., Sergi, M., Compagnone, D.,
Lorenzetti, S., Perugini, M., & Amorena, M. (2022). Oxysterols profile in zebrafish
embryos exposed to triclocarban and propylparaben — a preliminary study.
International Journal of Environmental Research and Public Health, 19(3), 1264.
https://doi.org/10.3390/ijerph19031264

93
Meyers, J. R. (2018). Zebrafish: development of a vertebrate model organism. Current
Protocols Essential Laboratory Techniques, 16(1), e19. https://doi.org/10.1002/cpet.19
Mikula, P., Kružíková, K., Dobšíková, R., Haruštiaková, D., & Svobodová, Z. (2009).
Influence of propylparaben on vitellogenesis and sex ratio in juvenile zebrafish (Danio
rerio).
Acta
Veterinaria
Brno,
78(2),
319–326.
https://doi.org/10.2754/avb200978020319
Ministry of Health of the People's Republic of China. National Standards of the
People's Republic of China GB 2760-2015. 20
https://apps.fas.usda.gov/newgainapi/api/report/downloadreportbyfilename?filename=
Standard%20for%20Food%20Additive%20Use%20-%20GB2760-2015_Beijing_China
%20-%20Peoples%20Republic%20of_4-28-2015.pdf
Miranowicz-Dzierżawska, K., Zapór, L., Skowroń, J., Chojnacka-Puchta, L., & Sawicka,
D. (2023). The effects of co-exposure to methylparaben and dibutyl phthalate on cell
line derived from human skin. Toxicological Research, 39(1), 71-89.
https://doi.org/10.1007/s43188-022-00151-3
Murata, W., Yamaguchi, Y., Fujita, K. I., Yamauchi, K., Tanaka, T., & Ogita, A. (2019).
Enhancement of paraben‐fungicidal activity by sulforaphane, a cruciferous
vegetable‐derived isothiocyanate, via membrane structural damage in Saccharomyces
cerevisiae.
Letters
in
applied
microbiology,
69(6),
403-410.
https://doi.org/10.1111/lam.13230
Nowak, K., Ratajczak–Wrona, W., Górska, M., & Jabłońska, E. (2018). Parabens and
their effects on the endocrine system. Molecular and cellular endocrinology, 474,
238-251. https://doi.org/10.1016/j.mce.2018.03.014
Nowak, K., Jabłońska, E., & Ratajczak-Wrona, W. (2021). Controversy around
parabens: Alternative strategies for preservative use in cosmetics and personal care
products.
Environmental
research,
198,
110488.
https://doi.org/10.1016/j.envres.2020.110488
Ortiz de García, S. A., Pinto Pinto, G., García-Encina, P. A., & Irusta-Mata, R. (2014).
Ecotoxicity and environmental risk assessment of pharmaceuticals and personal care
products in aquatic environments and wastewater treatment plants. Ecotoxicology,
23(8), 1517–1533. https://doi.org/10.1007/s10646-014-1293-8
Pedersen, K. L., Pedersen, S. N., Christiansen, L. B., Korsgaard, B., &
BjerregaardNote, P. (2000). The preservatives ethyl-, propyl- and butylparaben are

94
oestrogenic in an in vivo fish assay. Pharmacology and Toxicology, 86(3), 110–113.
https://doi.org/10.1034/j.1600-0773.2000.d01-20.x
Peng, X., Zheng, K., Liu, J., Fan, Y., Tang, C., & Xiong, S. (2018). Body
size-dependent bioaccumulation, tissue distribution, and trophic and maternal transfer
of phenolic endocrine-disrupting contaminants in a freshwater ecosystem.
Environmental
Toxicology
and
Chemistry,
37(7),
1811–1823.
https://doi.org/10.1002/etc.4150
Pereira, A. R., Gomes, I. B., & Simões, M. (2023). Impact of parabens on drinking
water bacteria and their biofilms: The role of exposure time and substrate materials.
Journal
of
Environmental
Management,
332,
117413.
https://doi.org/10.1016/j.jenvman.2023.117413
Perugini, M., Merola, C., Amorena, M., D'Angelo, M., Cimini, A., & Benedetti, E.
(2019). Sublethal exposure to propylparaben leads to lipid metabolism impairment in
zebrafish early‐life stages. Journal of Applied Toxicology, 40(4), 493–503.
https://doi.org/10.1002/jat.3921
Piao, C., Chen, L., & Wang, Y. (2014). A review of the extraction and chromatographic
determination methods for the analysis of parabens. Journal of Chromatography B,
969, 139–148. https://doi.org/10.1016/j.jchromb.2014.08.015
Pico, Y., Belenguer, V., Corcellas, C., Diaz-Cruz, M. S., Eljarrat, E., Farré, M.,
Gago-Ferrero, P., Huerta, B., Navarro-Ortega, A., Petrovic, M., Rodríguez-Mozaz, S.,
Sabater, L., Santín, G., & Barcelo, D. (2019). Contaminants of emerging concern in
freshwater fish from four Spanish Rivers. Science of the Total Environment, 659,
1186–1198. https://doi.org/10.1016/j.scitotenv.2018.12.366
Porto, V. A., da Rocha Júnior, E. R., Ursulino, J. S., Porto, R. S., da Silva, M., de
Jesus, L. W. O., Oliveira, M. J., Crispim, A.C., Santos, J. C. C. & de Aquino, T. M.
(2023). NMR-based metabolomics applied to ecotoxicology with zebrafish (Danio rerio)
as a prominent model for metabolic profiling and biomarker discovery: Overviewing the
most recent approaches. Science of The Total Environment, 868, 161737.
https://doi.org/10.1016/j.scitotenv.2023.161737
Qualhato, G., Vieira, L. G., Oliveira, M., & Rocha, T. L. (2023). Plastic microfibers as a
risk factor for the health of aquatic organisms: A bibliometric and systematic review of
plastic pandemic. Science of The Total Environment, 870, 161949.
https://doi.org/10.1016/j.scitotenv.2023.161949

95
Renz, L., Volz, C., Michanowicz, D., Ferrar, K., Christian, C., Lenzner, D., &
El-Hefnawy, T. (2013). A study of parabens and bisphenol A in surface water and fish
brain tissue from the Greater Pittsburgh Area. Ecotoxicology, 22(4), 632–641.
https://doi.org/10.1007/s10646-013-1054-0
Santos, M. M. d., Brehm, F. d. A., Filippe, T. C., Knapik, H. G., & Azevedo, J. C. R. d.
(2016). Occurrence and risk assessment of parabens and triclosan in surface waters of
southern Brazil: A problem of emerging compounds in an emerging country. RBRH,
21(3), 603–617. https://doi.org/10.1590/2318-0331.011616018
Shore, E. A., Huber, K. E., Garrett, A. D., & Pespeni, M. H. (2022). Four plastic
additives reduce larval growth and survival in the sea urchin Strongylocentrotus
purpuratus.
Marine
Pollution
Bulletin,
175,
113385.
https://doi.org/10.1016/j.marpolbul.2022.113385
Silva, D. C., Serrano, L., Oliveira, T. M. A., Mansano, A. S., Almeida, E. A., & Vieira, E.
M. (2018). Effects of parabens on antioxidant system and oxidative damages in Nile
tilapia (Oreochromis niloticus). Ecotoxicology and Environmental Safety, 162, 85–91.
https://doi.org/10.1016/j.ecoenv.2018.06.076
Silva, L. C., Moreira, R. A., Pinto, T. J., Ogura, A. P., Yoshii, M. P., Lopes, L. F.,
Montagner, C.C., Goulart, B.V., Daam, M.A. & Espindola, E. L. (2020). Acute and
chronic toxicity of 2, 4-D and fipronil formulations (individually and in mixture) to the
Neotropical cladoceran Ceriodaphnia silvestrii. Ecotoxicology, 29, 1462-1475.
https://doi.org/10.1007/s10646-020-02275-4
Sivapragasam, M., Moniruzzaman, M., & Goto, M. (2020). An overview on the
toxicological properties of ionic liquids toward microorganisms. Biotechnology journal,
15(4), 1900073. https://doi.org/10.1002/biot.201900073
Song, C., Lin, J., Huang, X., Wu, Y., Liu, J., & Wu, C. (2015). Effect of butyl paraben on
the development and microbial composition of periphyton. Ecotoxicology, 25(2),
342–349. https://doi.org/10.1007/s10646-015-1592-8
Terasaki, M., Abe, R., Makino, M., & Tatarazako, N. (2013). Chronic toxicity of
parabens and their chlorinated by-products in Ceriodaphnia dubia. Environmental
Toxicology, 30(6), 664–673. https://doi.org/10.1002/tox.21944
Terasaki, M., Makino, M., & Tatarazako, N. (2009). Acute toxicity of parabens and their
chlorinated by-products with Daphnia magna and Vibrio fischeri bioassays. Journal of
Applied Toxicology, 29(3), 242–247. https://doi.org/10.1002/jat.1402

96
Thakkar, S., Seetharaman, B., & Ramasamy, V. (2021). Impact of chronic sub-lethal
methylparaben exposure on cardiac hypoxia and alterations in neuroendocrine factors
in
zebrafish
model.
Molecular
Biology
Reports,
49(1),
331–340.
https://doi.org/10.1007/s11033-021-06878-w
Tomás‐Barberán, F. A., & Clifford, M. N. (2000). Dietary hydroxybenzoic acid
derivatives–nature, occurrence and dietary burden. Journal of the Science of Food and
Agriculture, 80(7), 1024-1032.
https://doi.org/10.1002/(SICI)1097-0010(20000515)80:7%3C1024::AID-JSFA567%3E3
.0.CO;2-S
Torres, T., Cunha, I., Martins, R., & Santos, M. (2016). Screening the toxicity of
selected personal care products using embryo bioassays: 4-MBC, propylparaben and
triclocarban. International Journal of Molecular Sciences, 17(10), 1762.
https://doi.org/10.3390/ijms17101762
Tran, C. M., Ra, J. S., Rhyu, D. Y., & Kim, K. T. (2023). Transcriptome analysis reveals
differences in developmental neurotoxicity mechanism of methyl-, ethyl-, and
propyl-parabens in zebrafish embryos. Ecotoxicology and Environmental Safety, 268,
115704. https://doi.org/10.1016/j.ecoenv.2023.115704
Tran-Lam, T. T., Phung, A. T. T., Pham, P. T., Bui, M. Q., Dao, Y. H., & Le, G. T. (2023).
Occurrence, biomagnification, and risk assessment of parabens and their metabolites
in
marine
fish: The case study of Vietnam. Chemosphere, 344,
140221.https://doi.org/10.1016/j.chemosphere.2023.140221
Vale, F., Sousa, C. A., Sousa, H., Santos, L., & Simões, M. (2022). Parabens as
emerging contaminants: Environmental persistence, current practices and treatment
processes. Journal
of
Cleaner
Production,
131244.
https://doi.org/10.1016/j.jclepro.2022.131244
Wang, L., Asimakopoulos, A. G., & Kannan, K. (2015). Accumulation of 19
environmental phenolic and xenobiotic heterocyclic aromatic compounds in human
adipose
tissue.
Environment
international,
78,
45-50.
https://doi.org/10.1016/j.envint.2015.02.015
Wang, N., Hu, X., Lu, S., Ma, S., Kang, L., Liao, S., & Yu, Y. (2019). Interrelationship of
anthropogenic activity and parabens in fish from Taihu Lake during 2009–2017.
Environmental Pollution, 252, 1002–1009. https://doi.org/10.1016/j.envpol.2019.06.041

97
Wei, F., Mortimer, M., Cheng, H., Sang, N., & Guo, L. H. (2021). Parabens as
chemicals of emerging concern in the environment and humans: A review. Science of
the Total Environment, 778, 146150. https://doi.org/10.1016/j.scitotenv.2021.146150
Wu Y, Sun Q, WangYW, Deng CX, Yu CP (2017) Comparative studies of aerobic and
anaerobic biodegradation of methylparaben and propylparaben in activated sludge.
Ecotoxicology
and
Environmental
Safety,
138:25–31.
https://doi.org/10.1016/j.scitotenv.2011.09.040
Xue, J., & Kannan, K. (2016). Accumulation profiles of parabens and their metabolites
in fish, black bear, and birds, including bald eagles and albatrosses. Environment
International, 94, 546–553. https://doi.org/10.1016/j.envint.2016.06.015
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. Environmental
Science & Technology, 51(2), 780–789. https://doi.org/10.1021/acs.est.6b05501
Yamamoto, H., Tamura, I., Hirata, Y., Kato, J., Kagota, K., Katsuki, S., Yamamoto, A.,
Kagami, Y., & Tatarazako, N. (2011). Aquatic toxicity and ecological risk assessment of
seven parabens: Individual and additive approach. Science of the Total Environment,
410-411, 102–111. https://doi.org/10.1016/j.scitotenv.2011.09.040
Yao, L., Lv, Y.-Z., Zhang, L.-J., Liu, W.-R., Zhao, J.-L., Liu, Y.-S., Zhang, Q.-Q., & Ying,
G.-G. (2018). Determination of 24 personal care products in fish bile using hybrid
solvent precipitation and dispersive solid phase extraction cleanup with ultrahigh
performance liquid chromatography-tandem mass spectrometry and gas
chromatography-mass spectrometry. Journal of Chromatography A, 1551, 29–40.
https://doi.org/10.1016/j.chroma.2018.04.003
Yao, L., Lv, Y.-Z., Zhang, L.-J., Liu, W.-R., Zhao, J.-L., Yang, Y.-Y., Jia, Y.-W., Liu, Y.-S.,
He, L.-Y., & Ying, G.-G. (2019). Bioaccumulation and risks of 24 personal care
products in plasma of wild fish from the Yangtze River, China. Science of the Total
Environment, 665, 810–819. https://doi.org/10.1016/j.scitotenv.2019.02.176
Yin, Y., Xie, Z., Sun, X., Wu, X., Zhang, J., Shi, H., Ding, L. & Hong, M. (2023). Effect
of Butyl Paraben on Oxidative Stress in the Liver of Mauremys sinensis. Toxics, 11(11),
915. https://doi.org/10.3390/toxics11110915
Zhao, X., Qiu, W., Zheng, Y., Xiong, J., Gao, C., & Hu, S. (2019). Occurrence,
distribution, bioaccumulation, and ecological risk of bisphenol analogues, parabens
and their metabolites in the Pearl River Estuary, South China. Ecotoxicology and
Environmental Safety, 180, 43–52. https://doi.org/10.1016/j.ecoenv.2019.04.083

