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DOI 10.2478/pjvs-2013-0111
Review
Alkylphenol ethoxylates and alkylphenols
– update information on occurrence,
fate and toxicity in aquatic environment
J. Kovarova
1
, J. Blahova
1
, L. Divisova
1
, Z. Svobodova
1
1
Department of Veterinary Public Health and Animal Welfare, Faculty of Veterinary Hygiene and Ecology,
University of Veterinary and Pharmaceutical Sciences Brno,
Palackeho tr. 1/3, 612 42 Brno, Czech Republic
Abstract
Alkylphenols and their precursors, alkylphenol etoxylates, are a group of manmade chemicals
used mainly as surfactants in domestic and industrial applications worldwide. It has been well estab-
lished that they have endocrine disruption activity, hepatotoxic, genotoxic and other negative effects
on animal and human health. In spite of the effort to reduce their use, they persist in the environment
not only in industrial but also in remote regions, and were detected in the variety of natural matrices
including air, water, soil as well as food products, and human blood and urine worldwide. This article
summarizes their occurrence, fate in natural conditions, and toxicity including mode of action. A sub-
ject of our concern was the aquatic environment as the most important reservoir and target of their
deleterious impact.
Key words:Nonylphenol, octylphenol, wastewater, xenoestrogen, fish, vitellogenin
Abbreviationsand units
AhR – aryl hydrocarbon receptor
AP – alkylphenol
APE – alkylphenol ethoxylates
EDC – endocrine disrupting chemicals
EO – ethoxylate portmon
ER – oestrogen receptor
NP – nonylphenol
OP – octylphenol
STP – sewage treatment plants
WWTP – wastewater treatment plant
Correspondence to: J. Blahova, email: blahovaj@vfu.cz, tel.: +420 541 562 785
Introduction
Alkylphenol ethoxylates (APE) are one of the
most widely used classes of surfactants. Recently, ap-
proximately 500,000 tons have been produced world-
wide annually (Renner 1997) and it makes APEs the
world’s third largest group of surfactants in terms of
production and use (Ying et al. 2002). They can be
used as detergents, wetting agents, dispersants, emul-
sifiers, solubilizers and foaming agents. APEs are im-
portant to a number of industrial applications, includ-
ing pulp and paper, textiles, coatings, agricultural pes-
Polish Journal of Veterinary Sciences Vol. 16, No. 4 (2013), 763–772
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R
OOOH
R
OOO
mOH
R
OOOH
m-1
OOH
R
O
OOH
R
OH
R
Alkylphenol polyethoxylate
(APnEO,n=m+1)
m
R = C H , nonyl
C H , octyl
R is usually branched
919
817
Alkylphenoxycarboxylic acid
(ApnEC)
Alkyphenol polyethoxylate
AP(n - 1)EO
Progressive shortening
of ethoxylate chain
Alkylphenoxyacetic acid
(AP 1 EC) Alkylphenol monoethoxylate
(AP 1 EO)
Alkylphenol (AP)
Ring cleavage, oxidation of alkyl chain
Fig. 1 Degradation pathway of alkylphenolethoxylate (Renner 1997).
ticides, lube oils and fuels, metals and plastics. Indus-
trial applications comprise 55% of the APE market.
The remaining uses include industrial and institu-
tional cleaning products (30%), household cleaning
products (15%) and other miscellaneous uses (<1%).
APEs are surfactants manufactured by reaction of al-
kylphenols (AP) with ethylene oxide. APE molecule
consists of two parts – AP is fairly non-polar portion
which allows to dissolve grease and other materials
that have small water solubility, and the ethoxylate
portion (EO) of the surfactant is water-soluble and
aids in the transfer of material to the aqueous phase.
This structure makes most of polar APEs soluble in
water and helps disperse dirt and grease from soiled
surfaces into water (Snyder et al. 2001).
APs, in addition to their role as a raw material for
APEs, are used in the preparation of phenolic resins,
polymers, heat stabilizers, antioxidants and curing
agents. From this group of chemicals, nonylphenol
(NP) is by far the most commercially important AP in
Europe, with an annual production of about 75,000
tons, 60% of which is used to make NPEs (RPA
1999). Moreover, NP as well as octylphenol (OP) and
AP mono- to triethoxylates (APE1, APE2 and APE3)
are more persistent and more toxic metabolites gener-
ated by degradation of APEs. These metabolites are
ubiquitous in the environment because of widespread
use of APEs surfactants and widespread lack of ad-
equate wastewater treatment, and these compounds
have been detected in air, water, sediment, soil and
biota at different levels in different parts of the world
(Giger et al. 1984).
