ArticlePDF Available

Abstract and Figures

A range of substances that are released into the environment, foodstuffs and drinking water as a result of human activity were originally considered relatively harmless, and it was only later that their adverse effects were discovered. In general the use of such substances is currently restricted, and they are often replaced by other substances. This applies also in the case of a range of endocrine disruptors. These substances have the capacity to disturb the balance of physiological functions of the organism on the level of hormonal regulation, and their pleiotropic spectrum of effects is very difficult to predict. Endocrine disruptors include the currently intensively studied bisphenol A (BPA), a prevalent environmental pollutant and contaminant of both water and foodstuffs. BPA has a significantly negative impact on human health, particularly on the regulation mechanisms of reproduction, and influences fertility. The ever increasingly stringent restriction of the industrial production of BPA is leading to its replacement with analogues, primarily with bisphenol S (BPS), which is not subject to these restrictions and whose impacts on the regulation of reproduction have not yet been exhaustively studied. However, the limited number of studies at disposal indicates that BPS may be at least as harmful as BPA. There is therefore a potential danger that the replacement of BPA with BPS will become one of the cases of regrettable substitution, in which the newly used substances manifest similar or even worse negative effects than the substances which they have replaced. The objective of this review is to draw attention to ill-advised replacements of endocrine disruptors with substances whose effects are not yet tested, and which may represent the same risks for the environment, for the reproduction of males and females, and for human health as have been demonstrated in the case of the originally used substances.
Content may be subject to copyright.
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
Supported by the Ministry of Agriculture of the Czech Republic (Project No. QJ1510138 and Project No.
MZeRO0714), by the Ministry of Education, Youth and Sports of the Czech Republic (Project No. LO1503 under
the NPU I program), by the Internal Grant Agency of the Czech University of Life Sciences Prague (CIGA) (Project
No. 20132035), and by the Grant Agency of the Charles University in Prague (PRVOUK P36 program).
Bisphenol S instead of bisphenol A: a story
of reproductive disruption by regretable substitution
– a review
T. Ž1, K. H1, J. N2,3, Š. P1, K. Z1,
T.K4, J. P2
1Department of Veterinary Sciences, Faculty of Agrobiology, Food and Natural Resources,
Czech University of Life Sciences Prague, Prague, Czech Republic
2Laboratory of Reproductive Medicine, Biomedical Center, Faculty of Medicine in Pilsen,
Charles University in Prague, Pilsen, Czech Republic
3Department of Histology and Embryology, Faculty of Medicine in Pilsen,
Charles University in Prague, Pilsen, Czech Republic
4Institute of Animal Science, Prague-Uhříněves, Czech Republic
ABSTRACT: A range of substances that are released into the environment, foodstuffs and drinking water as a
result of human activity were originally considered relatively harmless, and it was only later that their adverse effects
were discovered. In general the use of such substances is currently restricted, and they are often replaced by other
substances. is applies also in the case of a range of endocrine disruptors. ese substances have the capacity to
disturb the balance of physiological functions of the organism on the level of hormonal regulation, and their pleio-
tropic spectrum of effects is very difficult to predict. Endocrine disruptors include the currently intensively studied
bisphenol A (BPA), a prevalent environmental pollutant and contaminant of both water and foodstuffs. BPA has
a significantly negative impact on human health, particularly on the regulation mechanisms of reproduction, and
influences fertility. e ever increasingly stringent restriction of the industrial production of BPA is leading to its
replacement with analogues, primarily with bisphenol S (BPS), which is not subject to these restrictions and whose
impacts on the regulation of reproduction have not yet been exhaustively studied. However, the limited number of
studies at disposal indicates that BPS may be at least as harmful as BPA. ere is therefore a potential danger that
the replacement of BPA with BPS will become one of the cases of regrettable substitution, in which the newly used
substances manifest similar or even worse negative effects than the substances which they have replaced. e objec-
tive of this review is to draw attention to ill-advised replacements of endocrine disruptors with substances whose
effects are not yet tested, and which may represent the same risks for the environment, for the reproduction of males
and females, and for human health as have been demonstrated in the case of the originally used substances.
Keywords: human health; environment; endocrine disruptor; reproduction; oocyte; sperm
Many substances have been introduced into use
with great hopes, only for it to be demonstrated
earlier or later that they are harmful to the environ-
ment and/or human health. Notorious cases include
the mass use of DDT as an insecticide (http://apps., thalidomide
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
as a drug for pregnant women (McBride 1961), or
more recently neonicotinoid insecticides used for
the protection of fields against seed-destroying
insects (Blacquiere et al. 2012). Substances whose
negative effects on the environment or human health
were detected only after a long period of use also
include endocrine disruptors (Damstra et al. 2002).
The detection of the negative effects of abun-
dantly used substances leads to a dramatic restric-
tion of their use and their substitution with other
substances. In a range of cases this brings about
a genuine improvement. For example, chromated
copper arsenate (CCA) used for wood preserva-
tion was demonstrated to be a substance with
carcinogenic effects, and as a result was replaced
with alkaline copper quaternary (ACQ). ACQ does
not contain arsenic or chrome, and although it is
just as effective as CCA against wood destroy-
ing arthropods, its impacts on the environment
and human health are fundamentally less serious
(Landrigan et al. 2004).
On the other hand, we have been witnesses to
substitutions of harmful substances which have later
been shown to be highly problematic. For example,
2,3-butanedione, which occurs naturally in butter,
has been produced synthetically and added to foods
in order to impart a buttery flavour. When it was
demonstrated that 2,3-butanedione damaged lung
tissue, it was replaced by 2,3-pentanedione, which
however was subsequently proven to have similar
negative effects on lung tissue as 2,3-butanedione
(Hubbs et al. 2012). ere are far more similar ex-
amples of “regrettable substitutions” (Fahrenkamp-
Uppenbrink 2015; Zimmerman and Anastas 2015).
In these cases, negative impacts on reproduction are
often subsequently detected. For example, in the
case of pyrethroids, which replaced older insecticide
agents such as organocholorines, organophosphates
or carbamates, and which were considered harmless
to mammals, negative impacts were demonstrated on
the maturation of mammal oocytes (Petr et al. 2013).
From the perspective of reproductive risks, the
substitution of bisphenol A (BPA), a widely used
component of plastics and many other materials,
with its analogue bisphenol S (BPS) appears to be
potentially problematic. BPA has been proven to be
a strong endocrine disruptor, and its use has been
restricted. Many products are sold with a “BPA-free”
guarantee. Because BPA is substituted in a range
of cases by BPS, these products are not however
“bisphenol-free” (Glausiusz 2014), and their use
may be linked to significant reproductive risks. The
aim of this review is to point to the replacement of
BPA by BPS as a “regrettable substitution.
Endocrine disruptors
A less harmful substitute is currently searched
for a number of substances that had previously
been considered safe from a toxicological perspec-
tive and finally appeared to exert various negative
effects on health. This category of compounds
includes substances referred to summarily as en-
docrine disruptors (Clayton 2011). According to
the US Environmental Protection Agency, endo-
crine disrupting chemicals (EDCs) are defined as
“exogenous agent(s) that interfere(s) in synthesis,
secretion, transport, metabolism, binding action,
or elimination of natural blood-borne hormones
that are present in the body and are responsible
for homeostasis, reproduction, and developmental
processes” (Diamanti-Kandarakis et al. 2009).
EDCs manifest a range of particular properties.
Their hormone-like effects may be suppressed or
may fade away entirely in the case that the concen-
tration of EDCs is higher than the physiological
level of their hormonal counterpart. This ability
of agents to attain paradoxically stronger effects
in low doses than in high ones (vom Saal and
Welshons 2005) is termed the “low dose effect”
(Grasselli et al. 2010; Vandenberg et al. 2012). The
low dose hypothesis posits that exogenous che-
micals that interact with hormone action can do
so in a quite specific manner. In accordance with
that, mentioned traditional toxicological endpoints
are not capable to preclude adverse outcome, as
EDCs act with dose responses, that are nonlinear
and potentially non-monotonic (Vandenberg et al.
2012). In the case the relationship between dose
and response is nonlinear, any prediction is even
more complex. Therefore, the low dose definition
was extended by the effects of non monotonic
response curves. The mechanisms responsible
for the non-linear effects are described in detail
(Vandenberg et al. 2012), usually in connection
with an interaction between a ligand (hormone or
EDC) and a hormone receptor (Vandenberg 2014).
Non-linear dose-response patterns are com-
monly observed with endogenous and synthetic
agonists (e.g. numerous drugs, hormones, peptides)
that activate and inhibit receptor-mediated signal
pathways that affect various biological functions
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
(Calabrese and Baldwin 2001; Calabrese 2005). Ho-
wever, EDCs can also produce non monotonic dose
responses in which the slope of the curve changes
sign over the course of the dose-response (www.
and low dose effects are described for the majority
of EDCs (Birnbaum 2012; Vandenberg et al. 2012,
2013; Zoeller et al. 2012; Bergman et al. 2013).
The concept of endocrine-disrupting chemicals
was proposed after these compounds had been
observed to affect various reproductive functions
in wildlife and humans (Colborn et al. 1993). The
influence of several EDCs was demonstrated on the
course of development of male gametes, sperm (Li
et al. 2011; Knez et al. 2014) and female gametes,
oocytes, as well as embryonic development of males
and females (Mok-Lin et al. 2010; Xiao et al. 2011).
Moreover, the effect of EDCs on the reproduction
of adult individuals, including transgenerational
inheritance, has been described (Susiarjo et al.
2015; Ziv-Gal et al. 2015). Therefore, reproductive
functions represent crucial targets of the EDCs’
negative effects. Recently intensively studied EDCs,
interfering with the regulation of physiological re-
productive processes, include bisphenols, a family
of chemical compounds with two hydroxyphenyl
functional groups (Figure 1).
Bisphenol A
An example of a widely used substance, in which
endocrine-disrupting properties were detected
only later, is bisphenol A (BPA, 4,4'-(propane-2,2-
diyl)diphenol) (Vandenberg et al. 2009). BPA was
first synthesized in 1891, and as early as in 1936
it was demonstrated that it imitates the activity of
the hormone estradiol (Dodds and Lawson 1936).
Despite a very strong estrogen activity, BPA has
been commercially used since 1957, and despite
the fact that its endocrine-disrupting activity was
discovered (Krishnan et al. 1993), BPA has become a
high production volume chemical (Wang et al. 2012).
Worldwide annual production, which in the case
of BPA reached 4.6 million t in 2012, is constantly
increasing. Its production was estimated at 5.4 mil-
lion t in 2015 (Merchant Research & Consulting,
BPA is present especially in polycarbonate plas-
tics, epoxide resins, and several paper products
(Ehrlich et al. 2014), and as a result it is used in a
variety of commonly used consumer products such
as thermal recipes, cosmetics, dental materials,
medicinal tubes, utensils, toys, baby feeding bot-
tles and dummies, etc. Heat, UV radiation, alkaline
treatment or intensive washing causes a release of
BPA monomer. It is estimated that the worldwide
release of BPA into the environment is almost half
million kg per year (Mileva et al. 2014).
BPA is released into the environment either di-
rectly from chemical, plastic coating, and staining
manufacturers, from paper or material recycling
companies, foundries which use BPA in casting
sand, or indirectly leaching from plastic, paper, and
waste in landfills (Yang et al. 2015). BPA passes into
foodstuffs or water directly from the lining of food
and beverage cans, where it is used as an ingredi-
ent in the plastic used to protect the food from
direct contact with the can (Goodson et al. 2002;
Vandenberg et al. 2009). The main path of human
exposure is the consumption of such contaminated
foodstuffs, drinking water or via dermal contact
with thermal paper and cosmetics or inhalation
(Miyamoto and Kotake 2005; Huang et al. 2012).
It is therefore not surprising that a range of stud-
ies have now demonstrated the presence of BPA
in human tissue. Levels of BPA have been tested
in various populations worldwide, and the pres-
ence of BPA was demonstrated in 92.6% of Ameri-
cans (Wetherill et al. 2007) and 90% of Canadians
(Bushnik et al. 2010). Levels of BPA have been
demonstrated in various biological matrices, most
frequently in urine (Casas et al. 2013; Salgueiro-
Gonzalez et al. 2015), but also in blood serum.
Within the human reproductive system, levels of
BPA have been confirmed for example in testicle
tissue, seminal plasma (Manfo et al. 2014), in ovar-
ian follicular fluid (Ikezuki et al. 2002), mother’s
Figure 1. Chemical structure of bisphenol A (A), bisphe-
nolS (B), bisphenol F (C)
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
milk, fetal plasma (Shonfelder et al. 2002), amniotic
fluid (Yamada et al. 2002; Edlow et al. 2012), and the
placenta (Jimenez-Diaz et al. 2010; Cao et al. 2012)
(Table 1). Several studies have demonstrated a direct
correlation between exposure of the mother and the
BPA level of the fetus (Ikezuki et al. 2002; Kuruto-
Niwa et al. 2007). BPA may permeate the placenta
and thus influence the development of the fetus
(Edlow et al. 2012; Corbel et al. 2014). Newborns
may then be further exposed to the effect of BPA
during breastfeeding due to the presence of BPA in
mother’s milk (Mendonca et al. 2014).
The effects of BPA on humans are dependent not
only on the dose, but also on the window of exposure.
Exposure to BPA in the prenatal and neonatal period
probably affects the human organism in the most
receptive period (Fernandez et al. 2014).