98
Zhang, S., Wang, J., Liu, X., Qu, F., Wang, X., Wang, X., Li, Y. & Sun, Y. (2019).
Microplastics in the environment: A review of analytical methods, distribution, and
biological effects. TrAC Trends in Analytical Chemistry, 111, 62-72.
https://doi.org/10.1016/j.trac.2018.12.002
Zhang, Z. F., Wang, L., Zhang, X., Zhang, X., Li, Y. F., Nikolaev, A., & Li, W. L. (2021).
Fate processes of Parabens, Triclocarban and Triclosan during wastewater treatment:
assessment via field measurements and model simulations. Environmental Science
and
Pollution
Research,
28(36),
50602-50610.
https://doi.org/10.1007/s11356-021-14141-9
Zhu, H., Liao, D., Mehmood, M. A., Huang, Y., Yuan, W., Zheng, J., Ma, Y., Peng, Y.,
Tian, G., Xiao, X., Lan, C., Li, L., Xu, K., Lu, H. & Wang, N. (2023). Systolic heart
failure induced by butylparaben in zebrafish is caused through oxidative stress and
immunosuppression. Ecotoxicology and Environmental Safety, 268, 115692.
https://doi.org/10.1016/j.ecoenv.2023.115692
Zhu, R. G., Pan, C. G., Peng, F. J., Zhou, C. Y., Hu, J. J., & Yu, K. (2024). Parabens
and their metabolite in a marine benthic-dominated food web from the Beibu gulf,
South China Sea: Occurrence, trophic transfer and health risk assessment. Water
Research, 248, 120841. https://doi.org/10.1016/j.watres.2023.120841

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Capítulo 2
INTERACTIVE EFFECTS OF METHYLPARABEN AND MICROPLASTICS IN THE
DEVELOPING ZEBRAFISH (Danio rerio)

Abstract
Among the contaminants of emerging concern (CECs) are parabens, esters of the
p-hydroxybenzoic acid which are commonly used as preservatives in industrial
preparations. Microplastics, in turn, are particles up to 5 mm in diameter made up of
various types of organic polymers, among which polyethylene (PE) is one of the most
widely used and found in aquatic environments. These particles have the ability to
adsorb substances on their surface, potentially modifying the toxicity of the adsorbed
molecules. This study sought to understand the toxicity of the interaction between
methylparaben (MeP) and polyethylene microplastics (MPPE) on zebrafish (Danio rerio)
in its early stages of development. To achieve this, the embryolarval toxicity test (ZELT),
morphometric analysis and behavioral analysis were carried out using the
environmentally relevant concentrations of 0.01, 0.1 and 1 µM of MeP separately or in a
mixture with a suspension of MPPE at a concentration of 3.4 mg/L. There were no
differences between the treatments regarding mortality, hatching rates and spontaneous
movements during the 144 hours of experiment, but there was a significant difference in
the heart rate of groups exposed to the concentrations of parabens in comparison to
negative and solvent controls and to the group exposed to MPPE alone, which is an
evidence that MeP is capable of eliciting cardiotoxicity in zebrafish embryos at
environmentally relevant concentrations. However, the lower concentration of MeP in
mixture with MPPE didn’t elicit cardiotoxicity, which might be a signal that MPPE could,
in certain conditions, reduce the toxicity of contaminants adsorbed to its surface. This
phenomenon indicates not only that MPPE is, in fact, capable of interacting with MeP
and produce different effects than the isolate exposure, but also that the variability of
microplastics as a contaminant suite must be taken into account, especially regarding
size, shape and type of polymer. Further studies addressing such variables should be
carried in order to form a more comprehensive panoram of the interactions between
microplastics and parabens.
Keywords: contaminants of emerging concern, ecotoxicology, parabens, plastic
residues,conservatives, trojan horse effect.

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2.1 Introduction
Contaminants of emerging concern (CECs) are a set of substances and
microparticles of anthropogenic origin that are found ubiquitously in various
environmental matrices due to industrial development and population growth. They
include personal care products (PCPs) and pharmaceuticals, which have a wide range
of chemicals and new formulations continuously introduced to the market (Wilkinson et
al., 2017). Although present in low concentrations in their environments (µg/L or ng/L),
their behavior in the environment and ecotoxicological profile are not well understood.
This information gap raises concerns, especially since they are not typically removed by
wastewater treatment technologies (Gao et al., 2021). Furthermore, their potential
impacts have not been thoroughly explained in terms of their effects on individual,
population and community levels of aquatic ecosystems.
Parabens, which belong in the group of CECs, are a group of low molecular
weight organic substances generally defined as the alkyl and aryl esters of the
4-hydroxybenzoic acid (4-HB). They have been widely used as preservatives in
pharmaceutical preparations since the 1920s and are now also found in cosmetics,
foodstuffs, and other industrial products (Nowak et al., 2018; Bolujoko et al., 2022).
Among them, the most commonly used are methylparaben (MeP) and propylparaben
(PrP) — often found under the trade names of Nipagin™ and Nipasol™, respectively —,
either alone or in combination. They are stable in acidic aqueous solutions and their
properties vary according to the size/molecular weight of their alkyl group: longer chains
have greater antimicrobial activity and resistance to hydrolysis, but lower water solubility
(Błędzka et al, 2014). Parabens, like other PCPs, can be considered pseudopersistent
contaminants due to their continuous influx in low concentrations into surface waters
through anthropogenic effluents (Garric, 2013).
Parabens have been found in various aquatic environments, drinking water,
urban effluents, and agricultural soils as a result of research conducted in different
countries (Feng et al., 2019; Pompei et al., 2019). The presence of parabens in urban

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wastewater is well documented, and despite their low stability, high biodegradability
under aerobic conditions, and removal efficiency of over 90% during water treatment
processes (Haman et al, 2015), they are still found in treated effluents and in water and
sediments of water bodies, as well as their metabolite 4-HB (Feng et al., 2019).
Regarding the concentrations of parabens found in natural and anthropized
environments, there is great variability depending on the location studied. The highest
concentrations are found in urban effluents, reaching 76 900 ng/L (equivalent to ~ 0.5
µM) in a Southern California water treatment plant (Błędzka et al, 2014). In Brazil, MeP
was detected in streams in the city of Rio Grande/RS at concentrations between 7.6
and 29.8 µg/L (equivalent to ~ 0.05 to 0.2 µM), and in the city of Morro Redondo/RS,
between <1 and 134 µg/L (~ 0.88 µM) (Penha et al, 2021). Derisso et al (2020), when
analyzing seven points of the Monjolinho River in the city of São Carlos/SP, detected
MeP concentrations ranging from 0.11 to 0.98 µg/L. In an assessment of three rivers in
the Curitiba/PR region, Santos et al. (2016) found MeP concentrations of up to 2875
ng/L. In the Lobo reservoir, São Carlos/SP, MeP was detected in all water samples in a
mean concentration of 170.87 µg/L (~ 1.12 µM) and a maximum concentration of
1192.39 µg/L (~ 7.85 µM) (Pompei et al., 2019). In addition to their presence in surface
water, there is also evidence of bioaccumulation of parabens in animal tissues. Xue and
Kannan (2016) reported the presence of MeP and its metabolite 4-HB in the kidney,
liver and muscle tissue of bald eagles and albatrosses at concentrations ranging from
580 ng/g for MeP to between 35 and 300 ng/g for 4-HB, depending on the tissue and
species. In the same study, the presence of MeP, PrP and 4-HB was reported in the
liver and brain tissues of fish from the Florida coast, at concentrations ranging from 11.2
ng/g for MeP to 1130 ng/g for 4-HB. Tissue concentrations higher than plasma
concentrations are suggestive of bioaccumulation, especially in the liver (Xue; Kannan,
2016). So far, parabens have been detected in muscle, liver, kidney, fat tissue, muscle,
brain, plasma, gill, ovary and testicle samples of freshwater and marine fish collected
from multiple sites (da Silveira et al., 2024).
From the end of the 20th century, with the publication of studies suggesting
the estrogenic and antiandrogenic activity of parabens, concern grew about their
possible negative effects on the ecological balance (Błędzka et al, 2014). Their potential

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estrogenic activity is associated with the size of their alkyl chain, so that butylparaben
(BuP), heptylparaben (HeP) and benzylparaben (BzP) have a greater potential for
endocrine disruption, according to in silico, in vitro and in vivo approaches (Routledge et
al, 1998; Watanabe et al., 2013; Wei et al., 2022; Liang et al., 2023). 4-HB is considered
the final metabolite of the biotic degradation of parabens and also has estrogenic activity
reported in vitro and in vivo (Błędzka et al, 2014; Raja et al, 2019). The generation of
oxidative stress, embryotoxicity, neurotoxicity and cardiotoxicity are also possible
deleterious effects of these compounds (Merola et al, 2020; Penha et al, 2021; Fan et
al., 2022; Hu et al., 2023). However, studies aimed at clarifying the cellular, metabolic,
physiological and ecological consequences of fish exposure to sub-lethal concentrations
of parabens are still incipient.
Plastic pollution is a pressing environmental issue due to its prevalence as
the primary anthropogenic debris in aquatic environments, even in remote areas without
human presence (Chassignet et al, 2021). Although plastic has practical advantages,
such as malleability, durability, and lightness, its negative impact is becoming
increasingly apparent, since 8 to 11 million tons of plastic are dumped into the oceans
annually (Fava, 2022). The term 'microplastics' (MPs) was first introduced in 2004 by
Thompson et al., referring to plastic particles that are 5 mm or smaller in size. Since
then, researchers have been investigating their presence in surface waters, sediments,
and organisms, as well as their potential harmful effects on aquatic ecosystems
(Rezania et al., 2018). MPs can be of primary origin, such as plastic pellets used as raw
materials by industries or as exfoliating agents in PCPs. They can also be released by
the abrasion of tires or the washing of synthetic fabrics. Alternatively, they can be of
secondary origin, generated from the fragmentation of macroplastics by chemical,
luminous, or biotic degradation phenomena (Rezania et al., 2018; Schmid et al., 2020).
The surface of such polymers has a strong attraction to hydrophobic molecules,
enabling them to adsorb pollutants with this property, such as polycyclic aromatic
hydrocarbons (PAHs) (Hou et al., 2023), polychlorinated biphenyls (PCBs) (Llorca et al.,
2020), antibiotic drugs (Li et al., 2018), dyes (Xia et al., 2020) and steroid hormones (Lu
et al., 2020; Lara et al., 2021). Due to their small size, aquatic organisms can easily
ingest them, making them a potential carrier of highly toxic pollutants and pathogenic

103
microorganisms, acting as a 'Trojan Horse' (Trevisan et al., 2020; Kinigopoulou et al.,
2022; Rafa et al., 2024).
According

to

Plastics

Europe

(2020), low-density and high-density

polyethylene are the thermoplastic polymers with the second and third largest market
demand in Europe, respectively, after polypropylene. Majewski et al. (2016) analyzed
two wastewater samples from the city of Karlsruhe in Germany and reported that
polyethylene was the most common polymer, accounting for 34% (81 mg/m³) and 17%
(257 mg/m³) of each sample. Yong et al. (2020) lists the negative consequences of MPs
and NPs on embryos, larvae, and adult individuals of various fish species, including
zebrafish, medaka (Oryzias spp.), Japanese medaka (Tigriopus japonicas), carp
(Cyprinus carpio), sea bream (Carassius auratus), Pimephales promelas, and wild
species. Such adverse effects include inflammatory processes, damage to the digestive
system, reduced growth, oxidative stress, behavioral changes, and reduced viability and
survival of embryos and larvae. In addition, MPs can also be adsorbed by primary
producers in aquatic environments, such as algae and plankton, affecting their
photosynthesis and respiration rate, thus inhibiting their growth (Saud et al., 2023). The
bioaccumulation of multiple toxicants carried by MPs is also a growing concern,
especially concerning mollusks, which are filter feeders and more propense to
bioaccumulate substances (Rafa et al., 2024).
Considering such information, understanding how MPs and other organic
substances present in aquatic environments interact and exert toxic effects on the local
biota is an important matter that must be further researched. In this way, the present
study sought to understand the toxicity of the interaction between methylparaben (MeP)
and polyethylene microplastics (MPPE) on zebrafish (Danio rerio) in its early stages of
development. The expected outcomes are a higher toxicity of mixtures in comparison to
the isolated contaminants, as well as a concentration-dependent response to the
exposure to MeP.

104
2.2 Materials and Methods
2.2.1 Reagents
Polyethylene microplastics (MPPE) purchased from Sigma Aldrich were
used. They had been previously characterized by Dias (2023) through electron
microscopy, presenting a diameter of 35.46 µM ± 18.17 µM, heterogeneous shape and
irregular surface. The MPPE concentrations used in the treatments were prepared from
a stock solution with a concentration of 34 mg/L (10x). Polyethylene was chosen as
representative of MPs because it is one of the most widely used polymers in various
industrial applications, as well as one of the most found in the aquatic environment and
with the greatest resistance to biodegradation (Horton et al, 2017; Sequeira et al, 2020).
The other reagents such as methylparaben (MeP), dimethylsulfoxide (DMSO),
dichloroaniline (3,4-dichloroaniline) were also obtained from Sigma, with a high degree
of purity.
2.2.2 Maintenance of adult zebrafish and collection of eggs
The adult zebrafish of AB strain were kept in the Instituto de Patologia
Tropical e Saúde Pública of the Universidade Federal de Goiás (IPTSP — UFG) Fish
Nursery, in 3 L injected polycarbonate tanks (SLZF 110 - Scienlabor), at a ratio of 5
animals/L, at 27 ºC in a recirculating water system, on a 14:10h light-dark cycle, as
recommended by Reed and Jennings (2010). The tanks were filled with reconstituted
water (deionized water, sodium bicarbonate, magnesium sulphate, calcium chloride and
potassium chloride) and cleaned regularly. The fish were fed twice a day with flake feed
and once a day with brine shrimp. For breeding, the animals were placed in multiple
brooders (Tecniplast) with a ratio of 1:1 between males and females in each. After
breeding early the next morning, the embryos were collected and the viable embryos
separated. Different groups of animals were used for breeding in order to reduce the
genetic influence on the parameters evaluated. All the experiments were carried out
under the approval of UFAL's Ethics Committee on the Use of Animals — nº 15/2022.
2.2.3 Exposure