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Occurrence in environment
APEs and their degradation products are not pro-
duced naturally. Their presence in the environment is
solely a consequence of anthropogenic activity. Due
to their exceptional nature they can be considered as
amphophilic compounds. Lower APE oligomers (EO
<5) are usually described as water-insoluble or
lipophilic, whereas higher oligomers are described as
water-soluble and hydrophilic (Ahel and Giger 1993,
Jonsson et al. 2008). Another very important feature
of these chemicals is their ability to bind with organic
matter because of their very low partition coefficient
(log Kow for APEs metabolites between 3.90 and
4.48) (Ying et al. 2002). Therefore, APEs and/or their
metabolites were determined in whole range of
abiotic and biotic matrices all over the world. As men-
tioned above, APEs are readily biodegraded under
both aerobic and anaerobic conditions in the environ-
ment (Paasivirta and Rantio 1991) including microbial
degradation, losing most of EO units; the final break-
down products are APEs with one or two EO (APE1,
APE2), alkylphenoxy carboxylic acid, and APs
(Fig. 1). Many studies assessing levels of APEs and
their emerging metabolites have been published in re-
cent years and have shown the ubiquitous distribution
of these chemicals in sewage treatment plant dis-
charges, sediments and live organisms.
Occurence in wastewater discharges
This partial biodegradation was observed in a ma-
jority of sewage treatment plants (STP) during sewage
treatment processes that makes from sewage treatment
plant effluents and wastewater discharges an important
source of this type of compounds (David et al. 2009).
The efficiency of wastewater treatment plants
(WWTPs) in removal of NP was found to be highly
variable ranging from 11% to 99% depending on type
of treatment process unit employed (Berryman et al.
2004). A treatment process composed of ozonation and
subsequent activated carbon filtration with chlorination
was the most effective (removal of 95% of NP) (Pet-
rovic et al. 2003). If contaminants are adsorbed on ac-
tivated sludge particles, they accumulate in the
WWTPs sludge. In this case, the application of diges-
ted sludge, as fertilizer, on agricultural fields may cause
a potential contamination of soil and ground water
(Olea et al. 1996). Concentrations of ΣAPE meta-
bolites in treated wastewater effluents, for example in
the US, ranged from <0.1 to 369 μg/L (Rudel et al.
1998), in Spain they were between 6 and 343 μg/L (Sole
et al. 2000) and concentrations up to 330 μg/L were
foundintheUK(BlackburnandWaldock1995).
Occurrence in water
Many communities worldwide, such as Europe,
use surface or groundwater resources for drinking
water production, which contain a significant portion
of wastewater effluent (Sonnenschein and Soto 1998).
In some drinking wells these chemicals were detected
at concentrations ranging from <limit of detection to
32.9 μg/L (Rudel et al. 1998). To date, the amount of
APs pollution detected in aquatic environments most-
ly ranges from nanograms to milligrams per litre
(Uguz et al. 2003). Same data were presented also by
Naylor et al. (1992), who reported that water concen-
tration of NP seldom overwhelms 10 μg/L, though at
some “hot spots” concentrations reach 1000 μg/L
(Warhurst 1995).
In groundwater samples APs concentrations were
usually higher than those found in surface water. The
removalofcontaminantsisinfactveryslowinground-
water since chemical and biological characteristics in
the aquifers are not favourable for degradation pro-
cess. Groundwater temperatures are in the psych-
rophilic range and both carbon sources and oxygen
are limited. Microbiological resources of such ecosys-
tems are restricted and contaminants undergo ex-
tremely slow degradation process allowing con-
taminants to disperse up to several kilometers from
the contamination source and to exist for decades
(Soares et al. 2008).
In surface water the amount of this type of pollu-
tants is lower (Arditsoglou and Voutsa 2008). In Asia,
NP concentrations were found within the range of 0.3
to 2.8 μg/L (Basheer et al. 2004). Nevertheless, even
these numbers exceed environmental quality stan-
dards established by EU authorities to achieve the
good surface water quality status. The environmental
quality standards, as annual average concentration,
for NP and OP in surface water were proposed to 0.3
μg/L and 0.1 μg/L, respectively, however, environ-
mental quality standards are still not harmonised
throughout the EU (European Commission 2006).