Mechanism of BPA action
A typical feature of endocrine disruptors is their
wide spectrum of outcomes (Figure 2). Combi-
nation of their action in various target systems
in the organism is one of causes of their non-
linear effects. In this respect, BPA acts as a typical
endocrine disruptor with multi-level impacts (Khan
and Ahmed 2015). Nongenomic effects of BPA have
been described, thus influencing cellular signalling
Table 1. Bisphenol A (BPA) levels in human fluids
Sample Level of BPA References
Blood (ng/ml) 12.4–14.4 Bushnik et al. (2010)
Maternal blood (ng/ml) 0.63–14.36 Yamada et al. (2002)
Fetal blood (ng/ml) 0.2–9.2 Schonfelder et al. (2002)
Urine (ng/ml) 0.02–21.0 Liao et al. (2012c)
Saliva (ng/ml) 0.3 Joskow et al. (2006)
Follicular fluid (ng/ml) 2.4 ± 0.8 Ikezuki et al. (2002)
Amniotic fluid (ng/ml) 1.1–8.3 Ikezuki et al. (2002)
Placental tissue (ng/g) 1.0–104.9 Schonfelder et al. (2002)
Breast milk (ng/ml) 0.5–1.3 Mendonca et al. (2014)
Semen plasma (pg/ml) 66 (fertile men)
132–179 (infertile men) Vitku et al. (2015)
Figure 2. Possible mecha-
nisms of bisphenol action
and its potential impact
on human health
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
(Nakagawa and Tayama 2000), as well as genomic,
which affect transcription regulation (Trapphoff et
al. 2013), and also epigenetic, responsible for the
methylation and acetylation of DNA and core his-
tones (Bromer et al. 2010). It is precisely pronounced
estrogen activity of BPA in vitro (vom Saal et al. 2007;
Wetherill et al. 2007) and in vivo that contributes to
its immense potential to afflict the hormonal system
and act as an endocrine disruptor.
BPA inhibits the activity of natural endogenous
estrogens and thus disrupts estrogen nuclear
hormone receptor action (Kitamura et al. 2005;
Wetherill et al. 2007; Grignard et al. 2012). BPA
affects hormonal homeostasis, for example through
bonding to the classic nuclear estrogen receptors
α, β, γ (ERα, ERβ, ERγ), where it manifests a com-
bination of agonistic and/or antagonistic actions
in dependence on the target tissue, cell types, ER
subtypes, and differential cofactors recruited by
ER-ligand complexes (Kurosawa et al. 2002). BPA
also bonds to non-classical membrane ERs and
causes activation of the nuclear receptor gamma
(Takayanagi et al. 2006; Matsushima et al. 2007).
BPA has been identified as an antagonist of an-
drogen receptors (Kitamura et al. 2005; Wetherill et
al. 2007; Vinggaard et al. 2008; Molina-Molina et al.
2013). Its anti-androgenic activity has been docu-
mented in several studies, but with changing values
of the maximum inhibition concentration (Xu et al.
2005; Bonefeld-Jorgensen et al. 2007). In contrast
with other known androgen receptor antagonists,
BPA inhibits the effective nuclear translocation of
the androgen receptors, and disrupts their function
by means of a number of mechanisms (Teng et al.
2013). The endocrine-related BPA action mechanism
also involves a reduction of aromatase expression
(Zhang et al. 2011; Chen et al. 2014) and a decrease
in aromatase activity in vitro (Bonefeld-Jorgensen
et al. 2007). Within this context, it is of interest
that a decline in the synthesis of testosterone and
estradiol in vivo has been documented following
exposure to BPA (Akingbemi et al. 2004).
The epigenetic mechanisms of the effect of BPA
include the alteration of certain DNA methylation
samples (Dolinoy et al. 2007; Susiarjo et al. 2013).
Prenatal exposure to BPA alters the expression of
genes coding individual subtypes of ERs in a sex-
and brain region-specific manner (Kundakovic et
al. 2013) and disrupts the normal development of
the placenta (Susiarjo et al. 2013). As a result, it is
possible that BPA predetermines the response to
steroid hormones in the very early phase of devel-
opment (Wilson and Sengoku 2013). It has been
documented that BPA also disrupts the gene ex-
pression of the regulating factors that control the
stability and flexibility of epigenetic regulation,
and as a result has an adverse influence on the
development of functions of the controlling organ
of hormonal regulation, the hypothalamus (Warita
et al. 2013). The impacts of these changes have
transgenerational effects (Manikkam et al. 2013).
Further demonstrated actions of BPA in the
organism include the bonding to the glucuronide
receptor, suppression of the transcription receptor
of the thyroid hormone, reduction of the transport
of cholesterol via the mitochondrial membrane,
increase of oxidation of fatty acids, stimulation of
prolactin release (Machtinger and Orvieto 2014)
or an agonistic effect on the human pregnaneX
receptor (Sui et al. 2012).
BPA and human health
With such a wide spectrum of effects, it is evi-
dent that BPA has a negative influence on hu-
man health. Frequently discussed themes include
the possible association of BPA for example with
obesity (Trasande et al. 2012), diabetes (Lang et
al. 2008), neurobehavioural disorders (Jasarevic
et al. 2011), cancer (Jenkins et al. 2011), hepatic
(Peyre et al. 2014) and cardiovascular diseases,
hypertension, and disorders of the thyroid gland
function (Rochester 2013; Wang et al. 2013).
Especially in the area of reproduction in both
animal models and in humans, a wide range of
negative influences of BPA have been observed
(Kwintkiewicz et al. 2010; Trapphoff et al. 2013;
Zhang et al. 2014). BPA has varied and complex
mechanisms of action that may interfere with
normal reproductive development and functions.
In both males and females, BPA interferes with
hormonal regulation and influences the hypotha-
lamic–pituitary–gonadal axis on all levels (Navarro
et al. 2009; Patisaul et al. 2009; Xi et al. 2011).
Influences of BPA on reproduction of males
As a rule, endocrine-disrupting substances have
pronounced impacts on the reproduction of both
sexes. Several studies have shown detrimental
effects of BPA on spermatogenesis and semen
quality in fishes. The number of mature and im-
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
mature spermatozoa was decreased and increased,
respectively (Sohoni et al. 2001) and also the sperm
motility and concentration were reduced (Lahn-
steiner et al. 2005). There is a large evidence that
BPA can induce sex reversal from male to female
in aquatic animals. Changes in sex ratio were
observed at zebrafish during embryonic develop-
ment (Drastichova et al. 2005) and Xenopus larvae
through metamorphosis (Kloas et al. 1999).
Experimental studies on the effects of BPA on
the reproduction of male rodents have revealed
an adverse influence on the development of testes
(Vrooman et al. 2015) and on the spermatogenesis
of adult individuals following prenatal in utero
or early postnatal exposure. Exposure to BPA
during the period of development of the testes
is frequently linked to a range of negative effects
in adult testes, e.g. decreased levels of testicular
testosterone, decreased weights of the epididymis
and seminal vesicles, a decrease in daily sperm
production per gram testis, and increased weights
of the prostate and preputial (Richter et al. 2007).
Vrooman et al. (2015), with the help of transplan-
tation of spermatogonia from the testes of mice
exposed to the action of BPA into mice which were
not exposed, demonstrated permanent damage to
spermatogenesis. The influence of the exposure
of adult rodents to BPA on the quality of sperm
was also studied (Peretz et al. 2014).
Despite the differences in the experimental de-
signs used, certain findings appear repeatedly,
especially reduction in the number of sperm, reduc-
tion in the motility of sperm, increased amount of
apoptotic cells in the seminiferous tubules, changes
in the levels of hormones and steroid enzymes, and
damage to the DNA of sperm (Peretz et al. 2014).
Contemporary studies confirm that rodents are not
relevant for predicting the effect of low BPA concen-
trations on the endocrine function of human fetal
testis (N’Tumba-Byn et al. 2012). In a comparative
study by Maamar et al. (2015), the influence of BPA
was studied both on rats and on human fetal testes,
and it was determined that in both cases BPA had
dose-dependent anti-androgenic effects. Neverthe-
less, the authors urge caution in interpreting the
results obtained on rodents and their application
in human medicine (Maamar et al. 2015).
Unfortunately, there is only a limited number
of studies that have observed the influence of
exposure to BPA on the quality of sperm in adult
humans. In men exposed to BPA in the workplace
and patients in reproduction centres, a higher
level of BPA in urine was linked to a lower num-
ber, concentration, and motility of sperm (Knez
et al. 2014; Lassen et al. 2014). Nevertheless, in
a study conducted by Mendiola et al. (2010) on
fertile men, the concentration of BPA in urine did
not correlate with changes in semen parameters,
despite the fact that a significant correlation was
observed between the level of BPA in urine and
the volume of seminal plasma or markers of free
testosterone (Mendiola et al. 2010).
The following cohort study examined the re-
lationship between the concentration of BPA in
urine and the level of reproductive hormones
and semen in a group of 308 young healthy men.
It was determined that the concentration of BPA
strictly correlates with higher levels of selected
circulating reproductive hormones and reduced
motility of sperm. The results indicated that the
exposure to BPA on the level of environment has
an anti-androgenic and/or anti-estrogenic effect
due to the effect of BPA on the level of recep-
tors. The anti-estrogenic effect on the level of
the epididymis also explains the determined low
mobility of the sperm (Lassen et al. 2014).
Influences of BPA on reproduction of females
BPA markedly influences not only the reproduc-
tion of males, but also the reproduction of females.
In both in vitro and in vivo studies, the influence of
BPA has been demonstrated on fertility, function of
the womb i.e. formation of benign and malignant
lesions (Newbold et al. 2009), disruption apoptosis
of the uterine epithelium during estrus (Mendoza-
Rodriguez et al. 2011), function of ovaries and
quality of oocytes (Peretz et al. 2014), and defec-
tive folliculogenesis (Santamaria et al. 2016). In
females it is precisely the ovaries that are the key
organ responsible for reproductive and endocrine
functions, and BPA is frequently indicated as an
ovarian toxicant. BPA afflicts not only the overall
morphology and weight of the ovaries (Suzuki et al.
2002; Santamaria et al. 2016) but also demonstrably
reduces the quality of oocytes in both animal and
human models (Machtinger and Orvieto 2014).
During the course of the maturation of mouse
oocytes in vitro following treatment with BPA,
changes were documented in the configuration of
the meiotic spindle resulting in errors in chromosome
segregation and hyperploidy frequencies in mouse
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
oocytes (Hunt et al. 2003). Similary, it was reported
that BPA exposure altered chromosome and spindle
organization which resulted in hyperploidy of mouse
oocytes during meiosis (Can et al. 2005) and it was
also demonstrated that low BPA doses are related
with aberration during meiotic prophase, including
increased incidence of recombination (Susiarjo et
al. 2007) and failure formation of primordial follicle
by inhibiting meiotic progression of oocytes (Zhang
et al. 2012). In contrast, Eichenlaub-Ritter and her
colleagues found no evidence that low BPA doses
increased hyperploidy at meiosis II. On the other
hand they observed cell cycle delay and meiotic
spindle abnormalities, changes in the distribution
of pericentriolar material and chromosome align-
ment (Eichenlaub-Ritter et al. 2008). Exposure of
mice, from mid-gestation to birth, causes synaptic
abnormalities in oocytes and an increased amount of
recombination between homologous chromosomes.
It is also of interest that identical effects have been
observed in homozygous mice with an intentionally
disrupted gene coding the ERβ. In mouse oocytes,
epigenetic changes have also been documented fol-
lowing cultivation of follicles in the presence of
BPA, in which a disruption of the configuration of
chromosomes took place, as well as disorders of
meiosis caused by faulty genomic imprinting and
altered posttranslational modification of histones
(Trapphoff et al. 2013). Chronic exposure of oocytes
was linked to an increased incidence of aberrant
metaphases II and prematurely segregated chromatids
(Pacchierotti et al. 2008).
Bovine oocytes cultivated in the presence of
BPA have also manifested disorders of the meiotic
spindle and the chromosomal configuration (Ferris
et al. 2015). In Barbary Macaques, negative effects
of BPA have been demonstrated in various stages
of the oogenesis of developing ovaries. Oocytes
in the prophase of meiosis and in fetal ovaries
exhibited an increased number of recombination,
and an increased number of abnormally formed
follicles containing multiple oocytes was recorded
in perinatal ovaries (Hunt et al. 2012).
Similary as in the aforementioned studies on ro-
dents, cattle, and primates, an increased number of
crossing over and degenerations in oocytes have been
determined also in human oocytes cultivated in vitro
in the presence of BPA (Brieno-Enriquez et al. 2011).
In connected studies it has been demonstrated that
the exposure of human oocytes to BPA is linked to
up-regulation of genes involved in meiotic processes
connected to double strand breaks repair progression
(Brieno-Enriquez et al. 2012). A non-linear response
to BPA doses on the incidence of MII oocytes with
aligned chromosomes has also been determined
(Machtinger et al. 2013). e changes which have been
recorded in the development of oocytes exposed to
bisphenol may lead to disorders in the development
of embryos, fetal loss or genetic disorders (Rama
Raju et al. 2007; Ye et al. 2007; Tomari et al. 2011).
e result of maternal exposure to BPA may be the
disruption of the entire oogenesis in the developing
ovary (Susiarjo et al. 2007).
A number of cohort studies have been focused
on groups of persons who undergo treatment for
infertility through in vitro fertilization (IVF). The
measured levels of BPA in these persons were ex-
amined in connection with the ovarian response,
quality of embryos and implantation. A reduced
ovarian response was linked to a reduced success
rate of IVF (Mok-Lin et al. 2010). BPA also dis-
rupted embryonal development of fish via delay
hatching, yolk reabsorption, and larval growth of
trouts (Aluru et al. 2010), moreover lethality in
zebrafish larvae increased (Chan and Chan 2012).