105
The newly fertilized zebrafish embryos were placed in 24-well plates (Kasvi),
one embryo per well, in a total of 10 embryos for each treatment below: I) Negative
control (reconstituted water only — NC); II) Solvent control (0.05% DMSO) — SC; III)
Positive control (3,4-DCA at 4 mg/L — which promotes 30 to 100% embryo mortality) —
PC; IV) Polyethylene plastic microparticles (MPPE - 3.4 mg/L); V) MeP (0.01 µM); VI)
MeP (0.01 µM) + MPPE (3.4 mg/L); VII) MeP (0.1 µM); VIII) MeP (0.1 µM ) + MPPE (3.4
mg/L); IX) MeP (1 µM ); X) MeP (1 µM) + MPPE (3.4 mg/L).
​ The plates were incubated in an embryo chamber (Scienlabor) and kept at
the same temperature and photoperiod as the adults. Exposure was semi-static for 144
hours and the solutions were changed every 24 hours. The concentrations of 0.01 and
0.1 µM of MeP have environmental relevance (Penha et al, 2021), while the
concentration of 1 µM (equivalent to 152 µg/L) represents the highest average
concentrations found in current literature for surface waters (Bolujoko et al., 2022). The
concentration of 3.4 mg/L of MPPE also has environmental relevance (Koelmans et al.,
2019) and might elicit mild toxicity throughout the embryolarval period of zebrafish
(Malafaia et al., 2020).
2.2.4 Zebrafish embryo-larval toxicity test (ZELT)
The test was adapted from OECD standards (2013), assessing lethal,
non-lethal and teratogenic parameters. All tests were carried out in triplicate, using
batches of embryos obtained from different groups of adults. Throughout the exposure
period, the embryos/larvae were analyzed every 24 hours using a (ZEISSⓇ Stemi 508)
with an associated image capture system (ZEISSⓇ Axiocam 105 color). Embryo
mortality was assessed daily and considered based on four possible results, according
to OECD standards (2013): a) Embryo coagulation; b) Absence of somites; c)
Non-detachment of the tail; d) Absence of a heartbeat.
The non-lethal parameters analyzed were: Hatching rate (48, 72, 96 hours
post-fertilization — hpf); Spontaneous movements/min (24 hpf), related to neurotoxicity;
Heartbeats/min (48 hpf), related to cardiotoxicity; and embryo pigmentation (24, 48, 72,
96, 120 and 144 hpf). The teratogenic effects verified were: Presence of scoliosis (24 to
144 hpf); Vitelinic deformation (24 to 144 hpf); Growth retardation in general (24 to 144

106
hpf); Eye, otolith and tail defects (24 to 144 hpf); and pericardial and vitelinic edema (24
to 144 hpf).
2.2.5 Behavioral analysis
The methodology for behavioral analysis was adapted from Pinheiro-Da-Silva
et al. (2020). After 144 hours of exposure to MPPEs with or without the MeP
concentrations listed in section 2.3, 15 larvae (5 larvae from each replicate) were
transferred to 12-well microplates (KASVI®) with 3 mL of reconstituted water per well,
resulting in 15 larvae per experimental group, except for the positive control. The larvae
were acclimatized to the recording room temperature of 26 ºC for 30 minutes before the
1-minute recordings. The recordings were conducted in a mini-studio (24.5 cm x 24.5
cm x 24.5 cm) using a Logitech C922 Pro® webcam mounted on top of a Puluz Photo
Light Box®. The videos were analyzed using ZebTrack software, developed by
Pinheiro-da-Silva et al. (2016) and implemented in MATLAB (R2014a; MathWorks,
Natick, MA). The locomotor behavioral parameters assessed were total distance
traveled (DT), mean speed (MS), maximum speed (Vmax), and peripheral time (PTime).
2.2.6 Morphological analysis
After the ZELT and the recordings for behavioral analysis, the specimens
were euthanized by immersion in a 0.1% benzocaine solution and fixed in 4%
paraformaldehyde for 24 hours. After fixation, the specimens were washed thrice in 0.2
M PBS buffer at pH 7.2 and kept in 70% alcohol at 4 ºC until biometry was carried out.
Pictures from a lateral and dorsal view of each randomly selected individual (n = 15 per
treatment) were taken using a stereomicroscope (ZEISSⓇ Stemi 508) with an
associated image capture system (ZEISSⓇ Axiocam 105 color). The images were
analyzed using ImageJ software and morphometric parameters were divided into four
categories: i. sensory (eye diameter; maximum and minimum distances between the
eyes); ii. physiological (swimming bladder, yolk sac and pericardial sac diameters); iii.
skeletal structural (height, head width and depth, and distances from mouth to anus); iv.
muscle structural (angle and distance between myotomes), according to Malafaia et al

107
(2020) and Ribeiro et al (2020).
4.2.7 Statistical analysis
The fish parameters measured were compared between treatments through
Generalized Linear Model Analysis, using the best-fitted model. For some of these data,
the best-fit model was attained to Gaussian distribution. For the hatching rate, the
best-fitted model was attained to quasibinomial distribution. The predictive factors used
were treatments and, when appropriate, time. Pairwise comparisons were performed
using a posteriori Tukey tests. To verify the mortality rates in the different treatments
over time, a Survival Analysis was performed using the Kaplan-Meier curve.
A significance level of 5% was used. All analyses were performed in the R
environment (R core Team, 2024) using the following packages: multcomp (Hothorn et
al., 2018), survival (Therneau, 2024), ggsurvfit (Sjoberg et al., 2024), flexsurv (Jackson,
2016) and survminer (Kassambara et al., 2021).
2.3 Results and Discussion
Regarding the survival of the embryos (Fig. 1A), both the negative control
group (NC) and the solvent control group (SC) had a survival rate of 96,7% in the first
48 h, while the survival of the positive control group (PC) was significantly lower (p <
0.0001). At the 48 hpf time stamp, all the embryos from the PC group were dead.
Additionally, there were no differences between the survival rate of any treatment group
and the NC and SC groups. This result was expected, since previous studies have
established high LC50 values for MeP in zebrafish — ranging from 72.67 mg/L (Merola
et al., 2020a) to 211.12 mg/L for 96h larvae (Penha et al., 2021), which are
approximately 470 to 1,380 times higher than the highest concentration used in this
study. Furthermore, MeP is generally regarded as having lower toxicity in comparison to
other parabens due to its shorter side alkyl chain (Liang et al., 2023). In a study
conducted with zebrafish larvae (up to 24 hours post hatching — hph) and adults
exposed to an environmentally relevant concentration of MeP (30 µg/L, equivalent to ~
0.2 µM) for 168 h and 96 h, respectively, survival rate also wasn’t altered in comparison
to controls (Penha et al., 2021). Moreover, lower survival rate wasn’t observed in

108
zebrafish embryos exposed to MeP at the concentration of 1 mg/L (Merola et al.,
2020a), which is near the highest environmental findings for MeP (1192.39 µg/L)
(Pompei et al., 2019).
In accordance with our findings, the exposure of zebrafish embryos to
concentrations of MPPE (38.26 ± 15.64 µM) that ranged from 6.2 to 100 mg/L did not
alter the survival rates when a semi-static protocol (exposure solution changed every 24
h) was used (Malafaia et al., 2020). Contrasting with these results, a lower 24 h survival
rate was observed in zebrafish embryos exposed to MPPE (52 to 74 µM) in
concentrations that ranged from 102 to 106 particles/L (converted from ~ 0.04 to 437.47
mg/L) using a semi-static protocol (Chen et al., 2023). This result is explained by the
author as resulting from the low density of MPPE, which due to buoyancy form a kind of
hydrophobic film on the water surface and can potentially undermine gas exchange with
the atmosphere. Despite the similarities in concentrations, sizes, polymer and exposure
protocol, the MPPE used in both studies were different in a potentially relevant aspect:
in the study conducted by Malafaia et al. (2020), the MPPE had irregular shapes and a
rough surface, similarly to the ones used in this study; the ones used by Chen et al.
(2023), though, were nearly spherical and had a smoother surface. Further studies
should be conducted in order to investigate factors such as shape, surface texture, color
and aging on the toxicity of MPPE to aquatic organisms in early development stages,
given that their influence on harmful biological outcomes is not as well elucidated as
other aspects such as concentration, size and presence of sorbed contaminants.
In zebrafish, hatching occurs at 48 to 72 hpf, and individuals that have
spontaneously hatched generally are not more developmentally advanced than the
ones that remain in their chorions (Kimmel et al., 1995). The relative plasticity in
hatching time is thought to enhance fitness of zebrafish embryos by balancing the
benefits and costs of emerging as a free-swimming larva, as opposed to remaining
bound within the chorion (Wisenden et al., 2022). Hatching time can be influenced by
various environmental variables, such as temperature, oxygen availability, chemical
signals and the presence of environmental hazards, such as predators and pollutants,
being mainly driven by metabolic rate (Silva et al., 2022). The mechanisms that lead to
early hatching of zebrafish embryos exposed to environmental contaminants aren’t well

109
elucidated. However, it is speculated that MP adherence to the chorionic membrane
might lead to mechanical damage and to physiological changes, such as hindering of
gas exchange and consequent reduction of oxygen supply (Malafaia et al., 2020).
The hatching of eggs started at 48 hpf and ended at 96 hpf for all treatments
(Fig. 1B), except for the group exposed to MPPE alone, in which the start of hatching
was observed only at 72 hpf. In a study conducted by Merola et al. (2020a), the
exposure of zebrafish embryos to a higher, but still environmentally relevant
concentration of MeP (1 mg/L, ~ 6.5 µM) did not lead to alterations in hatching when
compared to controls. In a study conducted by La Pietra et al. (2024), the exposure to
MPPS (1 and 3 µM, 0.01 to 10 mg/L) also did not alter survival or the normal hatching
process in zebrafish embryos in any of the concentrations tested. In a study conducted
by Malafaia et al (2020), the exposure to MPPE at concentrations from 6.2 to 100 mg/L
in a static protocol resulted in early hatching and in lower survival rates in the
concentrations of 25, 50 and 100 mg/L, which was not observed in the semi-static
exposure system for any of the concentrations tested. This suggests that the exposure
protocol might have an equal or greater impact on detrimental development outcomes in
comparison to the concentrations of MPPE used and is in consonance with our findings.

110

Figure 1. A) Survival rate of zebrafish embryos and larvae exposed to methylparaben (MeP) in
environmentally relevant concentrations either alone or in mixture with polyethylene
microplastics (MPPE) during the course of the 144h. B) Hatching rate of zebrafish embryos
exposed to MeP with or without MPPE during the course of 144h.

In the first 24 hours of development of zebrafish embryos, an analysis of
spontaneous movements was conducted, being related to neurological development. At
that point, embryos exhibit slow and rhythmic movements driven by glutamatergic

111
signaling and tonic glycinergic impulses that cause contraction of muscle fibers
throughout the body, along with the development of the brain into five distinct lobes
(Norton, 2012; Mrinalini et al., 2023). There were no significant differences in
spontaneous movements between any treatment groups except for the positive control
(p < 0.01) (Fig. 2A), suggesting lack of neurotoxicity of both MPPE and MeP in
environmentally relevant concentrations as well as their mixtures. Dambal et al (2017)
and Merola et al (2020), in contrast, observed slowed or absent spontaneous
movements in zebrafish embryos exposed to MeP alone, though in much higher
concentrations (100 to 1000 µM and 1 to 80 mg/L — equivalent to 6.5 and 526 µM,
respectively). This result also contrasts with data obtained from adult zebrafish exposed
to environmentally relevant concentrations (1, 3 or 10 µg/L) of MeP for 28 days, in which
brain proteome disruption, oxidative stress, reduction of the brain-somatic index (BSI)
and neural signaling disturbances were observed, along with brain inflammation in the
male fish (Hu et al., 2023). The longer duration of exposure to MeP along with the use
of molecular biomarkers, which are generally more sensitive than those of higher
biological organization levels and can provide earlier signs of harmful effects (Ryan;
Hightower, 1996), may account for such differences. Regarding previous findings for
MPPE alone, Malafaia et al. (2020) also did not observe alteration of the number of
spontaneous movements per minute in 24 hpf zebrafish embryos exposed to
concentrations of MPPE that ranged from 6.2 to 100 mg/L.

112

Figure 2. A) Number of spontaneous movements per minute of 24 hpf zebrafish embryos
exposed to environmentally relevant concentrations of methylparaben (MeP) with or without
the presence of polyethylene microplastics (MPPE) (p < 0.01). B) Number of heartbeats per
minute of 48 hpf zebrafish embryos exposed to MeP and MPPE either alone or in mixture (p <
0.05).

After 48 hours of development, an analysis of embryonic heart rate was
conducted. In zebrafish, the heart development starts at 48 hpf, in which the peristaltic
linear heart tube is formed, and after another 48 hours a separate atrium and ventricle

113
are formed by looping (Teranikar et al., 2023). Malformations during that time period,
which can be induced by environmental pollutants, may affect the hemodynamic
performance of the embryonic heart, which comprises the heart rate and stroke volume
as components of total cardiac output (Teranikar et al., 2023). All the treatments with
MeP alone had a significantly higher heart rate when compared to both NC and SC
groups and also to the group exposed to MPPE alone (p < 0.05) (Fig. 2B). The two
mixture groups with higher concentrations of MeP (0.1 and 1 µM + MPPE) also had a
higher heart rate in comparison to the group exposed to MPPE alone (p < 0.001) (Fig.
2B). This suggests that environmentally relevant concentrations of MeP are capable of
eliciting toxicity effects over zebrafish heart function, even though there wasn’t a
difference in heartbeats per minute between MeP concentrations tested. This result
differs from previous findings in the sense that, although also altering heart function in
48 hpf zebrafish embryos, the exposure to MeP was observed to induce bradycardia
(Dambal et al., 2017; Merola et al., 2020a), while in our findings it appears to induce
tachycardia in comparison to controls. This discrepancy may be attributed to the fact
that the concentrations employed in the present study are 6.5 to 100 times lower than
those utilized in the aforementioned works. Consequently, it is plausible that a distinct
mechanism of cardiotoxicity may be involved depending on the MeP concentrations. In
the study conducted by Malafaia et al (2019), the exposure of zebrafish embryos to
MPPE in various concentrations did not elicit cardiotoxicity in neither a static or
semi-static exposure protocol, which is in accordance with our findings, since the group
exposed to MPPE alone did not differ from the control groups regarding heart rate.
On the other hand, the group exposed to 0.01 µM MeP alone had a
significantly lower heart rate in comparison to the group exposed to 0.01 µM MeP +
MPPE. This could suggest that the association of MPPE with lower concentrations of
MeP might reduce cardiotoxic effects of the latter in zebrafish embryos. In fact,
contrasting results regarding the effects of microplastics on the toxicity of associated
contaminants have been described, with size and shape of particles being relevant
variables to the biological outcomes. In juveniles of the marine fish Dicentrarchus
labrax, it was observed a reduction in immunotoxicological effects when animals were
fed with PFOS (4.83 µg/kg) adsorbed to MPPE (150 to 500 µM, 100 mg/kg) in