Furthermore, these environmentally realistic concen-
trations of NP may have toxic effect on aquatic fauna
(Ying et al. 2002). Interestingly, APs contamination in
surface water shows seasonal variations. NP and OP
concentrations were observed higher in the warmer
season than in the colder part of year (Tsuda et al.
2000, Isobe et al. 2001). It was suggested that a micro-
bial activity at warmer temperatures leads to an en-
hanced degradation of NPE and OPE (Li et al. 2004)
as well as photolysis induced by sunlight can influence
the concentration of APs in the surface layer of natu-
ral waters during summer (Ahel et al. 1994). Just as
temperature and light availability also salinity was de-
termined as a factor influencing the presence of APs
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in the water (Li et al. 2005). Sampling locality choice
influenced measured concentrations in a very import-
ant manner. A shallow part of continental shelf, along
the coast, is the most sensitive due to smaller volume
of water and proximity of point and diffuse sources of
wastewater discharge or other specific source of APs
pollution such as produced water. The produced
water is a by-product of current oil-production tech-
nology and consists of seawater containing an ex-
tremely complex mixture of dispersed oil, polycyclic
aromatic hydrocarbons, APs, organic acids, metals
and traces of production chemicals (Meier et al.
2011). In contrast, in deep sea areas concentrations
decrease rapidly as the distance from the coast in-
creases (Xie et al. 2006). Measurements of APs con-
tent in water column have obvious shortcomings
caused by the unfavourable physicochemical charac-
teristics. The environmental distribution is estimated
to 25% in water, more than 60% in sediments, more
than 10% in soil (Nordic Council of Ministers 1996).
Occurrence in water sediments
In the aquatic environment AP was found at high-
er concentrations in associations with sediments
rather than dissolved in the aqueous phase. Although,
organic content of the sediments was one of the im-
portant determinants of adsorption process. In sedi-
ments free of organic matter, adsorption was also ob-
served indicating that not only organic content is rel-
evant for NP. It was chased up that NP adsorption is
controlled by two major interactions: hydrophilic in-
teraction with mineral components and hydrophobic
interaction with organic matter (John et al. 2000). NP
concentrations of up to 20,700 ng/g were found in
sediments sampled near a sewage outlet in Tokyo Bay
(Kurihara et al. 2007). NP was also found at high
levels (13700 ng/g) in sediment in Jamaica Bay (USA)
impacted by sewage inputs (Fergusson et al. 2001).
Remarkably, in a study of Kelly et al. (2010) chemical
analysis of representative site samples of the Shannon
International River Basin District in Ireland also
identified APs in water and sediments in mg/L and
mg/kg concentrations, respectively. In sediments there
was determined another interesting phenomenon like
the vertical distribution of APEs and APs. It could be
influenced by many factors as vertical transport by
pore water diffusion or physical and/or bio mixing.
Peng et al. (2007) found a decline in NP concentra-
tions in the layer dated to the end of 1980s, corre-
sponding to the onset of wastewater treatment,
whereas the recorded increase in the NP concentra-
tion in the layer dated to 1990s could be explained by
the increased economic growth and the lack of effi-
cient wastewater facilities. Sediments are the main
medium, where persistent organic pollutants with
some lipophilic nature can be stored. For the NP
a risk of release from sediment has been evaluated
and it was concluded that, due to an unfavourable
ratio of sizes of the hydrophilic and hydrophobic parts
of the molecule, NP is not subject to micelle forma-
tion. Thus, NP is not expected to be able to mobilize,
e.g., hydrophobic pollutants as PAHs (Brix et al.
2001). In accordance with this findings the estimation
of APs half-lives was provided. Ekelund et al. (1993)
reported the half-lives of NP based on ultimate biode-
gradation of about 58 days in seawater and 35 days in
aerobic seawater plus sediment at 11
o
C while Shang et
al. (1999) reported NP’s half-life in sediments more
than 60 years.
Occurrence in biological matrices
The accumulation tendency of these group of sub-
stances is evident not only in inorganics but also in the
biological matrices, particularly aquatic biota (Ahel et
al. 1994). However, in a study presented by Arukwe et
al. (2000) it was suggested that NP does not accumu-
late in tissues and that a half-life of NP in the tissues
of the rainbow trout (Oncorhynchus mykiss) is about
24-48 hours. Based on many other studies it was con-
cluded that APs are able to accumulate in several
aquatic species, including aquatic plants, algae, fish
and mussels (Liber et al. 1999, Lintelman et al. 2003).