There is only a limited number of studies which
have observed the effects of BPA on the develop-
ment and quality of mammalian blastocysts. Failure
of embryonic development to mouse blastocyst
stage has been demonstrated after exposure of
females to BPA (Xiao et al. 2011). Disorder of
implantation of mouse blastocysts was also dem-
onstrated by Borman et al. (2015).
In human, Bloom et al. (2011) state a correlation
between the concentration of BPA in the urine
of men, though not in women, and a decline in
the quality of embryos generated by IVF. By con-
trast, in a study performed by Knez et al. (2014),
which confirms changes to the semen quality of
men with a determined environmental level of
BPA, undisrupted development of embryos into
blastocysts is described. As against this finding,
in women who have undergone IVF, a correlation
has been demonstrated between the concentration
of BPA in urine and a change to the formation of
blastocysts, though a reduced quality of embryos
was not recorded (Ehrlich et al. 2012).
The advent of BPS
The above-stated facts led to the necessity for
stringent regulation of the use of BPA, and in a
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
range of cases its substitution with another chemi-
cal. On the basis of the effects on human health
and reproduction demonstrated with the help of
standardized toxicological testing procedures,
government agencies in the United States (the US
Environmental Protection Agency, USEPA), Canada
(Health Canada), and Europe (the European Food
Safety Authority, EFSA) have established tolerable
daily intake levels, ranging from 25 to 50 g BPA/kg
of body weight (BW) per day (Rochester 2013).
With regard to the fact that several studies have
demonstrated BPA low dose effects (Vandenberg
et al. 2012), and that this possibility is unfortu-
nately not taken into account in the approach of
“traditional” toxicological studies, in which low
doses are not generally subjected to examination
(Vandenberg et al. 2012; Rochester 2013), scientists
have expressed concerns that the “safe” cut-off set
for BPA is too high (vom Saal and Hughes 2005).
In 2010 the Canadian government prohibited the
import, sale, and advertisement of baby feeding
bottles containing BPA. The European Union re-
sponded with a prohibition of the manufacture of
baby feeding bottles with BPA, which was passed
in 2011 (Commission Directive 2011). The Food
and Drug Administration (FDA) has indicated
BPA as a “chemical of concern, and in July 2012 a
blanket prohibition of BPA in baby feeding bottles
and sippy cups was recommended (FDA 2011).
However, new data and refined methodologies
have led EFSA experts to considerably reduce
the safe level of BPA from 50 µg/kg of BW/day to
4µg/kg of BW/day (EFSA 2014).
With regard to these restrictions and societal pres-
sures, manufacturers of plastics are now forced to
seek an alternative product which can replace BPA.
It is in the interest of chemical concerns that the
substitute which replaces BPA is inert or at least far
less toxic than BPA. Nevertheless, new chemicals
introduced onto the market are frequently untest-
ed, and may be equally or more harmful than the
originals, which are ultimately termed “regrettable
substitutions” (Rochester and Bolden 2015), as has
been the case of a number of perfluorinated chemi-
cals (Howard 2014), pesticides (Coggon 2002), and
self-extinguishing compounds (Bergman et al. 2012).
Manufacturers seeking BPA alternatives have turned
primarily to bisphenol S (BPS, 4,4'-sulfonyldiphe-
nol) (see Figure1), a structural analogue of BPA, to
produce “BPA-free” products (Grignard et al. 2012;
Barrett 2013). BPS is chemically more stable, worse
in terms of biodegradability than BPA, and shows
better dermal penetration than BPA (Ike et al. 2006;
Danzl et al. 2009; Liao et al. 2012a, b). It is discon-
certing that these properties may lead to a longer or
higher body burden or bioavailability of BPS versus
BPA (Helies-Toussaint et al. 2014). For these reasons,
too, at present the replacement of BPA with BPS is
considered a “regrettable substitution” (Fahrenkamp-
Uppenbrink 2015; Zimmerman and Anastas 2015).
With regard to the increase in production of BPS and
the indispensability of bisphenols in the production
of plastics, it is unfortunately possible to expect the
same widespread use of BPS as in the case of BPA
(Liao et al. 2012c). Now the presence of BPS can be
expected in almost all the consumer goods here in
which BPA was initially used (Mathew et al. 2014),
for example as a wash fastening agent in clearing
products, an electroplanting solvent, and a constitu-
ent of phenolic resins (Rochester and Bolden 2015).
One of the major industries that have replaced
BPA due its high occurrence (~3–22 g/kg) is that
of thermal paper (Mathew et al. 2014). In the
USA, Korea, Vietnam, Japan, and China (Liao et
al. 2012c), BPS has been detected in several differ-
ent “BPA free” paper products, including receipts
and paper money (Liao et al. 2012a). The presence
of BPS has been determined in tinned foodstuffs
(Vinas et al. 2010). The occurrence of BPS has
also been determined in indoor dust (Liao et al.
2012b), in fluvial water (Ike et al. 2006), surface
water, and waste waters (Song et al. 2014) (Table 2).
The main pathway to the human body is dermal,
dust ingestion, and dietary exposures (Liao et al.
2012b). Unfortunately, for example thermal paper
carries BPS into all recycled paper products, mak-
ing dermal exposure inevitable. Massive exposure
of the population to the effects of environmental
BPS has been demonstrated in a number of differ-
ent countries. Within the range of 0.02–21 ng/ml
(0.8–84nM) it has been detected in human urine
samples originating from seven Asian countries
and the USA (Liao et al. 2012a) in 81% of analyzed
samples. In the following study the presence of
BPS in urine was demonstrated in residents living
near a manufacturing plant in south China in a
concentration of 0.029 ng/ml (Yang et al. 2015).
Biological effects of BPS
Although nowhere near as much information
is available about BPS as about the endocrine-
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
disrupting effects of BPS, the substitution of BPA
with BPS is raising concerns. The limited number
of studies available at the present time, dealing
with the biological interactions of BPS with the
organism, indicate that BPS is also capable of
imitating properties of hormones, interacting with
ER (Delfosse et al. 2012; Rosenmai et al. 2014;
LeFol et al. 2015), and direct binding to nuclear
ERs (Yamasaki et al. 2004) and serum albumins
(Mathew et al. 2014) has been confirmed.
Some in vitro studies have demonstrated a weaker
estrogen activity of BPS than the activity manifested
by estradiol (Kuruto-Niwa et al. 2010; Grignard et
al. 2012; Molina-Molina et al. 2013; Rochester and
Bolden 2015). By contrast, a study conducted by
Vinas and Watson (2013a, b) demonstrated the same
or higher estrogen effectiveness than estradiol, BPS
was capable of stimulating the membrane recep-
tor pathways ordinarily up-regulated by estradiol.
After exposure to BPS there are also changes in
the expression of aromatase, the key enzyme in the
synthesis of estradiol (Kinch et al. 2015).
Like in the case of BPA, the androgenic activity
of BPS was confirmed (Kitamura et al. 2005), and
subsequently its anti-androgenic activity as well
(Molina-Molina et al. 2013). ese observations in
vitro have also been confirmed by in vivo studies.
Chen et al. (2002) described acute toxicity of BPS in
Daphnia magna and at the same time also demon-
strated estrogen activity of BPS in vitro. Yamasaki et
al. (2004) documented estrogen activity of BPS in vivo
in rats with the assistance of postnatal exposure to
BPS, which in both low and high doses induced the
growth of the womb (Owens and Ashby 2002). An
in vivo study on the effect of BPS in zebrafish docu-
mented not only changes in the mass of the gonads
and plasmatic levels of estrogen and testosterone,
but also a marked disruption of reproduction. e
study of Qiu and colleagues evaluated the impact of
BPA and BPS on the reproductive neuroendocrine
system during zebrafish embryonic development,
and explored potential mechanisms of action as-
sociated with ER, thyroid hormone receptor, and
enzyme aromatase pathways. All of these pathways
were necessary to observe the full effects of BPS on
the changes in gene expression in the reproductive
neuroendocrine axis (Qiu et al. 2016). ese data
were substantiated by a decrease in egg production
and hatchability and an increasing number of embryo
malformations (Ji et al. 2013). ese observations
were later extended upon by increased time to hatch,
reduced number of sperm, increasing number of
female to male ratio, and changes in the levels of
testosterone, estradiol, and vitellogenin (Naderi et
al. 2014). In further experiments provided in cell
cultures it has been demonstrated that BPS acts
cytotoxically, genotoxically (Lee et al. 2013), and
mutagenically (Fic et al. 2013).
The reason for these negative effects may be
for example binding to serum albumins or DNA
damage and subsequent influencing of several
signal cascades anywhere within the organism
(Lee et al. 2013; Mathew et al. 2014). Exposure to
BPS disrupts cellular signalling in the apoptotic
and survival pathways (Salvesen and Walsh 2014).
Evidently, it is possible to expect the interference
of BPS in signal pro-apoptotic pathways and signal
cascades described also in gametes, leading to an
altered cell cycle and cell death (Nevoral et al. 2013;
Sedmikova et al. 2013). Further studies focused on
the mechanism of BPS action are needed for a full
understanding its negative effect on reproduction
on the gamete level and cell cycle regulation.
In respect to previous regrettable substitution,
another bisphenols, such as bisphenol F (BPF,
bis(4-hydroxyphenyl)methane; see Figure 1), do
Table 2. Bisphenol S (BPS) levels in the personal care products and environment
Sample Level of BPS References
Canned food (ng/g) 8.9–17 Vinas et al. (2010)
ermal paper (mg/g) 0.0000138–22.0 Liao et al. (2012c)
Tickets (µg/g) 0.183–5.93 Liao et al. (2012c)
Currency bills (µg/g) 0.00–6.26 Liao et al. (2012c)
Other paper product types (µg/g) 0.00–8.38 Liao et al. (2012c)
Indoor dust (µg/g) 0.34 Liao et al. (2012b)
Municipal sawage sludge (ng/g dry weight) 0.17–110.00 Song et al. (2014)
River water (ng/l) 0.29–18.99 Yang et al. (2014)
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
not seem to be a suitable alternative. In addi-
tion to BPA and BPS, BPF has been described as
endocrine disruptor as well (Perez et al. 1998).
Surprisingly, natural presence of BPF has recently
been observed in mustard and, therefore, it is
a frequent compound of foodstuff (Zoller et al.
2016). Hence, BPF regulation is ambiguous for its
chronical intake by a major part of human popula-
tion (Dietrich and Hengstler 2016).
At present we are witnessing the substitution of
BPA with BPS in a whole range of materials, and
BPS is becoming a standard component of several
products. BPS is a substance which is structurally
very similar to BPA, it shows analogous effective-
ness and mechanism of in vitro action. Biological
changes occurring in the range of typical human
exposures were documented at doses below those
used in traditional toxicology. On the basis of the
described comparisons, it is possible to expect
that BPS, like BPA, is an endocrine disruptor,
and that it may have similar targets and manner
of action in vivo and may influence physiological
processes on several levels. With regard to its
slower degradation, BPS may act for a longer time
in the organism and thus interfere with the regu-
lation of reproduction of mammals in a yet more
dangerous manner than has been demonstrated
by a range of studies in the case of BPA.
The alarming results of the first reproduction
studies on BPS have generated an acute need for a
wider and at the same time more detailed assess-
ment of the impacts of BPS, with emphasis on the
area of reproduction of mammals, which is entirely
lacking at present. Should this not materialize,
due to the increasing industrial production of BPS
caused by the need to replace BPA, unfortunately
BPS may within the foreseeable future become
just as great an environmental health risk as BPA.
There is a need for very intensive research and
subsequently also legislative measures in order
to ensure that BPS will not become another “re-
grettable substitution” with pronounced negative
impacts on the environment and on human health,
including negative impacts on reproduction.
Acknowledgement. Professor František Jílek is
greatly acknowledged for his assistance in manu-
script writing.
Akingbemi B.T., Sottas C.M., Koulova A.I., Klinefelter G.R.,
Hardy M.P. (2004): Inhibition of testicular steroidogen-
esis by the xenoestrogen bisphenol A is associated with
reduced pituitary luteinizing hormone secretion and
decreased steroidogenic enzyme gene expression in rat
Leydig cells. Endocrinology, 145, 592–603.
Aluru N., Leatherland J.F., Vijayan M.M. (2010): Bisphe-
nolA in oocytes leads to growth suppression and altered
stress performance in juvenile rainbow trout. PLoS ONE,
5, e10741.
Barrett J.R. (2013): Assessing the safety of a replacement
chemical: nongenomic activity of bisphenol S. Environ-
mental Health Perspectives, 121, 97.
Bergman A., Ryden A., Law R.J., de Boer J., Covaci A., Alaee
M. (2012): A novel abbreviation standard for organo-
bromine, organochlorine and organophosphorus flame
retardants and some characteristics of the chemicals.
Environment International, 49, 57–82.
Bergman A., Heindel J.J., Jobling S., Kidd K.A ., Zoeller R.T.
(eds) (2013): State of the Science of Endocrine Disrupting
Chemicals – 2012. World Health Organization, Geneva,
Switzerland/United Nations Environment Programme,
Nairobi, Kenya.
Birnbaum L.S. (2012): Environmental chemicals: evaluating
low-dose effects. Environmental Health Perspectives,
120, 143–144.
Blacquiere T., Smagghe G., van Gestel C.A., Mommaerts V.
(2012): Neonicotinoids in bees: a review on concentra-
tions, side-effects and risk assessment. Ecotoxicology,
21, 973–992.