114
comparison to those fed with PFOS alone (Espinosa-Ruiz et al., 2023). In larvae of the
aquatic midge Chironomus riparius, MPPE either reduced or increased the toxicity of
the bioinsecticide Bacillus thuringiensis israelensis (Bti) depending on particle size
(Khan & Johnson, 2024). Our findings are also in consonance with a study conducted in
zebrafish embryos regarding the toxicity of mercury (Hg, 0.1 mg/L) alone or in
combination with NPPS (100 nM, 10 mg/L) or MPPS (157 µM, 10 mg/L). While no
difference in heart rate was observed in embryos exposed to MPPS either alone or in
combination with Hg in comparison to controls, exposure to Hg alone was able to
significantly decrease the heart rate of 48 hpf embryos. These results are also attributed
to the size of particles, which might be contained by the chorionic barrier, thus
decreasing the bioavailability of adsorbed Hg to zebrafish embryos (Wang et al., 2022).
The appearance of morphological malformations wasn’t observed in any of
the treatment groups (Fig. 3). Merola et al (2020a) observed multiple lethal and
sublethal alterations in zebrafish embryos and larvae exposed to MeP alone, such as
pericardial and yolk edema and notochord curvature. However, the concentrations in
which such alterations were observed in the aforementioned study (10 to 80 mg/L) were
from 65 to 526 times higher than the highest concentration used in our study and do not
have environmental relevance (Bolujoko et al., 2021). Malafaia et al. (2020) observed a
concentration-dependent response regarding malformations in embryos and larvae of
zebrafish exposed to MPPE in a semi-static manner, with stronger teratogenic effects in
concentrations ranging from 50 to 100 mg/L. The malformations found included
pericardial and vitelline sac edema, spinal curvature and caudal flexure, possibly being
related to chorionic pore clogging and consequent hypoxia induced by the MPPE. In the
study, MPPE were observed to adhere to the external surface and gastrointestinal tract
of the larvae, which might be associated with the toxicity mechanism regarding
teratogenesis. Such findings, nonetheless, contrast with our study, in which
malformations were not observed in the group exposed to MPPE alone. These
differences might be related to the lower concentration used in the present study, which
is in accordance with a concentration-dependent pattern of response.
Microplastics dispersed in the water might be ingested by aquatic organisms,
being detected in their gastrointestinal tract (Malafaia et al., 2019) and causing physical

115
damage and inflammation to the digestive organs (Rafa et al., 2024). The lack of
measurable alterations in the group exposed to MPPE alone, apart from the low
concentration used in the present study, might also be related to the fact that zebrafish
larvae only open their mouths widely after 72 hours of development (Kimmel et al.,
1995), thus allowing ingestion of particles present in the water, but reducing the
effective duration of exposure to MPPE. It is likely that smaller particles could cross the
chorion more promptly during the pre-hatching stages, bypassing the limitation inflicted
by the lack of active ingestion during, at least, the first 48 hours of zebrafish
development (Chen et al., 2024).

116

Figure 3. Photographic sheet of zebrafish embryos and larvae from five different treatment
groups (NC; SC; MPPE; MeP 1 µM; and MPPE + MeP 1 µM) at 24, 48 and 144 hpf. Scale bar
represents 1000 µM for 24 and 48 hpf embryos and 4000 µM for 144 hpf larvae.

117
As a matter of fact, the chorion of unfertilized zebrafish embryos have evenly
spaced pores with approximately 0.2 µM, while the chorionic pores of fertilized embryos
in the gastrula stage (5 to 10 hpf) have a diameter that varies between 0.5 to 0.7 µM
(Pelka et al., 2017) — comparatively, the MPPE used in this study were roughly 70
times larges (~ 35 µM). Whether a molecule is able to cross the chorion seems to
depend on its physical and chemical properties, as well as on its size and molecular
weight. Besides, the stability and permeability of the chorion appears to vary with age,
with blastula and gastrula-stage embryos requiring more force to puncture the chorion
than pre-hatching embryos, which is due to the increase of proteolytic activity with time
(Kim et al., 2024); On the other hand, the chorion of embryos in later stages of
development (> 24 hpf) seem to be more permeable to small molecules such as DMSO
(Kais et al., 2013). A dechorionation process of zebrafish embryos has been proposed
as a tool to enhance the sensitivity of ZELT, avoiding false-negatives due to the chorion
functioning as a barrier to certain materials and molecules, such as some nanomaterials
and bulkier polymeric structures (Pelka et al., 2017; Pereira et al., 2023). This process
could be performed after 24 hpf by a mechanical process with > 90% survival and < 5%
rate of sublethal effects, having shown to significantly increase the sensitivity of
embryos to the deleterial biological effects of a polymeric substance of high molecular
weight (Luviquat HM 552, ~ 400 kDa) (Henn & Braunbeck, 2011). Hence,
dechorionation could be performed in future studies that seek to evaluate the toxicity of
MPs and NPs in zebrafish embryos, aiming to analyze whether this measure has
significant impacts on test sensitivity and biological outcomes.
In contrast to our findings, a study made with smaller PS microplastics (1 and
3 µM in size) in different concentrations (0.01, 1 and 10 mg/L) was able to observe
malformations, tachycardia and apoptotic processes related to oxidative stress in
zebrafish embryos (La Pietra et al., 2024). A work conducted by Sun et al. (2021)
showed

significant

cardiovascular

dysfunction,

oxidative

stress

and systemic

inflammation elicited by NPPE of approximately 191 nm in diameter size; in comparison,
the plastic particles used by Malafaia et al. (2020) and in this study were nearly two
hundred times larger (~ 38 µM). In fact, particle size seems to have a relevant influence
on the type and intensity of toxicity elicited by micro and nanoplastics. In a study

118
conducted by Chen et al (2024) with different sizes of NPs in embryonic and juvenile
zebrafish, only the larger particles (500 nM) elicited oxidative stress, whereas only the
smaller particles (80 nM) were capable of increasing the expression of neural and
optical-specific mRNAs. The exposure of the clam Scrobicularia plana to MPPE (1
mg/L) of two different sizes (4-6 and 20-25 µM) alone or with PFOS adsorbed (55.7 ±
5.3 and 46.1 ± 2.9 μg/g) led to increased accumulation of larger MPPE in whole soft
tissues in comparison to the smaller MPPE. Furthermore, higher levels of lipid
peroxidation in the gills were detected for the larger MPPE, in both isolated form and
with PFOS absorbed, in comparison to the smaller MPPE (Islam et al., 2021). This
suggests that the size of microplastic particles has a significant influence on its toxicity
mechanisms and biological effects. To further investigate this phenomenon, it might be
relevant to carry a study with MPs and parabens using different particle sizes, rather
than, or in addition to, different concentrations of either contaminants. Further
investigation on the dynamics of MP diffusion, absorption and ingestion by embryos and
larvae could also be conducted by using particles dyed with specific stains, such as Nile
red, along with fluorescence microscopy (Bhagat et al., 2020; Malafaia et al., 2020;
Konings et al., 2024).
2.4 Conclusion
Based on the embryotoxicity data alone, the hypothesis that the mixture of
MeP and MPPE would cause greater toxicity than the exposure to either MeP or MPPE
alone could not be confirmed. Alterations in neurological, cardiac and general
development parameters, as well as in survival and hatching rates, were not observed
in the group exposed to MPPE alone when compared to negative and solvent controls.
However, an increased heart rate was observed for almost all the groups exposed to
MeP, with or without the presence of MPPE, except for the group exposed to 0.01 µM
MeP + MPPE. This suggests that MPPE in the size and concentration used in this study
have low toxicity to zebrafish embryos and larvae until 144 hpf and might even reduce
the cardiotoxicity of low MeP concentrations in such individuals. With the conclusion of
further analysis, which will comprise morphometric measurements and locomotor
behavioral assessment, it will be possible to form a better understanding of potential

119
outcomes of the exposure to these pollutants and how they might interact to elicit
toxicity in the earlier stages of zebrafish development.

REFERENCES
ARAÚJO, Amanda Pereira da Costa; LUZ, Thiarlen Marinho da; ROCHA, Thiago
Lopes; AHMED, Mohamed Ahmed Ibrahim; SILVA, Daniela de Melo; RAHMAN, Md
Mostafizur; MALAFAIA, Guilherme. Toxicity evaluation of the combination of emerging
pollutants with polyethylene microplastics in zebrafish: perspective study of genotoxicity,
mutagenicity, and redox unbalance. Journal Of Hazardous Materials, v. 432, p.
128691, jun. 2022. http://dx.doi.org/10.1016/j.jhazmat.2022.128691.
ATEŞ, Perihan Seda; ÜNAL, İsmail; ÜSTÜNDAĞ, Ünsal Veli; ALTURFAN, Ahmet Ata;
YIĞITBAŞĐ, Türkan; EMEKLI-ALTURFAN, Ebru. Methylparaben induces malformations
and alterations on apoptosis, oxidant-antioxidant status, ccnd1 and myca expressions in
zebrafish embryos. Journal Of Biochemical And Molecular Toxicology, v. 32, n. 3, p.
1-6, 23 jan. 2018. http://dx.doi.org/10.1002/jbt.22036.
BHAGAT, Jacky; ZANG, Liqing; NISHIMURA, Norihiro; SHIMADA, Yasuhito. Zebrafish:
an emerging model to study microplastic and nanoplastic toxicity. Science Of The Total
Environment,
v.
728,
p.
138707,
ago.
2020.
http://dx.doi.org/10.1016/j.scitotenv.2020.138707.
BŁĘDZKA, Dorota; GROMADZIńSKA, Jolanta; WĄSOWICZ, Wojciech. Parabens. From
environmental studies to human health. Environment International, v. 67, p. 27-42,
jun. 2014. http://dx.doi.org/10.1016/j.envint.2014.02.007.
BOLUJOKO, Nathaniel B.; UNUABONAH, Emmanuel I.; ALFRED, Moses O.;
OGUNLAJA, Aemere; OGUNLAJA, Olumuyiwa O.; OMOROGIE, Martins O.;
OLUKANNI, Olumide D.. Toxicity and removal of parabens from water: a critical review.
Science Of The Total Environment, v. 792, p. 148092, out. 2021.
http://dx.doi.org/10.1016/j.scitotenv.2021.148092.
BOUCHER, Julien; FRIOT, Damien. Primary Microplastics in the Oceans: a Global
Examination of Sources. International Union for Conservation of Nature (IUCN), Gland,
Suíça, 43 pp, 2017. http://dx.doi.org/10.2305/IUCN.CH.2017.01.en.
SILVA, Bruno William; LEITE-FERREIRA, Maria Elisa; MENEZES, Fabiano Peres;

120
LUCHIARI, Ana Carolina. Covariation among behavioral traits and hatching time in
zebrafish.
Behavioural
Processes,
v.
194,
p.
104546,
jan.
2022.
http://dx.doi.org/10.1016/j.beproc.2021.104546.
BUSCH, Wibke; DUIS, Karen; FENSKE, Martina; MAACK, Gerd; LEGLER, Juliette;
PADILLA, Stephanie; STRÄHLE, Uwe; WITTERS, Hilda; SCHOLZ, Stefan. The
zebrafish embryo model in toxicology and teratology, September 2–3, 2010, Karlsruhe,
Germany. Reproductive Toxicology, v. 31, n. 4, p. 585-588, maio 2011.
http://dx.doi.org/10.1016/j.reprotox.2011.02.010.
CHASSIGNET, Eric P.; XU, Xiaobiao; ZAVALA-ROMERO, Olmo. Tracking Marine Litter
With a Global Ocean Model: where does it go? Where does it come from?. Frontiers In
Marine Science, v. 8, p. 1-15, 23 abr. 2021. Frontiers Media SA.
http://dx.doi.org/10.3389/fmars.2021.667591.
CHEN, Jiangfei; LIANG, Qiuju; ZHENG, Yi; LEI, Yuhang; GAN, Xiufeng; MEI, He; BAI,
Chenglian; WANG, Haiyan; JU, Jingjuan; DONG, Qiaoxiang. Polystyrene nanoplastics
induced size-dependent developmental and neurobehavioral toxicities in embryonic and
juvenile zebrafish. Aquatic Toxicology, v. 267, p. 106842, fev. 2024.
http://dx.doi.org/10.1016/j.aquatox.2024.106842.
CHIESA, Luca Maria; PAVLOVIC, Radmila; PANSERI, Sara; ARIOLI, Francesco.
Evaluation of parabens and their metabolites in fish and fish products: a comprehensive
analytical approach using lc-hrms. Food Additives & Contaminants: Part A, v. 35, n.
12,
p.
2400-2413,
26
nov.
2018.
Informa
UK
Limited.
http://dx.doi.org/10.1080/19440049.2018.1544721.
CLASEN, Barbara; LORO, Vania L.; MURUSSI, Camila R.; TIECHER, Tadeu Luis;
MORAES, Bibiana; ZANELLA, Renato. Bioaccumulation and oxidative stress caused by
pesticides in Cyprinus carpio reared in a rice-fish system. Science Of The Total
Environment,
v.
626,
p.
737-743,
jun.
2018.
http://dx.doi.org/10.1016/j.scitotenv.2018.01.154.
DAMBAL, Vrinda Yatin; SELVAN, Karthikeyan Ponnien; LITE, Christy; BARATHI, S.;
SANTOSH, Winkins. Developmental toxicity and induction of vitellogenin in
embryo-larval stages of zebrafish (Danio rerio) exposed to methylparaben.
Ecotoxicology And Environmental SaZELTy, v. 141, p. 113-118, jul. 2017.
http://dx.doi.org/10.1016/j.ecoenv.2017.02.048.
DERISSO, Carolina Resende; POMPEI, Caroline Moço Erba; SPADOTO, Mariângela;

121
PINTO, Tiago da Silva; VIEIRA, Eny Maria. Occurrence of Parabens in Surface Water,
Wastewater Treatment Plant in Southeast of Brazil and Assessment of Their
Environmental Risk. Water, Air, & Soil Pollution, v. 231, n. 9, p. 1-13, 27 ago. 2020.
Springer
Science
and
Business
Media
LLC.
http://dx.doi.org/10.1007/s11270-020-04835-0.
DIAS, Felipe Cirqueira. Efeitos de concentrações ambientalmente relevantes de
microplásticos de polietileno no sistema sensorial e comportamento do zebrafish (Danio
rerio). 2024. Thesis (Masters in Animal Biodiversity) - Universidade Federal de Goiás,
Coordenação de Aperfeiçoamento de Pessoal de Nível Superior. Advising professor:
Thiago Lopes Rocha.
ESPINOSA-RUIZ, Cristóbal; GONZÁLEZ-FERNÁNDEZ, Carmen; CORMIER, Bettie;
KEITER, Steffen H.; VIEIRA, Luis R.; GUILHERMINO, Lúcia; CLÉRANDEAU,
Christelle; CACHOT, Jérôme; ESTEBAN, María A.; CUESTA, Alberto.
Immunotoxicological effects of perfluorooctanesulfonic acid on European seabass are
reduced by polyethylene microplastics. Fish & Shellfish Immunology, v. 137, p.
108793, jun. 2023. http://dx.doi.org/10.1016/j.fsi.2023.108793.
FAN, Zunpan; YANG, Yunyi; HU, Peixuan; HUANG, Yaochen; HE, Liting; HU, Rui;
ZHAO, Kai; ZHANG, Huiping; LIU, Chunyan. Molecular mechanism of ethylparaben on
zebrafish embryo cardiotoxicity based on transcriptome analyses. Science Of The
Total
Environment,
v.
842,
p.
156785,
out.
2022.
http://dx.doi.org/10.1016/j.scitotenv.2022.156785.
MARTA FAVA. UNESCO — Ocean Literacy Portal. Ocean plastic pollution, an overview:
data
and
statistics.
2022.
Disponível
em:
https://oceanliteracy.unesco.org/plastic-pollution-ocean/. Acesso em: 06 nov. 2024.
QIANG, Liyuan; CHENG, Jinping. Exposure to polystyrene microplastics impairs
gonads of zebrafish (Danio rerio). Chemosphere, v. 263, p. 128161, jan. 2021.
http://dx.doi.org/10.1016/j.chemosphere.2020.128161.
FENG, Jinglan; ZHAO, Jiahui; XI, Nannan; GUO, Wei; SUN, Jianhui. Parabens and
their metabolite in surface water and sediment from the Yellow River and the Huai River
in Henan Province: spatial distribution, seasonal variation and risk assessment.
Ecotoxicology And Environmental SaZELTy, v. 172, p. 480-487, maio 2019.
http://dx.doi.org/10.1016/j.ecoenv.2019.01.102.
GAO, Yanpeng; NIU, Xiaolin; WANG, Mei; LI, Guiying; AN, Taicheng. An inescapable
fact: toxicity increases during photo-driven degradation of emerging contaminants in
water environments. Current Opinion In Green And Sustainable Chemistry, v. 30, p.