In fish, the highest concentration of APs was found in
the fat of exposed organisms (Shiraishi et al. 1989)
and it was reported that NP bioaccumulates up to
410-fold in fish (Tsuda et al. 2000, Snyder et al. 2001).
High accumulation potential was also determined for
oysters and snails that both serve as transport media
and potential sources of APs in marine environment
(Cheng et al. 2006). It was also noted that an ability of
marine organisms to accumulate APs depends on
their feeding strategy. A higher proportion of APs was
found in organisms with the highest exposure to sedi-
ments (e.g. filter feeding organisms) and so APs con-
centrations in marine organisms were higher in be-
nthic and pelagic organisms that live close to the sedi-
ment than in fish (David et al. 2009). These findings
revealed that the main source of APs for marine or-
ganisms is not water column and intake through gills
but sediments and dietary uptake, both depending on
the AP distribution in the aquatic environment. On
the other hand, uptake through skin and gills was also
demonstrated (Ferreira-Leach and Hill 2001). A dis-
tribution of APs in live organisms was also inves-
tigated and, due to partly lipophilic nature of APs,
their target tissues in fish (such as the adipose tissue,
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brain, liver, kidney, muscle, skin and eye) are rich in
grease (Ferreira-Leach and Hill 2001). The study, us-
ing radio-labelled compounds demonstrating their
distribution, indicates that the liver and bile are
a main excretion route for absorbed APs, and the con-
centration of AP metabolites excreted in bile in-
creased with increasing degree of alkylation. It should
be emphasized that uptake, metabolism and elimin-
ation of xenobiotics, including APs, vary with com-
pound, exposure route and fish species (Tollefsen and
Nilsen 2008). The muscle tissue of the chub (Leucis-
cus cephalus) in the Czech Republic contains
1.92-6.11 ng/g w.w. of ΣNP + OP (Zlabek et al. 2006).
Some fish species from Adriatic Sea, Italy, have NP
levels from 12 to 1,431 ng/g w.w. and OP levels of
0.3-4.7 ng/g w.w. (Ferrera et al. 2005). Basheer et al.
(2004) presented concentrations of NP about 60.5
ng/g w.w. and OP about 31.4 ng/g w.w. in fish in Sin-
gapore. Fish in the UK were found to contain up to
0.8 ng/g w.w. of NP in the muscle tissue (Ying et al.
2002), whereas geometrical mean of NP in snails and
oysters was calculated in range of 380-1,560 ng/g and
92-1,080 ng/g d.w., respectively. Bioconcentration fac-
tors for oysters and snails exceed those in fish (Cheng
et al. 2006). Reported bioconcentration factor of NP
in whole fish ranged from 280 to 1,300
(Ferreira-Leach and Hill 2001). As mentioned above
the highest values of APs were determined in fish bile,
because glucuronic acid conjugates are the main ex-
cretion form of APs. In the study of Jonsson et al.
(2008) the levels of OP reached 579 ng/g and NP 19
ng/g of fish bile in the North Sea. The levels of NP
detected in fish bile in the UK reached 195-2,453
ng/mL and, in fish exposed to wastewater discharge,
values were even up to 12,678 ng/mL (Fenlon et al.
2010). It could be concluded that variability in APs
accumulation in water organisms is affected by feed-
ing habits, metabolism, levels of contamination in the
individual habitats, biotransformation and excreting
capacity.
Toxicity and risk factors
An increasing concern for APs occurrence in the
environment was initiated by investigations of
feminized male fish found near sewage outlets in sev-
eral rivers all over the world; a mixture of chemicals
containing APs that results from degradation of deter-
gents during sewage treatment seemed to be a causal
agent of this endocrine disruption (Purdom et al.
1994, Auriol et al. 2006). Even though suspicion of
toxicity of APs was declared in 1938 by Dodds and
Lawson (1938) and confirmed by Soto et al. (1991),
the aim to restrict the use of this group of chemicals
was accomplished in Europe in 2001 by their inclusion
in the list of priority hazardous substances for surface
waters (European Commission 2000) and in the USA
by developing “chronic criteria recommendations” for
NP by US EPA (2005). This procedure is in accord-
ance with the fact that the majority of chemicals that
have been in production before 1981 has not been
subjected to the present requirements for the environ-
mental evaluation of new chemicals (Wallstrom 2004).