Bloom M.S., vom Saal F.S., Kim D., Taylor J.A., Lamb J.D.,
Fujimoto V.Y. (2011): Serum unconjugated bisphenol A
concentrations in men may influence embryo quality
indicators during in vitro fertilization. Environmental
Toxicology and Pharmacology, 32, 319–323.
Bonefeld-Jorgensen E.C., Long M., Hofmeister M.V., Ving-
gaard A.M. (2007): Endocrine-disrupting potential of bis-
phenol A, bisphenol A dimethacrylate, 4-n-nonylphenol,
and 4-n-octylphenol in vitro: new data and a brief review.
Environmental Health Perspectives, 1, 69–76.
Borman E.D., Foster W.G., Greenacre M.K., Muir C.C., de
Catanzaro D. (2015): Stress lowers the threshold dose at
which bisphenol A disrupts blastocyst implantation, in
conjunction with decreased uterine closure and e-cadher-
in. Chemico-Biological Interactions, 237, 87–95.
Brieno-Enriquez M.A., Robles P., Camats-Tarruella N.,
Garcia-Cruz R., Roig I., Cabero L., Martinez F., Caldes
M.G. (2011): Human meiotic progression and recombina-
tion are affected by Bisphenol A exposure during in vitro
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
human oocyte development. Human Reproduction, 26,
Brieno-Enriquez M.A., Reig-Viader R., Cabero L., Toran
N., Martinez F., Roig I., Garcia Caldes M. (2012): Gene
expression is altered after bisphenol A exposure in human
fetal oocytes in vitro. Molecular Human Reproduction,
18, 171–183.
Bromer J.G., Zhou Y., Taylor M.B., Doherty L., Taylor H.S.
(2010): Bisphenol-A exposure in utero leads to epigenetic
alterations in the developmental programming of uterine
estrogen response. The FASEB Journal, 24, 2273–2280.
Bushnik T., Haines D., Levallois P., Levesque J., van Oostdam
J., Viau C. (2010): Lead and bisphenol A concentrations
in the Canadian population. Health Reports, 21, 7–18.
Calabrese E.J. (2005): Toxicological awakenings: the rebirth
of hormesis as a central pillar of toxicology. Toxicology
and Applied Pharmacology, 204, 1–8.
Calabrese E.J., Baldwin L.A. (2001): The frequency of
U-shaped dose-responses in the toxicological literature.
Toxicological Sciences, 62, 330–338.
Can A., Semiz O., Cinar O. (2005): Bisphenol-A induces cell
cycle delay and alters centrosome and spindle microtu-
bular organization in oocytes during meiosis. Molecular
Human Reproduction, 11, 389–396.
Cao X., Zhang J., Goodyer C.G., Hayward S., Cooke G.M.,
Curran I.H.A. (2012): Bisphenol A in human placental and
fetal liver tissues collected from Greater Montreal area
(Quebec) during 1998–2008. Chemosphere, 89, 505–511.
Casas M., Valvi D., Luque N., Ballesteros-Gomez A., Carsin
A.E., Fernandez M.F., Koch H.M., Mendez M.A., Sunyer
J., Rubio S., Vrijheid M. (2013): Dietary and sociodemo-
graphic determinants of bisphenol A urine concentra-
tions in pregnant women and children. Environment
International, 56, 10–18.
Chan W.K., Chan K.M. (2012): Disruption of the hypotha-
lamic–pituitary–thyroid axis in zebrafish embryo-larvae
following waterborne exposure to BDE-47, TBBPA and
BPA. Aquatic Toxicology, 108, 106–111.
Chen M.Y., Ike M., Fujita M. (2002): Acute toxicity, mu-
tagenicity, and estrogenicity of bisphenol A and other
bisphenols. Environmental Toxicology, 17, 80–86.
Chen S., Zhou D., Hsin L.Y., Kanaya N., Wong C., Yip R.,
Sakamuru S., Xia M., Yuan Y.C., Witt K., Teng C. (2014):
AroER tri-screen is a biologically relevant assay for en-
docrine disrupting chemicals modulating the activity of
aromatase and/or the estrogen receptor. Toxicological
Sciences, 139, 198–209.
Clayton R. (ed.) (2011): Endocrine Disrupters in the Environ-
ment. Foundation for Water Research, Marlow, UK, 3–22.
Coggon D. (2002): Work with pesticides and organophos-
phate sheep dips. Occupational Medicine, 52, 467–470.
Colborn T., vom Saal F.S., Soto A.M. (1993): Developmen-
tal effects of endocrine disrupting chemicals in wildlife
and humans. Environmental Health Perspectives, 101,
Corbel T., Gayrard V., Puel S., Lacroix M.Z., Berrebi A.,
Gil S., Viguie C., Toutain P.L., Picard-Hagen N. (2014):
Bidirectional placental transfer of Bisphenol A and its
main metabolite, Bisphenol A-Glucuronide, in the iso-
lated perfused human placenta. Reproductive Toxicology,
47, 51–58.
Damstra T., Barlow S., Bergman A., Kavlock R., Van Der
Kraak G. (eds) (2002): Global assessment of the state-of-the-
science of endocrine disruptors. International Programme
on Chemical Safety, World Health Organization. Available
endocrine_disruptors/en/ (accessed Aug 1, 2002).
Danzl E., Sei K., Soda S., Ike M., Fujita M. (2009): Biodeg-
radation of bisphenol A, bisphenol F and bisphenolS in
seawater. International Journal of Environmental Re-
search and Public Health, 6, 1472–1484.
Delfosse V., Grimaldi M., Pons J.L., Boulahtouf A., le Maire
A., Cavailles V., Labesse G., Bourguet W., Balaguer P.
(2012): Structural and mechanistic insights into bisphe-
nols action provide guidelines for risk assessment and
discovery of bisphenol A substitutes. Proceedings of the
National Academy of Sciences of the United States of
America, 109, 14930–14935.
Diamanti-Kandarakis E., Bourguignon J.P., Giudice L.C.
(2009): Endocrine-disrupting chemicals: an Endocrine So-
ciety scientific statement. Endocrine Reviews, 30, 293–342.
Dietrich D.R., Hengstler J.G. (2016): From bisphenol A to
bisphenol F and a ban of mustard due to chronic low-dose
exposures? Archives of Toxicology, 90, 489–491.
Dodds E.C., Lawson W. (1936): Synthetic estrogenic agents
without the phenanthrene nucleus. Nature, 137, 996.
Dolinoy D.C., Huang D., Jirtle R.L. (2007): Maternal nutrient
supplementation counteracts bisphenol A-induced DNA
hypomethylation in early development. Proceedings of
the National Academy of Sciences of the United States
of America, 104, 13056–13061.
Drastichova J., Svobodova Z., Groenland M., Dobsikova R.,
Zlabek V., Weissova D. (2005): Effect of exposure to bis-
phenol A and 17b-estradiol on the sex differentiation in ze-
brafish (Danio rerio). Acta Veterinaria Brno, 74, 287–291.
Edlow A.G., Chen M., Smith N.A., Lu C., McElrath T.F.
(2012): Fetal bisphenol A exposure: concentration of
conjugated and unconjugated bisphenol A in amniotic
fluid in the second and third trimesters. Reproductive
Toxicology, 34, 1–7.
Ehrlich S., Willliams P.L., Missmer S.A., Flaws J.A., Ye X.,
Calafat A.M., Petrozza J.C., Wright D., Hauser R. (2012):
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
Urinary bisphenol A concentrations and early reproduc-
tive health outcomes among women undergoing IVF.
Human Reproduction, 27, 3583–3592.
Ehrlich S., Calafat A.M., Humblet O., Smith T., Hauser R.
(2014): Handling of thermal receipts as a source of ex-
posure to bisphenol A. Journal of the American Medical
Association, 311, 859–860.
Eichenlaub-Ritter U., Vogt E., Cukurcam S., Sun F., Pac-
chierotti F., Parry J. (2008): Exposure of mouse oocytes
to bisphenol A causes meiotic arrest but not aneuploidy.
Mutation Research, 651, 82–92.
Commission Directive (2011): Commission Directive
2011/8/EU of 28 January 2011 amending Directive
2002/72/EC as regards the restriction of use of Bisphe-
nolA in plastic infant feeding bottles. Official Journal of
the European Union, L 26, 11–14.
EFSA (2014): Bisphenol A: EFSA consults on assessment of
risks to human health. European Food Safety Authority.
Available from (accessed Jan 17, 2014).
Fahrenkamp-Uppenbrink J. (2015): Using chemical design
to avoid regrets. Science, 347, 1213.
FDA (2011): Bisphenol A (BPA). US Food and Drug Ad-
ministration. Available from
ucm166145.htm (accessed March 24, 2011).
Fernandez M.F., Roman M., Arrebola J.P., Olea N. (2014):
Endocrine disruptors: time to act. Current Environmental
Health Reports, 1, 325–332.
Ferris J., Favetta L.A., King W.A. (2015): Bisphenol A expo-
sure during oocyte maturation in vitro results in spindle
abnormalities and chromosome misalignment in Bos
taurus. Cytogenetic and Genome Research, 145, 50–58.
Fic A., Zegura B., Sollner Dolenc M., Filipic M., Peterlin
Masic L. (2013): Mutagenicity and DNA damage of bis-
phenolA and its structural analogues in HepG2 cells. Ar-
chives of Industrial Hygiene and Toxicology, 64, 189–200.
Glausiusz J. (2014): Toxicology: the plastics puzzle. Nature,
508, 306–308.
Goodson A., Summerfield W., Cooper I. (2002): Survey of
bisphenol A and bisphenol F in canned foods. Food Ad-
ditives and Contaminants, 19, 796–802.
Grasselli F., Baratta L., Baioni L., Bussolati S., Ramoni R.,
Grolli S., Basini G. (2010): Bisphenol A disrupts granulosa
cell function. Domestic Animal Endocrinology, 39, 34–39.
Grignard E., Lapenna S., Bremer S. (2012): Weak estrogenic
transcriptional activities of Bisphenol A and BisphenolS.
Toxicology in Vitro, 26, 727–731.
Helies-Toussaint C., Peyre L., Costanzo C., Chagnon M.C.,
Rahmani R. (2014): Is bisphenol S a safe substitute for bis-
phenol A in terms of metabolic function? An in vitro study.
Toxicology and Applied Pharmacology, 280, 224–235.
Howard G.J. (2014): Chemical alternatives assessment: the
case of flame retardants. Chemosphere, 116, 112–117.
Huang Y.Q., Wong C.K.C., Zheng J.S., Bouwman H., Barra
R., Wahlstrom B., Wong M.H. (2012): Bisphenol A (BPA)
in China: a review of sources, environmental levels, and
potential human health impacts. Environment Interna-
tional, 42, 91–99.
Hubbs A.F., Cumpston A.M., Goldsmith W.T., Battelli L.A.,
Kashon M.L., Jackson M.C., Frazer D.G., Fedan J.S., Gora-
vanahally M.P., Castranova V., Kreiss K., Willard P.A.,
Friend S., Schwegler-Berry D., Fluharty K.L., Sriram K.
(2012): Respiratory and olfactory cytotoxicity of inhaled
2,3-pentanedione in Sprague-Dawley rats. The American
Journal of Pathology, 181, 829–844.
Hunt P.A., Koehler K.E., Susiarjo M., Hodges C.A ., Ilagan A.,
Voigt R.C., Thomas S., Thomas B.F., Hassold T.J. (2003):
Bisphenol A exposure causes meiotic aneuploidy in the
female mouse. Current Biology, 13, 546–553.
Hunt P.A., Lawson C., Gieske M., Murdoch B., Smith H., Marre
A., VandeVoort C.A. (2012): Bisphenol A alters early ooge-
nesis and follicle formation in the fetal ovary of the rhesus
monkey. Proceedings of the National Academy of Sciences
of the United States of America, 109, 17525–17530.
Ike M., Chen M.Y., Danzl E., Sei K., Fujita M. (2006): Bio-
degradation of a variety of bisphenols under aerobic and
anaerobic conditions. Water Science and Technology,
53, 153–160.
Ikezuki Y., Tsutsumi O., Takai Y., Kamei Y., Taketani Y.
(2002): Determination of bisphenol A concentrations in
human biological fluids reveals significant early prenatal
exposure. Human Reproduction, 17, 2839–2841.
Jasarevic E., Sieli P.T., Twellman E.E., Welsh T.H., Schachtman
T.R., Roberts R.M. (2011): Disruption of adult expression
of sexually selected traits by developmental exposure to
bisphenol A. Proceedings of the National Academy of Sci-
ences of the United States of America, 108, 11715–11720.
Jenkins S., Wang J., Eltoum I., Desmond R., Lamartinie-
re C.A. (2011): Chronic oral exposure to bisphenol A
results in a nonmonotonic dose response in mammary
carcinogenesis and metastasis in MMTV-erbB2 mice.
Environmental Health Perspectives, 119, 1604–1609.
Ji K., Hong S., Kho Y., Choi K. (2013): Effects of bisphenolS
exposure on endocrine functions and reproduction of
zebrafish. Environmental Science and Technology, 47,
Jimenez-Diaz I., Zafra-Gomez A., Ballesteros O., Navea
N., Navalon A., Fernandez M.F., Olea N., Vilchez J.L.
(2010): Determination of Bisphenol A and its chlori-
nated derivatives in placental tissue samples by liquid
chromatography-tandem mass spectrometry. Journal of
Chromatography B, 878, 3363–3369.
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
Joskow R., Barr D.B., Barr J.R., Calafat A.M., Needham L.L.,
Rubin C. (2006): Exposure to bisphenol A from bis-glyci-
dyl dimethacrylate-based dental sealants. The Journal of
the American Dental Association, 137, 353–362.