122
100472, ago. 2021. http://dx.doi.org/10.1016/j.cogsc.2021.100472.
GARRIC, Jeanne. Emerging Issues in Ecotoxicology: Pharmaceuticals and Personal
Care Products (PPCPs). In: FÉRARD, Jean-François; BLAISE, Christian (ed.).
Encyclopedia of Aquatic Ecotoxicology. Dordrecht: Springer, 2013. Cap. 38. p.
407-428.
HAMAN, Camille; DAUCHY, Xavier; ROSIN, Christophe; MUNOZ, Jean-François.
Occurrence, fate and behavior of parabens in aquatic environments: a review. Water
Research, v. 68, p. 1-11, jan. 2015. http://dx.doi.org/10.1016/j.watres.2014.09.030.
HENN, Kirsten; BRAUNBECK, Thomas. Dechorionation as a tool to improve the fish
embryo toxicity test (FET) with the zebrafish (Danio rerio). Comparative Biochemistry
And Physiology Part C: Toxicology & Pharmacology, v. 153, n. 1, p. 91-98, jan. 2011.
http://dx.doi.org/10.1016/j.cbpc.2010.09.003.
HORTON, Alice A.; WALTON, Alexander; SPURGEON, David J.; LAHIVE, Elma;
SVENDSEN, Claus. Microplastics in freshwater and terrestrial environments: evaluating
the current understanding to identify the knowledge gaps and future research priorities.
Science Of The Total Environment, v. 586, p. 127-141, maio 2017.
http://dx.doi.org/10.1016/j.scitotenv.2017.01.190.
HOTHORN, T., BRETZ, F., WESTFALL, P.. Simultaneous Inference in General
Parametric Models. Biometrical Journal, v. 50, n. 3, p. 346-363, 2008.
HOU, Guoqing; ZHAO, Xiaoli; ZHAO, Tianhui; WU, Xiaowei; PU, Shengyan; TANG, Zhi;
WU, Fengchang. The adsorption of PAHs on microplastics and desorption in the
simulated human digestive system. Chemical Engineering Journal, v. 473, p. 145157,
out. 2023. http://dx.doi.org/10.1016/j.cej.2023.145157.
HU, Chenyan; BAI, Yachen; SUN, Baili; ZHOU, Xiangzhen; CHEN, Lianguo. Exposure
to methylparaben at environmentally realistic concentrations significantly impairs
neuronal health in adult zebrafish. Journal Of Environmental Sciences, v. 132, p.
134-144, out. 2023. http://dx.doi.org/10.1016/j.jes.2022.07.012.
ISLAM, Naimul; FONSECA, Tainá Garcia da; VILKE, Juliano; GONÇALVES, Joanna M.;
PEDRO, Paulo; KEITER, Steffen; CUNHA, Sara C.; FERNANDES, José O.;
BEBIANNO, M.J.. Perfluorooctane sulfonic acid (PFOS) adsorbed to polyethylene
microplastics: accumulation and ecotoxicological effects in the clam Scrobicularia plana.
Marine
Environmental
Research,
v.
164,
p.
105249,
fev.
2021.

123
http://dx.doi.org/10.1016/j.marenvres.2020.105249.
JACKSON, Christopher. flexsurv: A Platform for Parametric Survival Modeling in R.
Journal
of
Statistical
Software,
v.
70,
n.8,
p.
1-33,
2016.
http://dx.doi.org/10.18637/jss.v070.i08
JIN, Xiaowei; WANG, Yeyao; JIN, Wei; RAO, Kaifeng; GIESY, John P.; HOLLERT,
Henner; RICHARDSON, Kristine L.; WANG, Zijian. Ecological Risk of Nonylphenol in
China Surface Waters Based on Reproductive Fitness. Environmental Science &
Technology, v. 48, n. 2, p. 1256-1262, 24 dez. 2013. American Chemical Society
(ACS). http://dx.doi.org/10.1021/es403781z.
JONKERS, Niels; SOUSA, Ana; GALANTE-OLIVEIRA, Susana; BARROSO, Carlos M.;
KOHLER, Hans-Peter E.; GIGER, Walter. Occurrence and sources of selected phenolic
endocrine disruptors in Ria de Aveiro, Portugal. Environmental Science And Pollution
Research, v. 17, p. 834-843, dez. 2009. http://dx.doi.org/10.1007/s11356-009-0275-5
KAIS, B.; SCHNEIDER, K.e.; KEITER, S.; HENN, K.; ACKERMANN, C.; BRAUNBECK,
T.. DMSO modifies the permeability of the zebrafish (Danio rerio) chorion-Implications
for the fish embryo test (FET). Aquatic Toxicology, v. 140-141, p. 229-238, set. 2013.
http://dx.doi.org/10.1016/j.aquatox.2013.05.022.
KALUEFF, Allan V.; STEWART, Adam Michael; GERLAI, Robert. Zebrafish as an
emerging model for studying complex brain disorders. Trends In Pharmacological
Sciences, v. 35, n. 2, p. 63-75, fev. 2014. http://dx.doi.org/10.1016/j.tips.2013.12.002.
KASSAMBARA, A., KOSINSKI, M., BIECEK. P.. survminer: Drawing Survival Curves
using
'ggplot2'
(2024).
R
package
version
0.5.0,
<https://CRAN.R-project.org/package=survminer>.
KHAN, Mohsin; JOHNSON, Kelly. Microplastics alter toxicity of the insecticide Bacillus
thuringiensis israelensis to chironomid larvae in different ways depending on particle
size. Science Of The Total Environment, v. 954, p. 176637, dez. 2024.
http://dx.doi.org/10.1016/j.scitotenv.2024.176637.
KIDD, K. A.; BLANCHFIELD, P. J.; MILLS, K. H.; PALACE, V. P.; EVANS, R. E.;
LAZORCHAK, J. M.; FLICK, R. W.. Collapse of a fish population after exposure to a
synthetic estrogen. Proceedings Of The National Academy Of Sciences, v. 104, n.
21, p. 8897-8901, 21 maio 2007. Proceedings of the National Academy of Sciences.
http://dx.doi.org/10.1073/pnas.0609568104.

124
KIM, Deok-Ho; SUN, Yu; YUN, Seok; KIM, Byungkyu; HWANG, Chang Nam; LEE,
Sang Ho; NELSON, B.J.. Mechanical property characterization of the zebrafish embryo
chorion. The 26Th Annual International Conference Of The Ieee Engineering In
Medicine
And
Biology
Society,
v.
4,
p.
5061-5064,
2004.
http://dx.doi.org/10.1109/iembs.2004.1404399.
KIMMEL, Charles B.; BALLARD, William W.; KIMMEL, Seth R.; ULLMANN, Bonnie;
SCHILLING, Thomas F.. Stages of embryonic development of the zebrafish.
Developmental
Dynamics,
v.
203,
n.
3,
p.
253-310,
jul.
1995.
http://dx.doi.org/10.1002/aja.1002030302.
KINIGOPOULOU,
Vasiliki;
PASHALIDIS,
Ioannis;
KALDERIS,
Dimitrios;
ANASTOPOULOS, Ioannis. Microplastics as carriers of inorganic and organic
contaminants in the environment: a review of recent progress. Journal Of Molecular
Liquids, v. 350, p. 118580, mar. 2022. http://dx.doi.org/10.1016/j.molliq.2022.118580.
KONINGS, Merel C.; ZADA, Liron; SCHMIDT, Robert W.; ARIESE, Freek. Optimization
of sample preparation, fluorescence- and Raman techniques for environmental
microplastics. Spectrochimica Acta Part A: Molecular and Biomolecular Spectroscopy,
v. 319, p. 124537, out. 2024. http://dx.doi.org/10.1016/j.saa.2024.124537.
LA PIETRA, Alessandra; FASCIOLO, Gianluca; LUCARIELLO, Daniela; MOTTA, Chiara
Maria; VENDITTI, Paola; FERRANDINO, Ida. Polystyrene microplastics effects on
zebrafish embryological development: comparison of two different sizes.
Environmental Toxicology And Pharmacology, v. 106, p. 104371, mar. 2024.
http://dx.doi.org/10.1016/j.etap.2024.104371.
LARA, Larissa Zacher; BERTOLDI, Crislaine; ALVES, Nortom Munhoz; FERNANDES,
Andreia Neves. Sorption of endocrine disrupting compounds onto polyamide
microplastics under different environmental conditions: behaviour and mechanism.
Science Of The Total Environment, v. 796, p. 148983, nov. 2021.
http://dx.doi.org/10.1016/j.scitotenv.2021.148983.
LAUGHLIN, Daniel C.; MESSIER, Julie. Fitness of multidimensional phenotypes in
dynamic adaptive landscapes. Trends In Ecology & Evolution, v. 30, n. 8, p. 487-496,
ago. 2015. http://dx.doi.org/10.1016/j.tree.2015.06.003.
LI, Jia; ZHANG, Kaina; ZHANG, Hua. Adsorption of antibiotics on microplastics.
Environmental
Pollution,
v.
237,
p.
460-467,
jun.
2018.

125
http://dx.doi.org/10.1016/j.envpol.2018.02.050.
LIANG, Jiefeng; LIU, Qian S.; REN, Zhihua; MIN, Ke; YANG, Xiaoxi; HAO, Fang;
ZHANG, Qing; LIU, Qian; ZHOU, Qunfang; JIANG, Guibin. Studying paraben-induced
estrogen receptor- and steroid hormone-related endocrine disruption effects via
multi-level approaches. Science Of The Total Environment, v. 869, p. 161793, abr.
2023. http://dx.doi.org/10.1016/j.scitotenv.2023.161793.
LIMONTA, Giacomo; MANCIA, Annalaura; BENKHALQUI, Assja; BERTOLUCCI,
Cristiano; ABELLI, Luigi; FOSSI, Maria Cristina; PANTI, Cristina. Microplastics induce
transcriptional changes, immune response and behavioral alterations in adult zebrafish.
Scientific
Reports,
v.
9,
n.
1,
p.
1-11,
31
out.
2019.
http://dx.doi.org/10.1038/s41598-019-52292-5.
LLORCA, Marta; ÁBALOS, Manuela; VEGA-HERRERA, Albert; ADRADOS, Miquel A.;
ABAD, Esteban; FARRÉ, Marinella. Adsorption and Desorption Behaviour of
Polychlorinated Biphenyls onto Microplastics’ Surfaces in Water/Sediment Systems.
Toxics, v. 8, n. 3, p. 59, 17 ago. 2020. http://dx.doi.org/10.3390/toxics8030059.
LU, Jian; WU, Jie; WU, Jun; ZHANG, Cui; LUO, Yongming. Adsorption and Desorption
of Steroid Hormones by Microplastics in Seawater. Bulletin Of Environmental
Contamination And Toxicology, v. 107, n. 4, p. 730-735, 7 jan. 2020.
http://dx.doi.org/10.1007/s00128-020-02784-2.
MAHARAJAN, Kannan; MUTHULAKSHMI, Sellamani; NATARAJ, Bojan; RAMESH,
Mathan; KADIRVELU, Krishna. Toxicity assessment of pyriproxyfen in vertebrate model
zebrafish embryos (Danio rerio): a multi biomarker study. Aquatic Toxicology, v. 196,
p. 132-145, mar. 2018. http://dx.doi.org/10.1016/j.aquatox.2018.01.010.
MAJEWSKY, Marius; BITTER, Hajo; EICHE, Elisabeth; HORN, Harald. Determination of
microplastic polyethylene (PE) and polypropylene (PP) in environmental samples using
thermal analysis (TGA-DSC). Science Of The Total Environment, v. 568, p. 507-511,
out. 2016. http://dx.doi.org/10.1016/j.scitotenv.2016.06.017.
MALAFAIA, Guilherme; SOUZA, Andreza Martins de; PEREIRA, Aryelle Canedo;
GONÇALVES, Stênio; ARAÚJO, Amanda Pereira da Costa; RIBEIRO, Renan Xavier;
ROCHA, Thiago Lopes. Developmental toxicity in zebrafish exposed to polyethylene
microplastics under static and semi-static aquatic systems. Science Of The Total
Environment,
v.
700,
p.
134867,
jan.
2020.
http://dx.doi.org/10.1016/j.scitotenv.2019.134867.