Therefore, only 3% of 2,500 high-volume chemicals
currently in use are well-tested, including their
long-term effects (Soares et al. 2008). Intensive re-
search focused on mode of action as well as other
toxicologic potential of these compounds was started
since the ability of NP, OP and APE1, APE2 and
APE3, present in the environment at sufficient levels,
to mimic natural hormones and disrupt endocrine
function was demonstrated (Tabata et al. 2001). Be-
sides the main detrimental xenoestrogenic effect,
these compounds were confirmed as potential car-
cinogens (Pedersen et al. 1999, Uguz et al. 2003),
hepatotoxins (Meier et al. 2007), genotoxins (Hwang
et al. 2010) and behavioural modulators affecting
basic survival reflexes such as locomotors activity or
aggression (Xia et al. 2010). On the molecular basis,
APs have been reported to bind to oestrogen recep-
tors (ER) of both fish and mammals (Gutendorf and
Westendorf 2001, Tollefsen 2002) and activate estro-
gen-responsive element regulating reporter genes
(Kuiper et al. 1998, Legler et al. 1999). They are ca-
pable to interfere with steroid hormone receptor ac-
tivity (Kudo and Yamaguchi 2005), steroid metab-
olism (Arukwe et al. 1997) and plasma endogenous
steroids in aquatic animals (Tollefsen 2007), too. It
indicates that this group of pollutants acts through
multiple mechanisms to cause potentially endocrine
disruption.
In general, the endocrine and reproductive effects
of chemicals are believed to exist due to their ability
to: (1) mimic the effect of endogenous hormones, (2)
antagonize the effect of endogenous hormones, (3)
disrupt the synthesis and metabolism of endogenous
hormones, and (4) disrupt the synthesis and metab-
olism of hormone receptors (Sonnenschein and Soto
1998). APs have been shown to cover more than two
of these effects. Very interesting outcomes that some
of the APs displayed apparent antagonistic activity at
low micromolar concentration whereas acting as weak
estrogens at high concentrations were presented by
Silva et al. (2002). Thus APs and alkylated non-phen-
olics act as both agonistic and antagonistic estrogens
depending on concentration. Antiandrogenic activity
through affecting aromatase activity and a function of
the aryl hydrocarbon receptor (AhR) was also con-
firmed by Bonefeld-Jogensen et al. (2007). It follows
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that, except to “classical” oestrogen receptors, APs
are able to bind to the AhR, which regulates some
ER-responsive genes (Safe and Gaido 1998); the
pregnant X receptor, which may regulate endogenous
steroid metabolism (Masuyama et al. 2000), the nic-
otinic receptor, where NP directly modulates ion cur-
rent (Nakazawa and Ohno 2001) and even cAMP re-
sponsive element as a key signalling route of many
hormones and substances (Cheshenko et al. 2008).
Some APs also induced significantly a p53 gene ex-
pression, as demonstrated in copepod, indicating the
involvement of p53 in such stress-responses (Hwang
et al. 2010). Recently, Yang et al. (2006) also assumed
that p53 gene would be closely related to the impacts
of endocrine disrupting chemicals (EDCs) on ecosys-
tem biota and human health. They were also deter-
mined as potent inhibitors of Ca
2+
-ATPase disrupting
Ca
2+
homeostasis in fish, perhaps contributing to cell
and tissue damage during key stages in the develop-
ment and thereby influencing normal endocrine func-
tion later in life (Kirk et al. 2003). A recently dis-
covered neuropeptide, kisspeptin, crucially important
substance for the triggering of puberty and for future
fertility in mammals (Oakley et al. 2009), and similar
function in teleost of fish (Katashi et al. 2009) is sus-
pected to be affected by APs, too.