Khan D., Ahmed S.A. (2015): Epigenetic regulation of
non-lymphoid cells by Bisphenol-A, a model endocrine
disrupter: potential implications for immunoregulation.
Frontiers in Endocrinology, 6, 91.
Kinch C.D., Ibhazehiebo K., Jeong J.H., Habibi H.R., Kur-
rasch D.M. (2015): Low-dose exposure to bisphenol A and
replacement bisphenol S induces precocious hypothala-
mic neurogenesis in embryonic zebrafish. Proceedings of
the National Academy of Sciences of the United States of
America, 112, 1475–1480.
Kitamura S., Suzuki T., Sanoh S., Kohta R., Jinno N., Sugi-
hara K., Yoshihara S., Fujimoto N., Watanabe H., Ohta S.
(2005): Comparative study of the endocrine-disrupting
activity of bisphenol A and 19 related compounds. The
Journal of Toxicological Sciences, 84, 249–259.
Kloas W., Lutz I., Einspanier R. (1999): Amphibians as a mo-
del to study endocrine disruptors: II. Estrogenic activity
of environmental chemicals in vitro and in vivo. Science
of the Total Environment, 225, 59–68.
Knez J., Kranvogl R., Breznik B.P., Voncina E., Vlaisavljevic
V. (2014): Are urinary bisphenol A levels in men related
to semen quality and embryo development after medi-
cally assisted reproduction? Fertility and Sterility, 101,
Krishnan A., Stathis P., Permuth S., Tokes L., Feldman D.
(1993): Bisphenol-A – an estrogenic substance is released
from polycarbonate flasks during autoclaving. Endocri-
nology, 132, 2279–2286.
Kundakovic M., Gudsnuk K., Franks B., Madrid J., Miller
R.L., Perera F.P., Champagne F.A. (2013): Sex-specific
epigenetic disruption and behavioral changes following
low-dose in utero bisphenol A exposure. Proceedings of
the National Academy of Sciences of the United States
of America, 110, 9956–9961.
Kurosawa T., Hiroi H., Tsutsumi O., Ishikawa T., Osuga
Y., Fujiwara T., Inoue S., Muramatsu M., Momoeda M.,
Taketani Y. (2002): The activity of bisphenol A depends
on both the estrogen receptor subtype and the cell type.
Endocrine Journal, 49, 465–471.
Kuruto-Niwa R., Tateoka Y., Usuki Y., Nozawa R. (2007):
Measurement of bisphenol A concentrations in human
colostrum. Chemosphere, 66, 1160–1164.
Kuruto-Niwa R., Nozawa R., Miyakoshi T., Shiozawa T.,
Terao Y. (2010): Estrogenic activity of alkylphenols, bis-
phenol S, and their chlorinated derivatives using a GFP
expression system. Environmental Toxicology and Phar-
macology, 19, 121–130.
Kwintkiewicz J., Nishi Y., Yanase T., Giudice L.C. (2010):
Peroxisome proliferator-activated receptor-γ mediates
bisphenol A inhibition of FSH-stimulated IGF-1, aro-
matase, and estradiol in human granulosa cells. Environ-
mental Health Perspectives, 118, 400–406.
Lahnsteiner F., Berger B., Kletz M., Weismann T. (2005):
Effect of bisphenol A on maturation and quality of semen
and eggs in the brown trout, Salmo trutta f. fario. Aquatic
Toxicology, 75, 213–224.
Landrigan P.J., Kimmel C.A., Eskenazi B. (2004): Children’s
health and the environment: public health issues and
challenges for risk assessment. Environmental Health
Perspectives, 112, 257–265.
Lang I.A., Galloway T.S., Scarlett A., Henley W.E., De-
pledge M., Wallace R.B. (2008): Association of urinary
bisphenolA concentration with medical disorders and
laboratory abnormalities in adults. The Journal of the
American Medical Association, 300, 1303–1310.
Lassen T.H., Frederiksen H., Jensen T.K., Petersen J.H.,
Joensen U.N., Main K.M., Andersson A.M. (2014): Uri-
nary bisphenol A levels in young men: association with
reproductive hormones and semen quality. Environmen-
tal Health Perspectives, 122, 478–484.
Lee S., Liu X., Takeda S., Choi K. (2013): Genotoxic poten-
tials and related mechanisms of bisphenol A and other
bisphenol compounds: a comparison study employing
chicken DT40 cells. Chemosphere, 93, 434–440.
Le Fol V., Ait-Aissa S., Cabaton N., Dolo L., Grimaldi M.,
Balaguer P., Perdu E., Debrauwer L., Brion F., Zalko D.
(2015): Cell-specific biotransformation of benzophe-
none-2 and bisphenol-S in zebrafish and human in vitro
models used for toxicity and estrogenicity screening.
Environmental Science and Technology, 49, 3860–3868.
Li D.K., Zhou Z., Miao M., He Y., Wang J., Ferber J., Her-
rinton L.J., Gao E., Yuan W. (2011): Urine bisphenol-A
(BPA) level in relation to semen quality. Fertility and
Sterility, 95, 625–630.
Liao C., Liu F., Alomirah H., Loi V.D., Mohd M.A., Moon
H.B. (2012a): Bisphenol S in urine from the United States
and seven Asian countries: occurrence and human ex-
posures. Environmental Science and Technology, 46,
Liao C., Liu F., Guo Y., Moon H.B., Nakata H., Wu Q. (2012b):
Occurrence of eight bisphenol analogues in indoor dust
from the United States and several Asian countries: im-
plications for human exposure. Environmental Science
and Technology, 46, 9138–9145.
Liao C., Liu F., Kannan K. (2012c): Bisphenol S, a new bis-
phenol analogue, in paper products and currency bills and
its association with bisphenol A residues. Environmental
Science and Technology, 46, 6515–6522.
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
Maamar M., Lesne L., Desdoits-Lethimonier C., Coiffec I.,
Lassurguere J., Lavoue V., Deceuninck Y., Antignac J.P.,
Le Bizec B., Perdu E., Zalko D., Pineau C., Chevrier C.,
Dejucq-Rainsford N., Mazaud-Guittot S., Jegou B. (2015):
An investigation of the endocrine-disruptive effects of
bisphenol A in human and rat fetal testes. PLoS ONE,
10, e0117226.
Machtinger R., Orvieto R. (2014): Bisphenol A, oocyte
maturation, implantation, and IVF outcome: review of
animal and human data. Reproductive BioMedicine On-
line, 29, 404–410.
Machtinger R., Combelles C.M., Missmer S.A ., Correia K.F.,
Williams P., Hauser R., Racowsky C. (2013): Bisphenol-A
and human oocyte maturation in vitro. Human Reproduc-
tion, 28, 2735–2745.
Manfo F.P., Jubendradass R., Nantia E.A., Moundipa P.F.,
Mathur P.P. (2014): Adverse effects of bisphenol A on
male reproductive function. Reviews of Environmental
Contamination and Toxicology, 228, 57–82.
Manikkam M., Tracey R., Guerrero-Bosagna C., Skinner
M.K. (2013): Plastics derived endocrine disruptors (BPA,
DEHP and DBP) induce epigenetic transgenerational
inheritance of obesity, reproductive disease and sperm
epimutations. PLoS ONE, 8, e55387.
Mathew M., Sreedhanya S., Manoj P., Aravindakumar C.T.,
Aravind U.K. (2014): Exploring the interaction of bisphe-
nol-S with serum albumins: a better or worse alternative
for bisphenol A? The Journal of Physical Chemistry B,
118, 3832–3843.
Matsushima A., Kakuta Y., Teramoto T., Koshiba T., Liu
X., Okada H., Tokunaga T., Kawabata S., Kimura M.,
Shimohigashi Y. (2007): Structural evidence for endocrine
disruptor bisphenol A binding to human nuclear receptor
ERR gamma. The Journal of Biochemistry, 142, 517–524.
McBride W.G. (1961): Thalidomide and congenital abnor-
malities. The Lancet, 278, 1358.
Mendiola J., Jorgensen N., Andersson A.M., Calafat A.M.,
Ye X., Redmon J.B. (2010): Are environmental levels of
bisphenol A associated with reproductive function in
fertile men? Environmental Health Perspectives, 118,
Mendonca K., Hauser R., Calafat A.M., Arbuckle T.E., Duty
S.M. (2014): Bisphenol A concentrations in maternal
breast milk and infant urine. International Archives of
Occupational and Environmental Health, 87, 13–20.
Mendoza-Rodriguez C.A., Garcia-Guzman M., Baranda-
Avila N., Morimoto S., Perrot-Applanat M., Cerbon M.
(2011): Administration of bisphenol A to dams during
perinatal period modifies molecular and morphological
reproductive parameters of the offspring. Reproductive
Toxicology, 31, 177–183.
Mileva G., Baker S.L., Konkle A.T., Bielajew C. (2014):
Bisphenol-A: epigenetic reprogramming and effects on
reproduction and behaviour. International Journal of En-
vironmental Research and Public Health, 11, 7537–7561.
Miyamoto K.I., Kotake M. (2005): Estimation of daily bis-
phenol A intake of Japanese individuals with emphasis on
uncertainty and variability. Environmental Sciences: an
International Journal of Environmental Physiology and
Toxicology, 13, 15–29.
Mok-Lin E., Ehrlich S., Williams P.L., Petrozza J., Wright
D.L., Calafat A.M., Ye X., Hauser R. (2010): Urinary bis-
phenol A concentrations and ovarian response among
women undergoing IVF. International Journal of Androl-
ogy, 33, 385–393.
Molina-Molina J.M., Amaya E., Grimaldi M., Saenz J.M.,
Real M., Fernandez M.F., Balaguer P., Olea N. (2013): In
vitro study on the agonistic and antagonistic activities
of bisphenol-S and other bisphenol-A congeners and
derivatives via nuclear receptors. Toxicology and Applied
Pharmacology, 272, 127–136.
Naderi M., Wong M.Y., Gholami F. (2014): Developmental
exposure of zebrafish (Danio rerio) to bisphenol-S impairs
subsequent reproduction potential and hormonal balance
in adults. Aquatic Toxicology, 148, 195–203.
Nakagawa Y., Tayama S. (2000): Metabolism and cytotox-
icity of bisphenol A and other bisphenols in isolated rat
hepatocytes. Archives of Toxicology, 74, 99–105.
Navarro V.M., Sanchez-Garrido M.A., Castellano J.M., Roa
J., Garcia-Galiano D., Pineda R., Tena-Sempere M. (2009):
Persistent impairment of hypothalamic KiSS-1 system after
exposures to estrogenic compounds at critical periods of
brain sex differentiation. Endocrinology, 150, 2359–2367.
Nevoral J., Krejcova T., Petr J., Melicharova P., Vyskocilova
A., Dvorakova M., Sedmikova M. (2013): The role of nitric
oxide synthase isoforms in aged porcine oocytes. Czech
Journal of Animal Science, 58, 453–459.
Newbold R.R ., Jefferson W.N., Padilla-Banks E. (2009): Pre-
natal exposure to bisphenol A at environmentally relevant
doses adversely affects the murine female reproductive
tract later in life. Environmental Health Perspectives,
117, 879–885.
N’Tumba-Byn T., Moison D., Lacroix M., Lecureuil C.,
Lesage L. (2012): Differential effects of bisphenol A and
diethylstilbestrol on human, rat and mouse fetal Leydig
cell function. PLoS ONE, 7, e51579.
Owens J.W., Ashby J. (2002): Critical review and evalua-
tion of the uterotrophic bioassay for the identification
of possible estrogen agonists and antagonists: in support
of the validation of the OECD uterotrophic protocols for
the laboratory rodent. Critical Reviews in Toxicology,
32, 445–520.
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
Pacchierotti F., Ranaldi R., Eichenlaub-Ritter U., Attia S.,
Adler I.D. (2008): Evaluation of aneugenic effects of bis-
phenol A in static and germ cells of the mouse. Mutation
Research, 651, 64–70.
Patisaul H.B., Todd K.L., Mickens J.A., Adewale H.B. (2009):
Impact of neonatal exposure to the ERα agonist PPT, bis-
phenol-A or phytoestrogens on hypothalamic kisspeptin
fiber density in male and female rats. Neurotoxicology,
30, 350–357.
Peretz J., Vrooman L., Ricke W.A., Hunt P.A., Ehrlich S.,
Hauser R., Padmanabhan V., Taylor H.S., Swan S.H., Van-
deVoort C.A., Flaws J.A. (2014): Bisphenol A and reproduc-
tive health: update of experimental and human evidence.
Environmental Health Perspectives, 122, 775–786.
Perez P., Pulgar R ., Olea-Serrano F., Villalobos M., Rivas A.,
Metzler M., Pedraza V., Olea N. (1998): The estrogenic-
ity of bisphenol A-related diphenylalkanes with various
substituents at the central carbon and the hydroxy groups.
Environmental Health Perspectives, 106, 167–174.
Petr J., Chmelikova E., Zalmanova T., Tumova L., Kheilova
K., Kucerova-Chrpova V., Jilek F. (2013): Pyrethroids cy-
permethrin, deltamethrin and fenvalerate have different
effects on in vitro maturation of pig oocytes at different
stages of growth. Animal, 7, 134–142.
Peyre L., Rouimi P., de Sousa G., Helies-Toussaint C., Carre
B., Barcellini S., Chagnon M.C., Rahmani R. (2014): Com-
parative study of bisphenol A and its analogue bisphe-
nolS on human hepatic cells: a focus on their potential
involvement in nonalcoholic fatty liver disease. Food and
Chemical Toxicology, 70, 9–18.