126
MEROLA, C.; PERUGINI, M.; CONTE, A.; ANGELOZZI, G.; BOZZELLI, M.;
AMORENA, M.. Embryotoxicity of methylparaben to zebrafish (Danio rerio) early-life
stages. Comparative Biochemistry And Physiology Part C: Toxicology &
Pharmacology,
v.
236,
p.
108792,
out.
2020.
http://dx.doi.org/10.1016/j.cbpc.2020.108792.
MEROLA, C.; LAI, O.; CONTE, A.; CRESCENZO, G.; TORELLI, T.; ALLORO, M.;
PERUGINI, M.. Toxicological assessment and developmental abnormalities induced by
butylparaben and ethylparaben exposure in zebrafish early-life stages. Environmental
Toxicology
And
Pharmacology,
v.
80,
p.
103504,
nov.
2020.
http://dx.doi.org/10.1016/j.etap.2020.103504.
MORCILLO, Patricia; ESTEBAN, María Á.; CUESTA, Alberto. Heavy metals produce
toxicity, oxidative stress and apoptosis in the marine teleost fish SAF-1 cell line.
Chemosphere,
v.
144,
p.
225-233,
fev.
2016.
http://dx.doi.org/10.1016/j.chemosphere.2015.08.020.
MRINALINI, R.; TAMILANBAN, T.; KUMAR, V. Naveen; MANASA, K.. Zebrafish – The
Neurobehavioural Model in Trend. Neuroscience, v. 520, p. 95-118, jun. 2023.
http://dx.doi.org/10.1016/j.neuroscience.2022.12.016.
NORTON, William H. J.. Measuring Larval Zebrafish Behavior: locomotion, thigmotaxis,
and startle. Neuromethods, p. 3-20, 2012.
http://dx.doi.org/10.1007/978-1-61779-597-8_1.
NOWAK, Karolina; RATAJCZAK–WRONA, Wioletta; GÓRSKA, Maria; JABłOŃSKA,
Ewa. Parabens and their effects on the endocrine system. Molecular And Cellular
Endocrinology,
v.
474,
p.
238-251,
out.
2018.
http://dx.doi.org/10.1016/j.mce.2018.03.014.
OECD. Test No. 236: Fish Embryo Acute Toxicity (FET) Test. OECD Guidelines For The
Testing Of Chemicals, Section 2, 2013. http://dx.doi.org/10.1787/9789264203709-en.
PARVEZ, Suhel; RAISUDDIN, Sheikh. Protein carbonyls: novel biomarkers of exposure
to oxidative stress-inducing pesticides in freshwater fish channa punctata (bloch).
Environmental Toxicology And Pharmacology, v. 20, n. 1, p. 112-117, jul. 2005.
http://dx.doi.org/10.1016/j.etap.2004.11.002.
PELKA, Katharina E.; HENN, Kirsten; KECK, Andreas; SAPEL, Benjamin;
BRAUNBECK, Thomas. Size does matter – Determination of the critical molecular size

127
for the uptake of chemicals across the chorion of zebrafish (Danio rerio) embryos.
Aquatic
Toxicology,
v.
185,
p.
1-10,
abr.
2017.
http://dx.doi.org/10.1016/j.aquatox.2016.12.015.
PENHA, Larissa Cristine de Carvalho; ROLA, Regina Coimbra; SILVA JUNIOR, Flávio
Manoel da; MARTINS, Camila de Martinez Gaspar. Toxicity and sublethal effects of
methylparaben on zebrafish (Danio rerio) larvae and adults. Environmental Science
And Pollution Research, v. 28, n. 33, p. 45534-45544, 18 abr. 2021.
http://dx.doi.org/10.1007/s11356-021-12800-5.
PLASTICSEUROPE. An analysis of European plastics production, demand and waste
data. Plastics — The facts. Association of Plastic Manufacturers. Bruxelas, Bélgica, 64
pp, 2020. Disponível em: https://www.plasticseurope.org/en/resources/market-data.
Acesso em: 17 out 2021.
PINHEIRO-DA-SILVA, Jaquelinne; SILVA, Priscila Fernandes; NOGUEIRA, Marcelo
Borges; LUCHIARI, Ana Carolina. Sleep deprivation effects on object discrimination
task in zebrafish (Danio rerio). Animal Cognition, v. 20, n. 2, p. 159-169, 19 set. 2016.
http://dx.doi.org/10.1007/s10071-016-1034-x.
PINHEIRO-DA-SILVA, Jaquelinne; AGUES-BARBOSA, Thais; LUCHIARI, Ana
Carolina. Embryonic Exposure to Ethanol Increases Anxiety-Like Behavior in Fry
Zebrafish. Alcohol And Alcoholism, v. 55, n. 6, p. 581-590, 4 set. 2020.
http://dx.doi.org/10.1093/alcalc/agaa087.
POMPEI, Caroline Moço Erba; CAMPOS, Luiza Cintra; SILVA, Bianca Ferreira da;
FOGO, José Carlos; VIEIRA, Eny Maria. Occurrence of PPCPs in a Brazilian water
reservoir and their removal efficiency by ecological filtration. Chemosphere, v. 226, p.
210-219, jul. 2019. http://dx.doi.org/10.1016/j.chemosphere.2019.03.122.
QIANG, Liyuan; CHENG, Jinping. Exposure to polystyrene microplastics impairs
gonads of zebrafish (Danio rerio). Chemosphere, v. 263, p. 128161, jan. 2021.
http://dx.doi.org/10.1016/j.chemosphere.2020.128161.
RAFA, Nazifa; AHMED, Bushra; ZOHORA, Fatema; BAKYA, Jannatul; AHMED,
Samiya; AHMED, Shams Forruque; MOFIJUR, M.; CHOWDHURY, Ashfaque Ahmed;
ALMOMANI, Fares. Microplastics as carriers of toxic pollutants: source, transport, and
toxicological effects. Environmental Pollution, v. 343, p. 123190, fev. 2024.
http://dx.doi.org/10.1016/j.envpol.2023.123190.

128
RAJA, Glancis Luzeena; SUBHASHREE, K. Divya; LITE, Christy; SANTOSH, Winkins;
BARATHI, S.. Transient exposure of methylparaben to zebrafish (Danio rerio) embryos
altered cortisol level, acetylcholinesterase activity and induced anxiety-like behaviour.
General And Comparative Endocrinology, v. 279, p. 53-59, ago. 2019.
http://dx.doi.org/10.1016/j.ygcen.2018.11.001.
REED, Barney; JENNINGS, Maggy. Guidance on the housing and care of zebrafish
(Danio rerio). Association for Assessment and Accreditation of Laboratory Animal Care
International (AAALAC International). Research Animals Department, Science
Group, RSPCA, nov. 2010. Atualizado em: maio 2011.
REZANIA, Shahabaldin; PARK, Junboum; DIN, Mohd Fadhil Md; TAIB, Shazwin Mat;
TALAIEKHOZANI, Amirreza; YADAV, Krishna Kumar; KAMYAB, Hesam. Microplastics
pollution in different aquatic environments and biota: a review of recent studies. Marine
Pollution
Bulletin,
v.
133,
p.
191-208,
ago.
2018.
http://dx.doi.org/10.1016/j.marpolbul.2018.05.022.
RIBAS, Laia; PIFERRER, Francesc. The zebrafish (Danio rerio) as a model organism,
with emphasis on applications for finfish aquaculture research. Reviews In
Aquaculture, v. 6, n. 4, p. 209-240, 31 jul. 2013. http://dx.doi.org/10.1111/raq.12041.
RIBEIRO, Renan Xavier; BRITO, Rafaella da Silva; PEREIRA, Aryelle Canedo;
MONTEIRO, Keyle Borges e Silva; GONÇALVES, Bruno Bastos; ROCHA, Thiago
Lopes. Ecotoxicological assessment of effluents from Brazilian wastewater treatment
plants using zebrafish embryotoxicity test: a multi-biomarker approach. Science Of The
Total
Environment,
v.
735,
p.
139036,
set.
2020.
http://dx.doi.org/10.1016/j.scitotenv.2020.139036.
RICHARDSON, Susan D.; KIMURA, Susana Y.. Water Analysis: emerging
contaminants and current issues. Analytical Chemistry, v. 92, n. 1, p. 473-505, 11 dez.
2019.
American
Chemical
Society
(ACS).
http://dx.doi.org/10.1021/acs.analchem.9b05269.
ROJE, Željka; ILIĆ, Krunoslav; GALIĆ, Emerik; PAVIČIĆ, Ivan; TURČIĆ, Petra;
STANEC, Zdenko; VRČEK, Ivana Vinković. Synergistic effects of parabens and plastic
nanoparticles on proliferation of human breast cancer cells. Archives Of Industrial
Hygiene And Toxicology, v. 70, n. 4, p. 310-314, 1 dez. 2019. Walter de Gruyter
GmbH. http://dx.doi.org/10.2478/aiht-2019-70-3372.
ROPER, Courtney; TANGUAY, Robert L.. Zebrafish as a Model for Developmental

129
Biology and Toxicology. Handbook Of Developmental Neurotoxicology, p. 143-151,
2018. http://dx.doi.org/10.1016/b978-0-12-809405-1.00012-2.
ROUTLEDGE, Edwin J.; PARKER, Joanne; ODUM, Jenny; ASHBY, John; SUMPTER,
John P.. Some Alkyl Hydroxy Benzoate Preservatives (Parabens) Are Estrogenic.
Toxicology And Applied Pharmacology, v. 153, n. 1, p. 12-19, nov. 1998.
http://dx.doi.org/10.1006/taap.1998.8544.
RYAN, J. A.; HIGHTOWER, L. E.. Stress proteins as molecular biomarkers for
environmental toxicology. Stress-Inducible Cellular Responses, p. 411-424, 1996.
http://dx.doi.org/10.1007/978-3-0348-9088-5_28.
SANTOS, Mauricius Marques dos; BREHM, Franciane de Almeida; FILIPPE, Tais
Cristina; KNAPIK, Heloise Garcia; AZEVEDO, Júlio César Rodrigues de. Occurrence
and risk assessment of parabens and triclosan in surface waters of southern Brazil: a
problem of emerging compounds in an emerging country. Rbrh, v. 21, n. 3, p. 603-617,
set. 2016. http://dx.doi.org/10.1590/2318-0331.011616018.
SAUD, Shah; YANG, Aoao; JIANG, Zhaoyu; DUNING,; FAHAD, Shah. New insights in
to the environmental behavior and ecological toxicity of microplastics. Journal Of
Hazardous
Materials
Advances,
v.
10,
p.
100298,
maio
2023.
http://dx.doi.org/10.1016/j.hazadv.2023.100298.
SCHMID, Chiara; COZZARINI, Luca; ZAMBELLO, Elena. Microplastic's story. Marine
Pollution
Bulletin,
v.
162,
p.
111820,
jan.
2021.
http://dx.doi.org/10.1016/j.marpolbul.2020.111820.
SEQUEIRA, Inês F.; PRATA, Joana C.; COSTA, João P. da; DUARTE, Armando C.;
ROCHA-SANTOS, Teresa. Worldwide contamination of fish with microplastics: a brief
global overview. Marine Pollution Bulletin, v. 160, p. 111681, nov. 2020.
http://dx.doi.org/10.1016/j.marpolbul.2020.111681.
SJOBERG D, BAILLIE M, FRUECHTENICHT C, HAESENDONCKX S, TREIS T.
ggsurvfit: Flexible Time-to-Event Figures (2024). R package version 1.1.0,
<https://CRAN.R-project.org/package=ggsurvfit>.

STANTON, Thomas; JOHNSON, Matthew; NATHANAIL, Paul; MACNAUGHTAN,
William; GOMES, Rachel L.. Freshwater microplastic concentrations vary through both
space and time. Environmental Pollution, v. 263, p. 114481, ago. 2020.

130
http://dx.doi.org/10.1016/j.envpol.2020.114481.
TERANIKAR, Tanveer; NGUYEN, Phuc; LEE, Juhyun. Biomechanics of cardiac
development in zebrafish model. Current Opinion In Biomedical Engineering, v. 27,
p. 100459, set. 2023. http://dx.doi.org/10.1016/j.cobme.2023.100459.
THERNEAU, T.. A Package for Survival Analysis in R (2024). R package version
3.7-0, <https://CRAN.R-project.org/package=survival>
THOMPSON, Richard C.; OLSEN, Ylva; MITCHELL, Richard P.; DAVIS, Anthony;
ROWLAND, Steven J.; JOHN, Anthony W. G.; MCGONIGLE, Daniel; RUSSELL,
Andrea E.. Lost at Sea: where is all the plastic?. Science, v. 304, n. 5672, p. 838-838, 7
maio 2004. American Association for the Advancement of Science (AAAS).
http://dx.doi.org/10.1126/science.1094559.
TREVISAN, Rafael; UZOCHUKWU, Daniel; GIULIO, Richard T. di. PAH Sorption to
Nanoplastics and the Trojan Horse Effect as Drivers of Mitochondrial Toxicity and PAH
Localization in Zebrafish. Frontiers In Environmental Science, v. 8, n. 78, p. 1-26, 24
jul. 2020. Frontiers Media SA. http://dx.doi.org/10.3389/fenvs.2020.00078.
WANG, Jun; LI, Yuejiao; LU, Lin; ZHENG, Mingyi; ZHANG, Xiaona; TIAN, Hua; WANG,
Wei; RU, Shaoguo. Polystyrene microplastics cause tissue damages, sex-specific
reproductive disruption and transgenerational effects in marine medaka (Oryzias
melastigma). Environmental Pollution, v. 254, p. 113024, nov. 2019.
http://dx.doi.org/10.1016/j.envpol.2019.113024.
WANG, Jing; WU, Jin; CHENG, Haodong; WANG, Yudi; FANG, Yanjun; WANG, Lei;
DUAN, Zhenghua. Polystyrene microplastics inhibit the neurodevelopmental toxicity of
mercury in zebrafish (Danio rerio) larvae with size-dependent effects. Environmental
Pollution, v. 314, p. 120216, dez. 2022. http://dx.doi.org/10.1016/j.envpol.2022.120216.
WATANABE, Yoko; KOJIMA, Hiroyuki; TAKEUCHI, Shinji; URAMARU, Naoto; OHTA,
Shigeru; KITAMURA, Shigeyuki. Comparative study on transcriptional activity of 17
parabens mediated by estrogen receptor α and β and androgen receptor. Food And
Chemical
Toxicology,
v.
57,
p.
227-234,
jul.
2013.
http://dx.doi.org/10.1016/j.fct.2013.03.036.
WEI, Fang; CHENG, Hefa; SANG, Nan. Comprehensive assessment of estrogenic
activities of parabens by in silico approach and in vitro assays. Science Of The Total
Environment,
v.
845,
p.
157194,
nov.
2022.

131
http://dx.doi.org/10.1016/j.scitotenv.2022.157194.
WILKINSON, John; HOODA, Peter S.; BARKER, James; BARTON, Stephen;
SWINDEN, Julian. Occurrence, fate and transformation of emerging contaminants in
water: an overarching review of the field. Environmental Pollution, v. 231, p. 954-970,
dez. 2017. http://dx.doi.org/10.1016/j.envpol.2017.08.032.
WISENDEN, Brian D.; PAULSON, Daniel C.; ORR, Megan. Zebrafish embryos hatch
early in response to chemical and mechanical indicators of predation risk, resulting in
underdeveloped swimming ability of hatchling larvae. Biology Open, v. 11, n. 12, p. 1-7,
29 nov. 2022. http://dx.doi.org/10.1242/bio.059229.
XUE, Jingchuan; KANNAN, Kurunthachalam. Accumulation profiles of parabens and
their metabolites in fish, black bear, and birds, including bald eagles and albatrosses.
Environment
International,
v.
94,
p.
546553,
set.
2016.
http://dx.doi.org/10.1016/j.envint.2016.06.015.
XUE, Xiaohong; XUE, Jingchuan; LIU, Wenbin; ADAMS, Douglas H.; KANNAN,
Kurunthachalam. Trophic Magnification of Parabens and Their Metabolites in a
Subtropical Marine Food Web. Environmental Science & Technology, v. 51, n. 2, p.
780-789,
5
jan.
2017.
American
Chemical
Society
(ACS).
http://dx.doi.org/10.1021/acs.est.6b05501.
YONG, Cheryl; VALIYAVEETTIL, Suresh; TANG, Bor. Toxicity of Microplastics and
Nanoplastics in Mammalian Systems. International Journal Of Environmental
Research And Public Health, v. 17, n. 5, p. 1509, 26 fev. 2020.
http://dx.doi.org/10.3390/ijerph17051509.