Qiu W., Zhao Y., Yang M., Farajzadeh M., Pan C., Wayne
N.L. (2016): Actions of bisphenol A and bisphenol S on
the reproductive neuroendocrine system during early
development in zebrafish. Endocrinology, 157, 636–647.
Rama Raju G.A., Prakash G.J., Krishna K.M., Madan K.
(2007): Meiotic spindle and zona pellucida characteristics
as predictors of embryonic development: a preliminary
study using PolScope imaging. Reproductive Biomedicine
Online, 14, 166–174.
Richter C., Birnbaum L.S., Farabollini F., Newbold R.R., Ru-
bin B.S., Talsness C.E., Vandenbergh J.G., Walser-Kuntz
D.R., vom Saal F.S. (2007): In vivo effects of bisphenol A
in laboratory rodent studies. Reproductive Toxicology,
24, 199–224.
Rochester J.R. (2013): Bisphenol A and human health: a
review of the literature. Reproductive Toxicology, 42,
Rochester J.R., Bolden A.L. (2015): Bisphenol S and F: a
systematic review and comparison of the hormonal ac-
tivity of bisphenol A substitutes. Environmental Health
Perspectives, 123, 643–650.
Rosenmai A.K., Dybdahl M., Pedersen M., van Vugt-Lus-
senburg B.M.A., Wedebye E.B., Taxvig C., Vinggaard
A.M. (2014): Are structural analogues to bisphenol a safe
alternatives? Toxicological Sciences, 139, 3–47.
Salgueiro-Gonzalez N., Turnes-Carou I., Vinas-Dieguez
L., Muniategui-Lorenzo S., Lopez-Mahia P., Prada-Rod-
riguez D. (2015): Occurrence of endocrine disrupting
compounds in five estuaries of the northwest coast of
Spain: ecological and human health impact. Chemosphe-
re, 131, 241–247.
Salvesen G.S., Walsh C.M. (2014): Functions of caspase 8:
the identified and the mysterious. Seminars in Immunol-
ogy, 26, 246–252.
Santamaria C., Durando M., Munoz de Toro M., Luque
E.H., Rodriguez H.A. (2016): Ovarian dysfunction in adult
female rat offspring born to mothers perinatally exposed
to low doses of bisphenol A. The Journal of Steroid Bio-
chemistry and Molecular Biology, 158, 220–230.
Schonfelder G., Wittfoht W., Hopp H., Talsness C.E., Paul
M., Chahoud I. (2002): Parent bisphenol A accumulation
in the human maternal–fetal–placental unit. Environ-
mental Health Perspectives, 110, 703–707.
Sedmikova M., Petr J., Dorflerova A., Nevoral J., Novotna
B. (2013): Inhibition of c-Jun N-terminal kinase (JNK)
suppresses porcine oocyte aging in vitro. Czech Journal
of Animal Science, 58, 535–545.
Sohoni P.C.R.T., Hurd K., Caunter J., Hetheridge M., Wil-
liams T., Woods C., Evans M., Toy R., Gargas M., Sumpter
J.P. (2001): Reproductive effects of long-term exposure to
bisphenol A in the fathead minnow (Pimephales promelas).
Environmental Science and Technology, 35, 2917–2925.
Song S., Song M., Zeng L., Wang T., Liu R., Ruan T. (2014):
Occurrence and profiles of bisphenol analogues in mu-
nicipal sewage sludge in China. Environmental Pollution,
186, 14–19.
Sui Y., Ai N., Park S., Rios-Pilier J., Perkins J.T., Welsh W.J.,
Zhou C. (2012): Bisphenol A and its analogues activate
human pregnane X receptor. Environmental Health Per-
spectives, 120, 399–405.
Susiarjo M., Hassold T.J., Freeman E., Hunt P.A. (2007):
Bisphenol A exposure in utero disrupts early oogenesis
in the mouse. PLoS Genetics, 3, e5.
Susiarjo M., Sasson I., Mesaros C., Bartolomei M.S. (2013):
Bisphenol A exposure disrupts genomic imprinting in the
mouse. PLoS Genetics, 9, e1003401.
Susiarjo M., Xin F., Bansal A., Stefaniak M., Li C., Simmons
R.A., Bartolomei M.S. (2015): Bisphenol A exposure dis-
rupts metabolic health across multiple generations in the
mouse. Endocrinology, 156, 2049–2058.
Suzuki A., Sugihara A., Uchida K., Sato T., Ohta Y., Katsu Y.,
Watanabe H., Iguchi T. (2002): Developmental effects of
Review Czech J. Anim. Sci., 61, 2016 (10): 433–449
doi: 10.17221/81/2015-CJAS
perinatal exposure to bisphenol-A and diethylstilbestrol
on reproductive organs in female mice. Reproductive
Toxicology, 16, 107–116.
Takayanagi S., Tokunaga T., Liu X., Okada H., Matsushima
A., Shimohigashi Y. (2006): Endocrine disruptor bisphe-
nol A strongly binds to human estrogen-related recep-
tor gamma (ERRgamma) with high constitutive activity.
Toxicology Letters, 167, 95–105.
Teng C., Goodwin B., Shockley K., Xia M., Huang R., Nor-
ris J., Merrick B.A., Jetten A.M., Austin C.P., Tice R.R.
(2013): Bisphenol A affects androgen receptor function
via multiple mechanisms. Chemico-Biological Interac-
tions, 203, 556–564.
Tomari H., Honjou K., Nagata Y., Horiuchi T. (2011): Re-
lationship between meiotic spindle characteristics in
human oocytes and the timing of the first zygotic cleav-
age after intracytoplasmic sperm injection. Journal of
Assisted Reproduction and Genetics, 28, 1099–1104.
Trapphoff T., Heiligentag M., Hajj N.E., Haaf T., Eichenlaub-
Ritter U. (2013): Chronic exposure to a low concentra-
tion of bisphenol A during follicle culture affects the
epigenetic status of germinal vesicles and metaphase II
oocytes. Fertility and Sterility, 100, 1758–1767.
Trasande L., Attina T.M., Blustein J. (2012): Association
between urinary bisphenol A concentration and obesity
prevalence in children and adolescents. The Journal of the
American Medical Association, 308, 1113–1121.
Vandenberg L.N. (2014): Non-monotonic dose responses in
studies of endocrine disrupting chemicals: bisphenolA
as a case study. Dose-Response, 12, 259–276.
Vandenberg L.N., Maffini M.V., Sonnenschein C., Rubin
B.S., Soto A.M. (2009): Bisphenol-A and the great divide:
a review of controversies in the field of endocrine disrup-
tion. Endocrine Reviews, 30, 75–95.
Vanderberg L.N., Colborn T., Hazes T.B., Heindel J.J.,
Jacobs D.R., Lee D.H., Shioda T., Soto A.M., vom Saal
F.S., Welshons W.V., Zoeller R.T., Myers J.P. (2012): Hor-
mones and endocrine-disrupting chemicals: low-dose
effects and nonmonotonic dose response. Endocrine
Reviews, 33, 378–455.
Vandenberg L.N., Colborn T., Hayes T.B., Heindel J.J., Jacobs
D.R., Lee D.H., Myers J.P., Shioda T., Soto A.M., Vom
Saal F.S., Welshons W.V., Zoeller R.T. (2013). Regulatory
decisions on endocrine disrupting chemicals should be
based on the principles of endocrinology. Reproductive
Toxicology, 38, 1–15.
Vinas P., Campillo N., Martinez-Castillo N., Hernandez-
Cordoba M. (2010): Comparison of two derivatization-
based methods for solid-phase microextraction–gas
chromatography–mass spectrometric determination of
bisphenol A, bisphenol S and biphenol migrated from
food cans. Analytical and Bioanalytical Chemistry, 397,
Vinas R., Watson C.S. (2013a): Bisphenol S disrupts es-
tradiol-induced nongenomic signaling in a rat pituitary
cell line: effects on cell functions. Environmental Health
Perspectives, 121, 352–358.
Vinas R., Watson C.S. (2013b): Mixtures of xenoestrogens
disrupt estradiol-induced non-genomic signaling and
downstream functions in pituitary cells. Environmental
Health, 12, 26.
Vinggaard A.M., Niemela J., Wedebye E.B., Jensen G.E.
(2008): Screening of 397 chemicals and development of
a quantitative structure–activity relationship model for
androgen receptor antagonism. Chemical Research in
Toxicology, 21, 813–823.
Vitku J., Sosvorova L., Chlupacova T., Hampl R., Hill M.,
Sobotka V., Heracek J., Bicikova M., Starka L. (2015):
Differences in bisphenol A and estrogen levels in the
plasma and seminal plasma of men with different degrees
of infertility. Physiological Research, 64, 303–311.
vom Saal F.S., Hughes C. (2005): An extensive new literature
concerning low-dose effects of bisphenol A shows the
need for a new risk assessment. Environmental Health
Perspectives, 113, 926–933.
vom Saal F.S., Welshons W.V. (2005): Large effects from
small exposures. II. The importance of positive controls
in low-dose research on bisphenol A. Environmental
Research, 100, 50–76.
vom Saal F.S., Akingbemi T.B., Belcher S.M., Birnbaum
L.S., Crain D.A., Eriksen M., Farabollini F., Guillette Jr.
L.J., Hauser R., Heindel J.J., Ho S.-M., Hunt P.A., Iguchi
T., Jobling S., Kanno J., Keri R.A., Knudsen K.E., Laufer
H., LeBlanc G.A., Marcus M., McLachlan J.A ., Myers J.P.,
Nadal A., Newbold R .R., Olea N., Prins G.S., Richter C.A.,
Rubin B.S., Sonnenschein C., Soto A.M., Talsness C.E.,
Vandenbergh J.G., Vandenberg L.N., Walser-Kuntz D.R.,
Watson C.S., Welshons W.V., Wetherill Y., Zoeller R.T.
(2007): Chapel Hill bisphenol A expert panel consensus
statement: integration of mechanisms, effects in animals
and potential to impact human health at current levels of
exposure. Reproductive Toxicology, 24, 131–138.
Vrooman L.A., Oatley J.M., Griswold J.E., Hassold T.J., Hunt
P.A. (2015): Estrogenic exposure alters the spermato-
gonial stem cells in the developing testis, permanently
reducing crossover levels in the adult. PLoS Genetics,
11, e1004949.
Wang F., Hua J., Chen M., Xia Y., Zhang Q., Zhao R., Zhou
W., Zhang Z., Wang B. (2012): High urinary bisphenolA
concentrations in workers and possible laboratory ab-
normalities. Occupational and Environmental Medicine,
69, 679–684.
Czech J. Anim. Sci., 61, 2016 (10): 433–449 Review
doi: 10.17221/81/2015-CJAS
Wang T., Lu J., Xu M., Xu Y., Li M., Liu Y., Ning G. (2013):
Urinary bisphenol A concentration and thyroid function
in Chinese adults. Epidemiology, 24, 295–302.
Warita K., Mitsuhashi T., Ohta K., Suzuki S., Hoshi N., Miki
T., Takeuchi Y. (2013): Gene expression of epigenetic
regulatory factors related to primary silencing mechanism
is less susceptible to lower doses of bisphenol A in em-
bryonic hypothalamic cells. The Journal of Toxicological
Sciences, 38, 285–289.
Wetherill Y.B., Akingbemi B.T., Kanno J., McLachlan J.A.,
Nadal A., Sonnenschein C., Watson C.S., Zoeller R.T.,
Belcher S.M. (2007): In vitro molecular mechanisms of bi-
sphenol A action. Reproductive Toxicology, 24, 178–198.
Wilson M.E., Sengoku T. (2013): Developmental regulation
of neuronal genes by DNA methylation: environmental
influences. International Journal of Developmental Neu-
roscience, 31, 448–451.
Xi W., Lee C.K.F., Yeung W.S.B., Giesy J.P., Wong M.H.,
Zhang X., Wong C.K. (2011): Effect of perinatal and post-
natal bisphenol A exposure to the regulatory circuits at
the hypothalamus–pituitary–gonadal axis of CD-1 mice.
Reproductive Toxicology, 31, 409–417.
Xiao S., Diao H., Smith M.A., Song X., Ye X. (2011): Preim-
plantation exposure to bisphenol A (BPA) affects embryo
transport, preimplantation embryo development, and
uterine receptivity in mice. Reproductive Toxicology,
32, 434–441.
Xu L.C., Sun H., Chen J.F., Bian Q., Qian J., Song L., Wang
X.R. (2005): Evaluation of androgen receptor transcrip-
tional activities of bisphenol A, octylphenol and nonyl-
phenol in vitro. Toxicology, 216, 197–203.
Yamada H., Furuta I., Kato E.H., Kataoka S., Usuki Y., Kob-
ashi G., Fujimoto S. (2002): Maternal serum and amniotic
fluid bisphenol A concentrations in the early second
trimester. Reproductive Toxicology, 16, 735–739.
Yamasaki K., Noda S., Imatanaka N., Yakabe Y. (2004): Com-
parative study of the uterotrophic potency of 14 chemicals
in a uterotrophic assay and their receptor-binding affinity.
Toxicology Letters, 146, 111–120.
Yang O., Kim H.L., Weon J.I., Seo Y.R. (2015): Endocrine-
disrupting chemicals: review of toxicological mechanisms
using molecular pathway analysis. Journal of Cancer
Prevention, 20, 12–24.
Yang Y., Lu L., Zhang J., Yang Y., Wu Y., Shao B. (2014):
Simultaneous determination of seven bisphenols in envi-
ronmental water and solid samples by liquid chromatog-
raphy–electrospray tandem mass spectrometry. Journal
of Chromatography A, 1328, 26–34.