132
CONSIDERAÇÕES GERAIS

Os microplásticos são uma classe de contaminantes extremamente variável
em suas características físicas e químicas, havendo evidências crescentes de sua
ubiquidade em ambientes aquáticos, bem como de sua ecotoxicidade e dos efeitos
biológicos de sua interação com outros tipos de contaminantes orgânicos e inorgânicos
encontrados em águas superficiais. Dentre os contaminantes que poderiam interagir
com microplásticos e alterar sua toxicidade, os parabenos se destacam por serem
substâncias detectadas em águas residuais, águas superficiais e em tecidos animais de
organismos aquáticos de diversas localizações, com efeitos tóxicos documentados a
partir de estudos experimentais realizados em microrganismos, invertebrados e
vertebrados de variadas espécies. Esta dissertação provê uma extensa revisão da
literatura acerca de todo o conhecimento atual que diz respeito à bioacumulação e
ecotoxicidade de parabenos e seus metabólitos em organismos aquáticos. Além disso,
conduz um estudo preliminar que busca investigar a interação de um polímero plástico
de ampla produção e descarte com o parabeno mais extensamente encontrado em
águas superficiais a partir de concentrações ambientalmente relevantes de ambos os
contaminantes. Dessa forma, traz conhecimento inovador a respeito do comportamento
e ecotoxicidade destes contaminantes ambientais emergentes em um cenário realista,
buscando contribuir para com o arcabouço de informações que sustenta regulações e
políticas públicas acerca da produção, descarte e remediação de tais contaminantes em
matrizes ambientais.

133
APÊNDICE I
Figura 1. Captura de tela da primeira página do artigo de revisão publicado na revista
Environmental Pollution em novembro de 2024.

134
APÊNDICE II
Table 1. Chemical and physical properties of main parabens∗.
CHARACTERISTICS/
PARABENS

4-HB/PHBA

MeP

EtP

PrP

BuP

PhP

BzP

HeP

Chemical name

4-hydroxybenzoic acid

Methylparaben

Ethylparaben

Propylparaben

Butylparaben

Phenylparaben

Benzylparaben

Heptylparaben

Chemical formula

C7H6O3

C8H8O3

C9H10O3

C10H12O3

C11H14O3

C13H10O3

C14H12O3

C14H20O3

Molecular weight (g M-1)

138.12

152.15

166.17

180.2

194.23

214.22

228.25

236.31

CAS n.

99-96-7

99-76-3

120-47-8

94-13-3

94-26-8

17696-62-7

94-18-8

6259-77-4

Solubility in water at 25°C (mg
L-1) (MG/l)

5000

2500

885

500

∼150

∼10

∼20

∼1

Log octanol-water (Log KOW)

1.58

1.96

∼2.47

3.04

3.57

∼3.5

3.71

4.52

Log acid dissoc. constant (pKA)

∼4.54

∼8.17

8.1

∼8.2

8.37

∼8.4

∼8.4

∼8.4

Half-life (hours)

—

35.2#

27.5#

20.3#

9.6#

–

10–19h##

–

Biodegradability (in water) ∗∗

Readily

Readily

Readily to
slower rate

Readily to slower
rate

Readily to slower
rate

Slow rate

Slow rate

Slow rate

135
Data obtained from Barabasz et al. (2019) Gonzalez-Marino et al. (2011)a,b, Kim et al. (2023), Nowak et al. (2018), Vale et al. (2022), and Yamamoto et al. (2007). (**) Paraben’s longer
alkyl chain makes it more lipophilic and less water-soluble, which can lead to a slower degradation rate in the environment. ( #) Data from wastewater test. ( ##) Data from activated
sludge batch. (¡) Data not available.

APÊNDICE III
Table 2. Summary of experimental studies with parabens in aquatic organisms regarding the types and concentrations of
parabens used, biomarkers investigated and model species.
Species

Stage of
development

Type of
parabens

Concentrations

Biomarkers

Exposure
period

Exposure
Type

Analytica
l control

Bioaccum
ulation
Analysis

Effects

Reference

MICROORGANISMS
Vibrio fischeri

—

Pseudokirchneriella
subcapitata

—

MeP, EtP, PrP,
iPrP, BuP, iBuP,
BzP and their
chlorinated
compounds
MeP, EtP, PrP,
iPrP, BuP, iBuP,

Unspecified

Bioluminescence

5, 15 min

Static

No

No

↓bioluminescen
ce

Terasaki et al.,
2009

Unspecified
(established

Growth rate

72 h

Static

No

No

↓growth rate for
longer alkyl

Yamamoto et
al., 2011

136
BzP

according to
LOEC)

chain PBs at
high
concentrations

—

MeP

Unspecified (EC50
= 35.25 mg/L)

Growth inhibition

72 h

Static

No

No

↑growth
inhibition

Di Poi et al.,
2017

Aliivibrio fischeri

—

MeP, EtP, PrP,
4-HB

Unspecified

5, 15 min

Static

No

No

↓bioluminescen
ce, respiration

Ortíz de García
et al., 2014

Aliivibrio fischeri

__

MeP, PrP

5, 15 min

__

No

No

↓bioluminescen
ce

Dailianis et al.,
2023

Acinetobacter
calcoaceticus

__

MeP, PrP, BuP,
mixture

1.75, 2.5, 5, 10
μg/L and 0.25, 0.5,
1, 3.75, 7.5, 15, 30
mg/L
150 ng/L

Bioluminescence,
respiration
inhibition
Bioluminescence

7-26 d

Semi-static
for (2 d) for
26
d-exposure

No

No

↑cellular
culturability and
density

Pereira et al.,
2023

Freshwater biofilm

__

MeP, BuP

Unspecified

Virulence factors,
cellular
culturability,
density, and
thickness
Type of cell deatlh,
cell wall damage,
ROS generation

24 h

Static

No

No

Liu et al., 2023b

Periphyton biofilm

—

BuP

0,5, 50 and 5000
µg/L

Total biomass,
chlorophyll a,
algae diversity and
biovolume,
photosynthetic
efficiency, carbon
source utilizing
capacity

32 d

Flowthrough

No

No

BuP promoted
necrosis while
MeP
apoptosis/both
↑ cell wall
dagame and
ROS production
No effect at
environmentally
relevant
concentrations
(0,5 µg/L),
↓algae growth,
algal diversity
and
photosynthetic
efficiency at the
highest
concentration

(12h)

Song et al.,
2016

137
Stenotrophomonas
maltophilia

—

MeP, PrP, BuP,
mixture

150 ng/L

Virulence factors,
cellular
culturability,
density, and
thickness

7-26 d

Semi-static
for (2 d) for
26
d-exposure

No

No

↑cellular density
and thickness,
protease and
gelatinase
production

Pereira et al.,
2023

INVERTEBRATES
Daphnia magna

Neonates

Neonates

MeP, EtP, PrP,
iPrP, BuP, iBuP,
BzP and their
mono- and
dichlorinated
compounds
MeP, EtP, PrP,
iPrP, BuP, iBuP,
BzP

Unspecified

EC50
(immobilization)

48 h

Static

No

No

↑mortality with
chlorinated and
longer alkyl
chain PBs

Terasaki et al.,
2008

Unspecified
(calculated
accordingly to
LC50)

LC50, mortality,
reproduction

48 h (acute
exposure),
10 d
(subchronic
exposure)

Static

Yes

No

↑mortality for
higher MeP
concentration
(12 mg/L),
↓growth for
longer alkyl
chain PBs
↑immobilization
for longer alkyl
chain PBs and
higher
concentrations
of shorter alkyl
chain PBs
↑transcription of
genes related to
oxidative stress
(with UV light),
↓growth and
number of
offspring in a
dose-dependen
t manner,
↓population
growth rate
↑immobilization

Dobbins et al.,
2009

Neonates

MeP, EtP, PrP,
iPrP, BuP, iBuP,
BzP

Unspecified
(calculated
accordingly to
LOEC)

Immobilization

48 h (acute
exposure),
21 d
(chronic
exposure,
MeP only)

Static

No

No

Neonates

MeP

Acute exposure: 0,
0.37, 0.75, 1.50,
3.12, and 6.25
mg/L
Chronic exposure:
0, 0.1, 0.3, 1.0,
3.2, and 10.0 mg/L
with or without UV
light

Immobilization,
growth, number of
living offspring,
gene expression

48 h (acute
exposure),
21 d
(chronic
exposure)

Static

Yes

No

Neonates

MeP

Immobilization

48 h

Static

No

No

Juvenile

MeP, EtP, PrP,
BuP

Unspecified (EC50
= 41.23 mg/L)
Acute exposure: 0,
0.1, 1.0, 10 mg/L

Behavior,
neurotoxicity,

48 h

Static

No

No

Disruption of
cardio and

Yamamoto et
al., 2011

Lee et al., 2017

Di Poi et al.,
2017
Eghan et al.,
2023

138

Acartia tonsa

Nauplii

Artemia franciscana

cardiotoxicity and
gene expression
Larval
development
LC50, oxidative
stress, AChE

Nauplii

MeP, EtP, PrP,
BuP, BzP
MeP

Unspecified (use
of wastewater)
0.0085 and 0.017
mg/L

Tigriopus japonicus

Nauplii

MeP, EtP, PrP

Acute exposure: 0,
5000, 7500,
10,000, 15,000,
and 20,000 µg/L
(MeP and EtP); 0,
100, 200, 300,
400, and 500 µg/L
(PrP)
Chronic exposure:
0, 10, 100, 1000,
and 10,000 µg/L
(MeP); 0, 3.75,
37.5, 375, and
3750 µg/L (EtP); 0,
0.05, 0.5, 5, and
50 µg/L (PrP)

LC50,
development,
reproduction rate

Ceriodaphnia dubia

Adults and
neonates

MeP, PrP, iPrP,
BzP and their
chlorinated
compounds

Unspecified

Paracentrotus
lividus

Embryos and
larvae

PrP

Crassostrea gigas

Embryos and
larvae

MeP

neurobehavioral
functions
↑inhibition of
development
↓CAT

5d

Static

No

No

Kusk et al.,
2011
Comeche et al.,
2017

24 h (LC50)
and 9 d
(chronic
exposure)
96 h (acute
exposure)

Static

No

No

Static

No

No

Developmental
delay,
↓reproduction
rate in higher
concentrations

Kang et al.,
2019

EC50 (Mortality,
offspring number
and first brood)

7d

Static

Yes

No

↓offspring
number for
MeP, BzP and
dichlorinated
BzP
↑mortality with
longer alkyl
chains and
nonchlorinated
PBs

Terasaki et al.,
2013

10, 64, 100, 160,
400, 1000 and
10000 µg/L

Larvae length,
malformations

48 h

Static

No

No

↓larvae length
↑malformations

Torres et al.,
2016

Unspecified (EC50
= 18.57 mg/L —
embryotoxicity;
EC50 = 7.88 —

Embryotoxicity,
metamorphosis

36 h
(embryotoxi
city), 24 h
(metamorph

Static

No

No

↑abnormal
larvae,
↓metamorphosi
s rate

Di Poi et al.,
2017

139
Dugesia japonica

—

MeP, EtP, PrP,
BuP

Mytilus
galloprovicialis

—

MeP, PrP

Strongylocentrotus
purpuratus

Larvae

MeP

metamorphosis)
0, 6.25, 12.5, 25,
50 (MeP); 0, 50,
100, 200, 400
(EtP); 0, 1, 5, 10,
25 (PrP); 0, 1, 10,
25, 50 (BuP)
0.3, 2.0 µg/L

0.1, 0.5, 1 and 5
mg/L

Mobility

osis)
10 min

Static

Yes

No

↓mobility in
higher
concentrations

Li, 2020

Oxidative stress,
lysosomal
membrane
integrity,
genotoxicity

4 days

Semi-static
(24 h)

Yes

No

disturbance of
lysosomal
membrane
integrity, ↑
superoxides,
nitric oxides,
lipid
peroxidation,
superoxides,
and
micronuclei.

Dailianis et al.,
2023

Fertilization
success, survival,
development,
body length

40 min
(fertilization)
,
3 h (larval
developmen
t), 96 h
(larval
growth and
survival)

Static

No

No

↓survival, ↓body
length in the
highest dose

Shore et al.,
2022

VERTEBRATES
Oncorhynchus
mykiss

Juvenile

EtP, PrP, BuP,
4-HB

Juvenile

PrP

Juvenile

BuP

100 and 300
mg/kg (EtP, PrP
and 4-HB), 50, 150
and 200 mg/kg
(BuP)
7 to 1830 mg/kg/2
days (oral
administration)
50 and 225 µg/L
(water exposure)
4 to 74 mg/kg/2
days (oral
administration)

VTG

Intraperitone
al injections
at days 0
and 6

Static

No

No

↑VTG and
mortality

Pedersen et al.,
2000

VTG,
toxicokinetics

10 d (oral
administrati
on), 12 d
(water
exposure)
10 d (oral
administrati
on), 12 d

Static

No

Liver and
muscle

↑VTG

Bjerregaard et
al., 2003

Static

No

Liver and
muscle

↑VTG

Alslev et al.,
2005

LC50, VTG, tissue
concentration

140

Cyprinus carpio

Juvenile

Embryos

MeP

MeP, PrP, BuP

Pimephales
promelas

Larvae

MeP, EtP, PrP,
iPrP, BuP, iBuP,
BzP

Oryzias latipes

Adults

PrP

Larvae

MeP, EtP, PrP,
iPrP, BuP, iBuP,
BzP

Embryos,
eleuthero
embryos and
larvae

PrP

35 and 201 µg/L
(water exposure)
0, 0.84, 1.68 and
4.20 mg/L

LC50,
histopathology,
VTG, behavior,
organ mass,
bioaccumulation,
hepatic enzymes

(water
exposure)
96 h (acute
exposure),
28 d
(subchronic
exposure)

MeP 0.5; 50;
500; 5000;
100,000 µg/L; PrP
0.1; 10; 100; 1000;
100,000 µg/L; BuP
0.1; 10; 100; 1000;
100,000 µg/L
Unspecified
(calculated
accordingly to
LC50)

Mortality, hatching,
development,
oxidative damage,
and gene
expression.