Ye J., Coleman J., Hunter M.G., Craigon J., Campbell K.H.S.,
Luck M.R. (2007): Physiological temperature variants and
culture media modify meiotic progression and develop-
mental potential of pig oocytes in vitro. Reproduction,
133, 877–886.
Zhang H.Q., Zhang X.F., Zhang L.J., Chao H.H., Pan B., Feng
Y.M., Shen W., Li L., Sun X.F. (2012): Fetal exposure to
bisphenol A affects the primordial follicle formation by
inhibiting the meiotic progression of oocytes. Molecular
Biology Reports, 39, 5651–5657.
Zhang T., Qin X.S., Zhou Y., Zhang X.F., Wang L.Q., Felici
M., Chen H., Qin G.Q., Shen W. (2014): Di-(2-ethylhexyl)
phthalate and bisphenol A exposure impairs mouse primor-
dial follicle assembly in vitro. Experimental and Molecular
Mutagenesis, 55, 343–353.
Zhang X., Chang H., Wiseman S., He Y., Higley E., Jones
P., Wong C.K.C., Al-Khedhairy A., Giesy J.P., Hecker M.
(2011): Bisphenol A disrupts steroidogenesis in human
H295R cells. Toxicological Sciences, 121, 320–327.
Zimmerman J.B., Anastas P.T. (2015): Toward designing
safer chemicals. Science, 347, 215.
Ziv-Gal A., Wang W., Zhou C., Flaws J.A. (2015): The ef-
fects of in utero bisphenol A exposure on reproductive
capacity in several generations of mice. Toxicology and
Applied Pharmacology, 284, 354–362.
Zoeller R.T., Brown T.R., Doan L.L., Gore A.C., Skakkebaek
N.E., Soto A.M., Woodruff T.J., vom Saal F.S. (2012):
Endocrine-disrupting chemicals and public health pro-
tection: a statement of principles from the Endocrine
Society. Endocrinology, 153, 4097–4110.
Zoller O., Bruschweiler B.J., Magnin R., Reinhard H., Rhyn
P., Rupp H., Zelter S., Felleisen R. (2016): Natural oc-
currence of bisphenol F in mustard. Food Additives and
Contaminants: Part A, 33, 137–146.
Received: 2015–10–31
Accepted after corrections: 2016–05–03
Corresponding Author
Ing. Tereza Žalmanová, Ph.D., Czech University of Life Sciences Prague, Faculty of Agrobiology, Food and Natural
Resources, Department of Veterinary Sciences, Kamýcká 129, 165 21 Prague 6-Suchdol, Czech Republic
Phone: +420 224 382 942, e-mail:
... Bisphenol S is also used in different food compounds like canned eating, items related to livestock, and sea foods, etc. Bisphenol S accumulates into surrounding, and its toxicity in the laboratory and inside the model organisms has been evaluated in many researches (Qiu et al. 2019). In case of DNA, bisphenol interacts with groove of DNA and therefore it also known as "groove binder" (Qing et al. 2014 pathways that are involved in the exposure of bisphenol S are skin, ingestion of dust, and manifestation through food (Zalmanova et al. 2016). ...
Full-text available
Bisphenol S (BPS) is an analog of bisphenol A, which is used as substitute of BPA in many products like airport luggage tags, baby bottles, plastics, and epoxy resins etc. Bisphenol S can cause toxic effects in different organisms, i.e., mice, rat, zebrafish, and C.elegans, etc. Bisphenol S is also known as “endocrine disruptor” due to its ability to mimic the endocrine receptors. So, the aim of this study was to evaluate the cytotoxic and genotoxic effects of bisphenol S on meristematic cells present in onion root tips through Allium cepa (A.cepa) and comet tests. Root growth inhibition was evaluated by root growth inhibition assay. Mitotic index (MI) and chromosomal aberrations (CAs) were assessed by A.cepa assay. DNA damage was evaluated by comet assay. Root growth of A.cepa was inhibited due to bisphenol S. LC50 value calculated by root growth inhibition assay for bisphenol S was (2.6±0.63, 50 μg/ml). Mitotic index was reduced, and chromosomal aberrations were observed, i.e., stickiness, polyploidy, and disturbed ana-telophase in anaphase and telophase stages of mitosis. In case of comet assay, DNA damage was increased in statistically significant manner (p ≤ 0.05). It was concluded that bisphenol S constitutes cytotoxic and genotoxic effects on A. cepa root meristematic cells. Moreover, it is suggested to explore more toxicity studies of bisphenol S at molecular level.
... In the past ten years, many "BPA free" products have been considered as safety products and sold in the markets. However, BPA analogs, in particular BPS and BPF, also can be found in most of the "BPA free" products [15][16][17]. BPs are featured with a scaffold consisting of two phenol rings, functionalized with different groups, linked by a variously substituted carbon atom [3]. The shared core scaffold is believed directly responsible of the endocrine activity. ...
Full-text available
A reliable and affordable QuEChERS (quick, easy, cheap, effective, rugged, and safe) methodology in combination with ultra-high performance liquid chromatography-tandem mass spectrometry (UPLC–MS/MS) was successfully developed and validated for the determination of eight bisphenols (BPs) residues containing in meats (chicken, duck, beef, pork, fish, shrimp, and mutton). A novel QuEChERS method optimization was carried out in terms of process efficiency (PE), matrix effect (ME), and extraction recovery (RE). After a simple vortex extraction of the samples with acetonitrile, 1 g sodium acetate was used for salting out (NaAC), and 100 mg primary secondary amine (PSA) purifying reagents were used for purification. The properties of the sorbents were assessed by the obtained parameters, such as matrix effect (ME), linearity, sensitivity, accuracy, and precision. Under the optimal conditions, BPs were well separated on an ACQUITY UPLC BEH ® C18 column in 8 min by gradient elution, and exhibited a good linear relationship (R² > 0.9988) in the linear range. Moreover, the limits of detection (LODs) and the limits of quantification (LOQs) were located in the range of 0.01– 0.11 μg/kg and 0.03 – 0.37 μg/kg, respectively. The developed method was satisfactory in terms of accuracy (relative recoveries: 76.1% – 113.7%) and precision (relative standard deviations below 10.3%). Finally, the developed method was successfully employed to identify and quantify BPs residues in 28 real meat samples. The proposed QuEChERS-UPLC–MS/MS method is simple, high efficiency, cost-effective, practical, and susceptible to being implemented in routine laboratories to quickly detect the BPs in meats (chicken, duck, beef, pork, fish, shrimp, and mutton). In this sense, the method is useful for obtaining BPs residue data to evaluate the contamination status of BPs in meat food and provide scientific support for scientific supervision.
... These chemicals can migrate from plastics to a surrounding medium during their lifecycle, presenting many short-and long-term human and environmental hazards Present studies have looked at the presence of BPA in plastic MCPs other than PC and epoxy resins, focusing particularly on polyvinyl chloride (PVC) where BPA is an IAS (Wang et al., 2021), and in polyethylene terephthalate (PET) where BPA is found as a NIAS (Dreolin et al., 2019). Furthermore, BPA in protective glasses, infant incubators, thermal paper, has also been investigated (Ćwiek-Ludwicka, 2015;Huang et al., 2012;Shelby, 2008;Žalmanová et al., 2016). Nevertheless, insights on human exposure to BPA via interaction with the plastic MCPs, and the synergistic relationship of BPA with other chemical substances in MCPs are limited. ...
With over 95% of BPA used in the production of polycarbonate (PC) and epoxy resins, termed herein as BPA‐based plastic materials, components and products (MCPs), an investigation of human exposure to BPA over the whole lifecycle of BPA‐based plastic MCPs is necessary. This mini‐review unpacks the implications arising from the long‐term human exposure to BPA and potential accumulation across the lifecycle of BPA‐based plastics (production, use and management). This investigation is timely and necessary in promoting a sustainable circular economy model. BPA restrictions in the form of bans and safety standards are often specific to products, while safety limits rely on traditional toxicological and biomonitoring methods that may underestimate human health implications and therefore the ‘safety’ of BPA exposure. Controversies in regards to the: a) dose‐response curves; b) the complexity of sources, release mechanisms and pathways of exposure; and/or c) the quality and reliability of toxicological studies, appear to currently stifle progress toward the regulation of BPA‐based plastic MCPs. Due to the abundance of BPA in our MCPs production, consumption and management systems, there is partial and inadequate evidence on the contribution of BPA‐based plastic MCPs to human exposure to BPA. And yet, the production, use and end‐of‐life management of plastic MCPs constitute the most critical BPA source and potential exposure pathways that require further investigation. Active collaboration among risk assessors, government, policy‐makers, and researchers is needed to explore the impacts of BPA in the long term and introduce restrictions to BPA‐based MCPs. This article is protected by copyright. All rights reserved.
Through a survey with European companies and expert interviews we study how REACH authorisation affects the phase-out of hazardous chemicals focusing on trichloroethylene, a well-studied solvent used in metal parts cleaning. We find that most of the firms have substituted trichloroethylene by perchloroethylene, which has similar chemical characteristics. This allows them to continue to use the same machines and routines at low costs. Although perchloroethylene is only classified as a suspected rather than a proven carcinogenic substance in Europe, the “improvement” as the result of much regulatory effort must be considered fairly limited, particularly in the light of less hazardous alternatives being used on the market for a long time. Our survey shows that the REACH authorisation process has some effect. Many firms cited as their main reason for substitution that they wanted to avoid the renewed application process. Still, the fact that many firms report using old machines reinforces the impression that some firms are not feeling enough pressure to modify routines and engage in a more fundamental substitution process. The results illustrate the limited effectiveness of a substance by substance approach in chemical risk management. When companies can substitute chemicals of concern to substances with similar chemical characteristics, the health and environmental objectives of chemical regulation are not achieved. An important policy conclusion is that additional incentives need to be introduced in order to realize the ambition of a non-toxic environment in the European Chemicals Strategy for Sustainability. Increased use of measures targeting broader groups of structurally similar hazardous chemicals, in combination with fees that incentivize substitution, are promising avenues for a more sustainable European chemicals strategy.
Emerging micro-pollutants (EMPs) have potential threats to human health and eco-environment. It is necessary to seek alternative treatment technologies when the existing processes have limitations to treat EMPs. This paper aimed to investigate the use of ferrate (Fe(VI)) to treat imidacloprid (IMP), bisphenol-S (BS) and azithromycin (AZM) by assessing the toxicity of the EPMs before and after Fe(VI) treatment, generating reaction kinetic rate constants and proposing degradation pathways and oxidation products. BS was readily removed at a lower Fe(VI) dose of 0.009 mM, while both IMP and AZM were not be effectively degraded at such a low Fe(VI) dose. The resulting toxicity of BS was reduced after Fe(VI) treatment, whereas that of IMP and AZM treated by Fe(VI) increased; this is due to changes in structure and property of EMPs and the formation of oxidation products (OPs), which induced either low or high toxicity. The pseudo-first-order reaction rates of IMP, BS, and AZM with Fe(VI) were derived as 0.071, 0.148, and 0.076 s⁻¹ respectively, which are consistent with the trend of the EMPs’ removing by Fe(VI).
Full-text available
Idiopathic infertility is a serious problem, which can be caused and explained by exposure to endocrine disruptors, such as bisphenols. In our study, we studied transactional exposure to bisphenol and its effects on newborn male mice throughout their reproductive life. Newborn male mice were exposed to bisphenol S and bisphenol F through maternal milk from post-natal day 0 to post-natal day 15 at concentrations of 0.1 ng.g/bw/day and 10 ng.g/bw/day, respectively. Although there were minimal differences between the control and experimental groups in testicular tissue quality and spermatozoa quality, we discovered an interesting influence on early embryonic development. Moderate doses of bisphenol negatively affected cleavage of the early embryo and subsequently, the blastocyst rate, as well as the number of blastomeres per blastocyst. In our study, we focused on correlations between particular stages from spermatogenesis to blastocyst development. We followed epigenetic changes such as dimethylation of histone H3 and phosphorylation of histone H2 from germ cells to blastocysts; we discovered the transfer of DNA double-strand breaks through the paternal pronucleus from spermatozoa to blastomeres in the blastocyst. We elucidated the impact of sperm DNA damage on early embryonic development, and our results indicate that idiopathic infertility in adulthood may have causes related to the perinatal period.
Bisphenol A (BPA) and bisphenol S (BPS) are agonists of hERα receptors and due to BPA regulations in many countries, several substitutes that are close analogs to BPA and BPS were developed. In the presented study, we have determined human estrogen receptor (hER)α agonist and antagonist activities with the validated OECD assay with the hERα-Hela9903 cell line for five different chemical classes of BPA and BPS analogs. This study also defined clear structure-activity relationships for agonist and antagonist activities of the 12 bisphenols on hERα, which are supported by molecular docking studies. These data show that classical analogs of BPA (e.g., bisphenols B, C, AP, E) have comparable or superior estrogenic agonist potencies compared to BPA and BPS. The most potent of these hERα agonists were even more potent than BPA, as bisphenol B and C, with IC50 values of 0.31 μM and 0.48 μM, respectively. Among these selected bisphenols, 4-4'-methylenebis(oxyethylenethio)diphenol was the most potent hERα antagonist, with an IC50 of 0.39 μM. The estrogenic agonist and antagonist potencies of these different chemical classes of BPA and BPS analogs are mutually comparable and can be used as a basis for further structure-activity relationships studies and human risk assessment.
Full-text available
Bisphenol A (BPA) and its analogs, bisphenol S (BPS) and bisphenol F (BPF), might impact fertility by altering oxidative stress pathways. Here, we hypothesize that bisphenols-induced oxidative stress is responsible for decreased gamete quality. In both female (cumulus-oocyte-complexes—COCs) and male (spermatozoa), oxidative stress was measured by CM-H2DCFDA assay and key ROS scavengers (SOD1, SOD2, GPX1, GPX4, CAT) were quantified at the mRNA and protein levels using qPCR and Western blot (COCs)/immunofluorescence (sperm). Either gamete was treated in five groups: control, vehicle, and 0.05 mg/mL of BPA, BPS, or BPF. Our results show elevated ROS in BPA-treated COCs but decreased production in BPS- and BPF-treated spermatozoa. Additionally, both mRNA and protein expression of SOD2, GPX1, and GPX4 were decreased in BPA-treated COCs (p < 0.05). In sperm, motility (p < 0.03), but not morphology, was significantly altered by bisphenols. SOD1 mRNA expression was significantly increased, while GPX4 was significantly reduced. These results support BPA’s ability to alter oxidative stress in oocytes and, to a lesser extent, in sperm. However, BPS and BPF likely act through different mechanisms.
Full-text available
The endocrine disruptors are chemicals with the capacity to influence physiological processes in the organism, most often through hormonal control. They are present in the environment and in the products of daily use. They are often found in food, released from plastic bottles for water, present in cosmetics or fertilizers. Latest research suggests that they can be released from plastics used in the IVF laboratories and can be even present in the manipulation and cultivation media used for isolation and fertilization of gametes and subsequent cultivation of embryos. Permanent and long-term utilization of these substances has adverse effects in human reproductive health, mainly by the means of interfering with synthesis and action mechanisms of reproductive hormones. Moreover, some endocrine disruptors show a range of adverse effects directly on the gametes or embryos cultured in the in vitro conditions. The article provides an overview on bisphenols detected in plastics and media commonly used in the IVF laboratory and considers their possible impact on effectiveness of the IVF methods in a human laboratory.
Today, the interest in hair as alternative matrix for human biomonitoring of environmental pollutants has increased, but available data on chemical levels in hair remain scarce. In this study, the measurement of 2 bisphenols (A and S), 3 parabens (methyl-, ethyl- and propylparabens) and 8 perfluroralkyl compounds (PFCs) namely perfluoroctanesulfonate (PFOS), perfluorohexanesulfonate (PFHxS), perfluorooctanoic acid (PFOA), perfluorononanoic acid (PFNA), perfluorodecanoic acid (PFDA), perfluroroheptanoic acid (PFHpA), perfluoropentanoic acid (PFPeA) and perfluorohexanoic acid (PFHxA) was carried out, using a thoroughly validated UPLC-MS/MS method, in the hair from 114 adults living in Liege (Belgium) and surrounding areas. The most frequently quantified compounds in the population were: bisphenol S (97.4%, median = 31.9 pg·mg⁻¹), methylparaben (94.7%, median = 28.9 pg·mg⁻¹), bisphenol A (93.9%, median = 46.6 pg·mg⁻¹), ethylparaben (66.7%, median = 5.2 pg·mg⁻¹), propylparaben (54.8%, median = 16.4 pg·mg⁻¹) and PFOA (46.4%, median < 0.2 pg·mg⁻¹). The other PFCs were detected only in few samples although current exposure of the Belgian population to PFCs was previously demonstrated using blood analyses. Nonparametric statistical analyses were performed to evaluate the influence of gender, hair treatments and hair length, but no significant difference was observed. Only age was positively correlated with the propylparaben contamination. Although blood seems to remain more suitable for PFCs exposure assessment, the results of this study suggest that hair can be an appropriate matrix for biomonitoring of organic pollutants such as parabens or bisphenols.
Full-text available
The general population is potentially exposed to many chemicals that can affect the endocrine system. These substances are called endocrine disruptors (EDs), and among them bisphenol A (BPA) is one of the most widely used and well studied. Nonetheless, there are still no data on simultaneous measurements of various EDs along with steroids directly in the seminal fluid, where deleterious effects of EDs on spermatogenesis and steroidogenesis are assumed. We determined levels of BPA and 3 estrogens using LC-MS/MS in the plasma and seminal plasma of 174 men with different degrees of infertility. These men were divided according their spermiogram values into 4 groups: (1) healthy men, and (2) slightly, (3) moderate, and (4) severely infertile men. Estradiol levels differed across the groups and body fluids. Slightly infertile men have significantly higher BPA plasma and seminal plasma levels in comparison with healthy men (p<0.05 and p<0.01, respectively). Furthermore, seminal BPA, but not plasma BPA, was negatively associated with sperm concentration and total sperm count (-0.27; p<0.001 and -0.24; p<0.01, respectively). These findings point to the importance of seminal plasma in BPA research. Overall, a disruption of estrogen metabolism was observed together with a weak but significant impact of BPA on sperm count and concentration.
Full-text available
Bisphenol A (BPA) is a well-known environmental endocrine-disrupting chemical and bisphenol S (BPS) has been considered a "safer" alternative for BPA-free products. The present study aims to evaluate the impact of BPA and BPS on the reproductive neuroendocrine system during zebrafish embryonic and larval development, and to explore potential mechanisms of action associated with estrogen receptor (ER), thyroid hormone receptor (THRs) and enzyme aromatase (AROM) pathways. Environmentally relevant, low levels of BPA exposure during development led to advanced hatching time, increased numbers of Gonadotropin-Releasing Hormone 3 (GnRH3) neurons in both terminal nerve and hypothalamus, increased expression of reproduction-related genes (kiss1, kiss1r, gnrh3, lhβ, fshβ, and erα) and a marker for synaptic transmission (sv2). Low levels of BPS exposure led to similar effects: increased numbers of hypothalamic GnRH3 neurons and increased expression of kiss1, gnrh3, and erα. Antagonists of ER, THRs and AROM blocked many of the effects of BPA and BPS on reproduction-related gene expression, providing evidence that those three pathways mediate the actions of BPA and BPS on the reproductive neuroendocrine system. This study demonstrates that alternatives to BPA used in the manufacture of BPA-free products are not necessarily safer. Further, this is the first study to describe the impact of low-level BPA and BPS exposure on the Kiss/Kissr system during development. It is also the first report of multiple cellular pathways (ERα, THRs and AROM) mediating the effects of BPA and BPS during embryonic development in any species.
Full-text available
In the sphere of reproductive biotechnologies, the demand for sufficient numbers of high-quality oocytes is still increasing. In some cases, this obstacle is overcome by in vitro prolonged cultivation. However, a prolonged oocyte culture is accompanied by changes called ageing. Ageing is manifested by spontaneous parthenogenetic activation, programmed cell death or lysis. Various substances, such as caffeine or dithio- threitol, have been tested for ageing suppression. In this respect, research into gasotransmitters (hydrogen sulphide, carbon monoxide, and nitric oxide) has currently been intensified. The objectives of the present study were to localize nitric oxide synthases (NOS) and to evaluate NOS inhibition of aged porcine oocytes. We demonstrated the presence of NOS isoforms in oocyte cultivation prolonged by 24, 48, and 72 h. After 72 h of prolonged cultivation, NOS inhibition by the non-specific inhibitor L-NAME or the specific inhibitor aminoguanidine caused suppression both of programmed cell death and lysis. Although NOS amount rapidly decreased after the 72-h cultivation, changes induced by NOS inhibition were statistically significant. We can presume that NOS play an important physiological role in porcine oocyte ageing.
Full-text available
Oocyte ageing is a complex of processes that occur when matured in vitro oocytes are, after reaching the metaphase II stage, exposed to further in vitro culture. Aged oocytes remaining at the metaphase II stage undergo spontaneous parthenogenetic activation, or cellular death, through apoptosis (fragmentation) or lysis. The key factor in apoptotic pathway regulation is c-Jun-N-terminal kinase (JNK), stress kinase from the mitogene-activated protein kinase (MAPK) family. To investigate the effect of JNK inhibition on porcine oocytes ageing, cleavage rate, and embryonic development after parthenogenetic activation, DNA fragmentation, and pro-apoptotic factor Bax expression, we cultured in vitro matured oocytes for another 1-4 days in the presence of a JNK inhibitor. The inhibition of JNK significantly protected the oocytes from fragmentation (0% of fragmented oocytes under JNK inhibition vs. 13.4% of fragmented oocytes in the control group, 2nd day of ageing) and increased the percentage of parthenogenetically activated oocytes (82 vs 57.7%, 2nd day of ageing). The embryonic development of oocytes parthenogenetically activated after 24 h of ageing was influenced by JNK inhibition as well. The percentage of oocytes at the morula stage, after seven days of cultivation, was significantly increased when oocytes aged in the presence of a JNK inhibitor (42.5%) by comparison to the percentage of oocytes exposed to ageing in an inhibitor-free medium (23.3%). DNA fragmentation was significantly suppressed by JNK inhibition from the 1st day of ageing, but the expression of pro-apoptotic factor Bax in the oocytes was not influenced. On the basis of our experiments, we can conclude that JNK inhibition suppresses apoptosis and DNA fragmentation of aged oocytes and improves their embryonic development following the parthenogenetic activation. However, to completely eliminate all ageing related processes is insufficient.
Full-text available
Bisphenol F (BPF) was found in mustard up to a concentration of around 8 mg kg(-1). Contamination of the raw products or caused by the packaging could be ruled out. Also the fact that only the 4,4'-isomer of BPF was detected spoke against a contamination from epoxy resin or other sources where technical BPF is used. Only mild mustard made of the seeds of Sinapis alba contained BPF. In all probability BPF is a reaction product from the breakdown of the glucosinolate glucosinalbin with 4-hydroxybenzyl alcohol as an important intermediate. Hot mustard which was made only from brown mustard seeds (Brassica juncea) or black mustard seeds (Brassica nigra) contained no BPF. BPF is structurally very similar to bisphenol A and has a similar weak estrogenic activity. The consumption of a portion of 20 g of mustard can lead to an intake of 100 to 200 µg of BPF. According to a preliminary risk assessment, the risk of BPF in mustard for the health of consumers is considered as low, but available toxicological data are insufficient for a conclusive evaluation. It is a new and surprising finding, that BPF is a natural food ingredient and that this is the main uptake route. This insight sheds new light on the risk linked to the family of bisphenols.
Full-text available
Endocrine disrupting chemicals (EDC) abound in the environment since many compounds are released from chemical, agricultural, pharmaceutical and consumer product industries. Many of the EDCs such as Bisphenol A (BPA) have estrogenic activity or interfere with endogenous sex hormones. Experimental studies have reported a positive correlation of BPA with reproductive toxicity, altered growth and immune dysregulation. Although the precise relevance of these studies to the environmental levels is unclear, nevertheless, their potential health implications remain a concern. One possible mechanism by which BPA can alter genes is by regulating epigenetics, including microRNA, alteration of methylation and histone acetylation. There is now wealth of information on BPA effects on non-lymphoid cells and by comparison, paucity of data on effects of BPA on the immune system. In this mini review, we will highlight BPA regulation of estrogen receptor-mediated immune cell functions and in different inflammatory conditions. In addition, BPA-mediated epigenetic regulation of non-lymphoid cells is emphasized. We recognize that most of these studies are on non-lymphoid cells, and given that BPA also affects the immune system, it is plausible that BPA could have similar epigenetic regulation in immune cells. It is hoped that this review will stimulate studies in this area to ascertain whether or not BPA epigenetically regulates the cells of the immune system.
Reproductive tract development is influenced by estrogen. The aim of this study was to determine the effects of an environmental estrogenic chemical bisphenol-A (BPA) on prenatal and postnatal development of female mouse reproductive organs. In the prenatal treatment group, BPA or the synthetic estrogen diethylstilbestrol (DES) were given by subcutaneous (s.c.) injections to pregnant mice during gestational days 10–18. Some offspring treated prenatally with 10 and 100 mg/kg bw BPA or 0.67 and 67 μg/kg bw DES were ovariectomized at 30 days and sacrificed at 40 days of age. Vaginal smears were examined in the remaining offspring, then these offspring were mated with normal males. Prenatal exposure to 10 mg/kg BPA reduced the number of mice with corpora lutea compared to sesame oil controls at 30 days, but more than 80% of mice from either prenatally exposed BPA group were fertile at 90 days. Mice exposed prenatally to maternal doses of 67 μg/kg DES were sterile and showed ovary-independent vaginal and uterine epithelial stratification; however, mice exposed prenatally to BPA did not show ovary-independent vaginal and uterine changes. The number of offspring and litter sex ratio from mice exposed prenatally to BPA (10 or 100 mg/kg) or 0.67 μg/kg DES were not different compared to controls.
The study of oral exposure to the environmental estrogen Bisphenol A (BPA) during the perinatal period and its effects on ovarian functionality in adulthood has generated special interest. Thus, our objective was to investigate ovarian folliculogenesis and steroidogenesis in adult female rat offspring born to mothers exposed to low doses of BPA (BPA50: 50μg/; BPA0.5: 0.5μg/ by the oral route during gestation and breastfeeding. Ovaries from both BPA-treated groups showed reduced primordial follicle recruitment and a greater number of corpora lutea, indicating an increased number of ovulated oocytes, coupled with higher levels of mRNA expression of 3β-hydroxysteroid dehydrogenase and serum progesterone. BPA50-treated animals had lower expression of androgen receptor (AR) at different stages of the growing follicle population. BPA0.5-treated rats evidenced an imbalance of AR expression between primordial/primary follicles, with higher mRNA-follicle-stimulating hormone receptor expression. These results add to the growing evidence that folliculogenesis and steroidogenesis are targets of BPA within the ovary.