96 h

LC50, mortality,
growth

0.055, 0.55, 5.5
and 55 mM
Unspecified
(established
accordingly to
LOEC)

VTG, gene
expression
VTG, gene
expression

48 h (acute
exposure), 7
d
(subchronic
exposure)
1 week

40, 400, 1000 and
4000 µg/L

Embryotoxicity,
histopathology,
EROD,
post-eclosion
development

Static

No
Testes,
liver, brain,
gills and
muscle

Semi-static

↓ASP, ACP,
testiculosomatic
index, ↑ALT,
ALK, liver size,
VTG,
bioaccumulatio
n in testis, liver,
brain, gills and
muscle

Barse et al.,
2010

No

No

↑Mortality,
↓hatching rates,
↑malformations,
gene
expression
downregulation

Medkova et al.,
2023

Static

No

No

↓growth for
longer alkyl
chain PBs

Dobbins et al.,
2009

Static

No

No

Inui et al., 2003

96 h (acute
exposure),
14 d
(chronic
exposure,
MeP only)

Static

No

No

↑VTG, ERS1
and CHG
↑VTG in males,
upregulation of
12 genes
(including vtg2,
choriogenin and
esr1),
downregulation
of 10 genes

10 d

Static

No

No

↓development
rate,
↑malformations,
weakness,
histological
defects and
mortality at the
highest
concentration,

González-Donc
el et al., 2014

(24h)

Yamamoto et
al., 2011

141
EROD,
gallbladder area

Oreochromis
niloticus

Adults

MeP and PrP

0.400, 2.00 and
10.0 mg/L (MeP);
0.320, 1.00 and
3.20 mg/L (PrP)

Adults and
parental
generation

2-EHHB

Unspecified
(established
according to
OCSPP MEOGRT
890.2200)

Adults

MeP, EtP, PrP,
BuP, BzP, MeP
+ PrP

Juvenile

BzP

Adults

BuP

4.0 mg/L (all
parabens alone),
6.0 mg/L MeP +
1.7 mg/L PrP
0, 5, 50, 500 and
5000 ng/L

5, 50, 500 and
5000 ng/L

Fecundity and
fertility (spawning
status), VTG,
secondary sex
characteristics,
body length and
weight, ESR1, Arβ
Fecundity and
fertility (spawning
status), VTG,
secondary sex
characteristics,
body length and
weight

21 d

Static

Yes

No

Agonistic
activity towards
ESR1, ↑VTG in
males,
↓fecundity

Kawashima et
al., 2021

32 weeks

Flowthrough

Yes

No

Testicular
hypoplasia/atro
phy, reduced
liver glycogen,
and
effects on body
weight and
length

Matten et al.,
2023

↓GSH (6 days),
↑SOD, GPx,
GR, GSH (12
days)
↑metabolic
disorders of
hepatic glycerol
phospholipids,
glycerolipids
and
sphingomyelins,
crude fat
content,
oxidative stress
and liver tissue
inflammation
↓brain AChE

Silva et al.,
2018

(72 min)

LC50, oxidative
stress in gills and
liver

12 d

Semi-static

Yes

No

Lipid metabolism,
hepatic
morphology,
oxidative stress,
brain AChE

8 weeks

Semi-static

No

No

Histology,
neurotransmitters,
gene expression

56 d

Yes

No

(24h)

Semi-static
(24h)

Darker skin
pigmentation,
↑Tyr, Arr3a
↓dopamine,
Asip2,

Lin et al., 2022

Liu et al., 2023a

142
Danio rerio

Juvenile

PrP

500, 1000 or 2000
mg/kg

Length and
weight, VTG, sex
ratio

20 d (VTG),
45 d (sex
ratio)

Semi-static

Yes

No

↑proportion of
females in the
lowest
concentration

Mikula et al.,
2009

(24h)

Embryos

PrP

10, 100, 1000,
3500, 6000, 8500
and 10000 µg/L

Embryotoxicity

80 h

Static

No

No

↑developmental
delay,
malformations,
mortality at the
highest
concentration
↓heart rate,
hatching rate

Torres et al.,
2016

Embryos and
larvae

MeP

100, 200, 400, 800
and 1000 µM

96 h

Static

No

No

MeP

50 mg/L

68 h

Static

No

No

↑malformations,
VTG-I
↓heart rate,
hatching rate
↑malformations,
mortality, lipid
peroxidation,
myca and
ccnd1
↓GST, NO,
distance swam

Dambal et al.,
2017

Embryos and
larvae

LC50,
embryotoxicity,
gene expression
(VTG-I only)
Embryotoxicity,
oxidative stress,
gene expression,
behavior,
apoptosis

Adults

MeP

0.001, 0.01, 1 and
10 mg/L

Survival, length,
weight,
gonadosomatic
index,
histopathology of
the testis

21 d

Semi-static

No

No

Testicular
atrophy,
↓gonadosomati
c index,
multinucleated
gonocytes,
impaired germ
cells, Leydig
cell hyperplasia,
interstitial
fibrosis,
apoptosis of
Sertoli cells

Hassanzadeh,
2017

NOEC/LOEC,
embryotoxicity,
pancreatic
malformations,

165 h

No

No

↑malformations,
beta cell area,
aberrant
pancreatic islet

Brown et al.,
2018

Embryos and
larvae

BuP

0, 250, 500, 1000
and 3000 nM

(24h)

Static

Ateş et al.,
2017

143
oxidative stress,
gene expression,
Embryos and
larvae

MeP

0.1, 1, 10 and 100
ppb

Embryotoxicity,
neurotoxicity
(AChE, cortisol),
behavior

142 h

Static

No

No

Embryos and
larvae

MeP and PrP

1, 10, 25, 50, 100
and 200 µM
(embryotoxicity), 1
and 10 µM (gene
expression)

Embryotoxicity,
gene expression

118 h

Static

No

No

Embryos and
larvae

MeP

1, 10, 30, 60 and
80 mg/L

LC50,
embryotoxicity

96 h

Static

No

No

Embryos and
larvae

EtP, BuP

5, 10, 20 and 30
mg/L (EtP), 1, 2.5
mg/L (BuP)

LC50,
embryotoxicity,
behavior

96 h

Static

No

No

Embryos and
larvae

PrP

1, 2, 4, 6 and 8
mg/L

Embryotoxicity,
lipid metabolism

96 h

Static

No

No

morphologies,
GSH, gsr
↓pdx1
↑cortisol
↓AChE, latency
to reach and
time spent in
the upper part
of the tank,
heart rate,
hatching rate
↑malformations,
mortality
↓hatching rate,
altered
expression of
30 genes
related to cell
cycle, DNA
damage,
inflammation,
fatty acid
metabolism and
endocrine
function
↑malformations
↓heart rate,
survival at the
highest
concentrations
↑malformations,
behavioral
abnormalities
↓heart rate,
blood
circulation,
hatching rate
↑malformations,
yolk sac size
↓hatching rate,
swim bladder
size, embryo
length, head
length, survival

Luzeena-Raja
et al., 2018

Bereketoglu
and Pradhan,
2019

Merola et al.,
2020a

Merola et al.,
2020b

Perugini et al.,
2020

144

Embryos and
larvae

MeP, EtP, BuP

Embryos and
larvae

MeP, EtP, PrP,
BuP

Larvae and
adults

MeP

Adults

MeP

Embryos and
larvae

PrP

100, 1000 e 10000
µg/L (MeP), 50,
500 e 5000 µg/L
(EtP), 5, 50 e 500
µg/L (BuP)
20 to 200 µM
(MeP), 20 to 100
µM (EtP), 5 to 20
µM (PrP),
2 to 10 µM (BuP)
30 µL, 50 mg/L
(adults), 50 mg/L
(larvae)

1, 10 and 110 ppb

10 and 10000 µg/L

Behavior

96 h

Static

No

No

Embryotoxicity,
endocrine
dysregulation
(thyroid), gene
expression
LC50, NOEC,
oxidative stress,
genotoxicity, gut
microbiome

120 h

Static

No

No

96 h
(adults), 168
h (larvae)

Static

No

No

Neurotoxicity
(AChE, 5-HT),
gene expression,
behavior

30 d

Semi-static

No

No

Oxysterols

24 h

No

No

(24h)

Static

at the highest
concentrations,
PLA2
↑thigmotaxis
(EtP and BuP)

↓survival (EtP,
PrP and BuP),
T3 and T4
↑malformations,
cell proliferation
↓EROD in
adults (gills) at
high
concentration
↑lipid
peroxidation,
kidney nuclei
and micronuclei
(erythrocytes) in
adults, carbon
sources utilized
by gut
microbiota in
adults
↓AChE
↑5-HT,
anxiety-like
behavior in
females
Dysregulation
of cardiac
hypoxia and
neuronal
differentiation-re
lated genes
↑27-OH,
↓7a-OH and
7b-OH at 8hpf,
↑24-OH at 24
hpf,
non-detection of

Merola et al.,
2021

Liang et al.,
2021

Penha et al.,
2021

Thakkar et al.,
2022

Merola et al.,
2022

145

Embryos and
larvae

PrP, BuP

0.1, 1 and 10 ppb

Embryotoxicity,
anxiety behavior,
oxidative stress in
the brain,
apoptosis in the
head, AChE, NO

96 h

Static

No

No

Adults

MeP

0, 1, 3 and 10 µg/L

Gene expression,
histopathology of
liver, oxidative
stress,
metabolomic
profile of liver

28 d

Semi-static

No

No

Growth, gut
microbiome, 5-HT,
TJP2, goblet cells,
proinflammatory
genes, oxidative
stress

28 d

No

No

Heart morphology
and
histopathology,
heart rate, gene

90 h

No

No

Adults

Embryos

MeP

EtP

0, 1, 3 and 10 µg/L

0.1, 0.5, 1, 2, 3, 4,
5, 6, 10, 15, 18,
20, 25, 30, 42, 50,
80 and 100 mg/L

(24h)

Semi-static
(24h)

Static

22-OH and
25-OH at 8 and
24 hpf
↓hatching rate
(PrP), heart
rate, SOD, CAT,
GPx, GST,
GSH, AChE
↑mortality,
malformations,
scototaxis, NO,
ROS, lipid
peroxidation,
apoptosis in the
head
↑hepatic cortisol
in males,
↓synthesis and
conjugation of
primary bile
acid,
↑degradation of
estradiol and
retinoic acid,
↑hepatocellular
vacuolization,
↑redox
imbalance
↑body length
and weight in
females,
↑proinflammator
y cytokines in
females,
↑intestinal
dysbiosis,
↑goblet cells in
males, ↓TJP2
and serotonin in
females
↑abnormalities
in heart
morphology and
function,

Lite et al., 2022

Hu et al., 2022a

Hu et al., 2022b

Fan et al., 2022

146
expression

Embryos and
Adults

MeP

0, 1, 3 and 10 µg/L

Embryotoxicity,
gonad histology,
sex hormones,
gene expression

28 d

Semi-static

No

No

(24h)

Embryos

MeP, EtP, PrP
and BuP

20, 100 and 200
µM (MeP), 20, 50
and 100 µM (EtP),
2, 5 and 10 (PrP),
1, 2 and 5 µM
(BuP)

Hormones, gene
expression

120 h

Static

No

No

Adults

MeP

1, 3 and 10 µg/L

Neural proteome,
oxidative stress,
AChE, glutamate,
gene expression

28 d

Semi-static

No

No

(24h)

disruption of
retinoic acid
signaling
pathway, ↓gene
related to
myocardial
contraction
↑gonadosomati
c index,
↑mortality of
offspring, early
hatching,
blockage of
oogenesis,
disbalance of
sex hormones,
upregulation of
17 genes,
downregulation
of 29 genes
Downregulation
of 14 genes,
upregulation of
6 genes, ↑VTG,
↑estradiol
(BuP),
↓testosterone
Downregulation
of 33 differential
proteins in
males and 88 in
females,
upregulation of
31 differential
proteins in
males and 89 in
females,
downregulation
of two genes,
↑glutamate in
the male brain,
ROS
↓glutamate in
the female

Hu et al., 2022c

Liang et al.,
2023a

Hu et al., 2023

147

Mauremys sinensis

brain, CAT,
GSH
Downregulation
of four genes
(MeP),
upregulation of
two genes
(EtP), ↑AChE in
two exposure
groups ↓ACTH,
total movement
distance, mean
velocity
↑Mortality,
↓hatching rates,
↑malformations
(at higher
concentrations),
gene
expression
downregulation
Abnormal
embryonic
development;
behavioral
hyperactivity

Embryos and
larvae

MeP, EtP, PrP,
BuP

20, 100 and 200
µM (MeP), 20, 50
and 100 µM (EtP),
2, 5 and 10 µM
(PrP), 1, 2 and 5
µM (BuP)

Swimming
behavior, AChE,
cortical hormones,
gene expression

118 h

Static

No

No

Embryos

MeP, PrP, BuP

MeP 0.5; 50;
500; 5000;
100,000 µg/L; PrP
0.1; 10; 100; 1000;
100,000 µg/L; BuP
0.1; 10; 100; 1000;
100,000 µg/L

Lethal and
sublethal
endpoints and
gene expression

96 h

Static

No

No

Embryos and
Larvae

MeP, EtP, PrP

Embryotoxicity,
neurotoxicity,
behavior, gene
expression

120 h

Static

No

No

Embryos

BuP

5, 10, 20, 40, 80,
150 and 300 µM
(embryotoxicity
test)
Non-specified for
other tests (based
on embryonic
mortality)
0.6 mg/L, 1.2
mg/L, and 1.8
mg/L

Cardiotoxicity,
oxidative stress,
and gene
expression

72 h

Static

No

No

Cardiac
morphological
defects and
functional
impairment;
Cardiac
oxidative stress
and
immunosuppres
sion

Zhu et al., 2023

Adults

BuP

Gut microbiome,
gut histopathology,
cytokines, gene
expression

20 weeks

Semi-static

No

Intestine

Gut microbiome
dysbiosis;
inflammatory
response,

Ding et al.,
2023

5, 50 and 500 µg/L

(48h)

Liang et al.
2023b

Medkova et al.,
2023

Tran et al.,
2023

148

Adults

Xenopus laevis

Embryos and
larvae

BuP

MeP, PrP, BuP

5, 50 and
500 µg/L

MeP 0.5; 50; 500;
5000; 100,000
µg/L; PrP 0.1; 10;
100; 1000;
100,000 µg/L; BuP
0.1; 10; 100; 1000;
100,000 µg/L

Liver oxidative
stress and gene
expression

20 weeks

Lethal and
sublethal
endpoints and
gene expression

96 h

Semi-static

No

No

No

No

(48h)

Static

shortened
intestinal villi
↑ MDA, ↓ SOD,
CAT, GSH,
dysregulation of
genes related to
Nrf2-Keap1
pathway,
inflammatory
and apoptosis,
several
histopathologic
al findings
↑Mortality (at
higher
concentrations),
gene
expression
downregulation

Yin et al.. 2023

Medkova et al.,
2023