It has been reported that NP exerts its biochemi-
cal effects by down regulating an activity of micro-
somal cytochrome P4501A while stimulating an in-
crease in the cytochrome P4503A protein in the liver
of mammals and fish (Uguz et al. 2003). It was shown
that 220 μg/L of NP induce neoplastic proliferation of
reticuloendothelial cells in fish hepatopancreas after
2 weeks and an increase in glutathione S-transferase
activity after 1 week and replacement of the normal
liver parenchyma by the fibrous tissue after 4 weeks
(Arukwe et al. 2000). Thus, it could be assumed that
NP has also the ability to cause variations in isoforms
of hepatic cytochrome P450-dependent steroid and
xenobiotic metabolizing enzymes (Servos 1999). Final-
ly, not all APs are able to disrupt the reproductive
function in a live body. Interesting observations per-
taining to structure-function relationships were made:
(i) the alkyl chain must have at least 3 carbons, (ii)
only the p-isomers are estrogenic, (iii) polyalkylated,
hindered phenols like butylated hydroxytoluene and
Irganox 1640 (Ciba-Geigy) are not estrogenic while
they are effective antioxidants, and (iv) fused rings
like naphthols are not estrogenic in spite of being an
integral part of the A and B ring of natural steroids
(Sonnenschein and Soto 1998). The estrogenicity is
dependent on ligand-binding affinity, transcriptional
and post-transcriptional regulation of ER-dependent
genes, and the toxicokinetics of the compound (Beres-
ford et al. 2000, Katzenellenbogen et al. 2000).
Since the toxicity potential of APs was deter-
mined, monitoring strategies to assess these com-
pounds and their biological impact in the environment
have been developed. As endocrine disrupting chemi-
cals they can affect a number of reproductive par-
ameters in fish, including gonadal development
(Meier et al. 2007), induction of plasma vitellogenin
in male and juvenile fish (Jobling and Sumpter 1993),
inhibition of spermatogenesis (Jobling and Sumpter
1993) and oogenesis (Weber et al. 2003), all of them
could be used as markers of endocrine disruption. The
hepatic induction of vitellogenin production in male
and juvenile oviparous fish, which occurs normally
only in maturing females under stimulation of 17β-es-
tradiol, has been proposed as a sensitive biomarker of
the exposure to estrogenic chemicals of endogenous
and exogenous origin (Sumpter and Jobling 1995,
Randak et al. 2009). Interestingly, the calculated con-
tribution of the anticipated NP and OP to the poten-
tial estrogenicity of discharge from healthcare facili-
ties was found greater than 65% of the total potential
estrogenicity (Nagarnaik et al. 2010). Healthcare facil-
ities like hospitals, nursing facilities, assisted living fa-
cilities, and independent living facilities, were suspec-
ted of releasing immense amount of natural or syn-
thetic hormone metabolites and of responsibility for
endocrine disruption. According to the study of
Nagarnaik et al. (2010) their share of endocrine dis-
ruption is based more on releasing of APs that have
weaker estrogenic effectsthan hormones but the APs
are released in higher amounts. APs were present in
part per billion concentrations in water samples taken
near healthcare facilities compared to the
part-per-trillion concentrations of hormones. An in-
tensive use of detergents in healthcare facilities and
industry could be more important source of NPs in
the environment because NPE based detergents enter
the municipal system and enhance their participation
in endocrine disruption in the aquatic ecosystem. Fi-
nally, levels in the environment that are capable to
induce reproduction impairment were determined.
The acute toxicity with 96hLC50 calculated for NP
has been determined for at least 22 different species
of fish with reported 96hLC50s ranging from 17 to
3,000 μg/L, although most of the values ranges from
100 to 300 μg/L (Gray and Metcalfe 1997, Dwyer et al.
1999). Chronic toxicity values (NOEC) are as low as
6μg/L in fish and 3.7 μg/L in invertebrates (Servos
1999) and LOEC was at nominal dose of 4 μg/kg body
weight of AP for the effects of timing of puberty, and
20 μg/kg body weight of AP for the delay effect in
mature fish (Meier et al. 2011). Likewise, Bellingham
et al. (2010) demonstrated that developing foetus is
sensitive to environmentally relevant concentrations
of EDCs that may have serious impact on the future
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development of the reproductive system. The body
burden corresponds to water exposure in the ng/L
concentration range and this is 100times lower than
the APs concentrations (μg/L) that have previously
been reported to induce vitellogenin in male or juven-
ile fish and have been documented in surface water all
over the world (Meier et al. 2011). Further indicators
of EDCs toxicity could be sex steroid-binding protein
determination (Tollefsen and Nilsen 2007) or behav-
ioural changes assessment (Xia et al. 2010). It was
demonstrated that APs and alkylated non-phenolics
are able to bind successfully to the rainbow trout sex
steroid-binding protein. Unfortunately, the most ubi-
quitous AP like NP was not proved to bind to rainbow
trout sex steroid-binding protein (Tollefsen and Niel-
sen 2007). On the other hand, measurements of sev-
eral behavioural endpoints, especially locomotors ac-
tivity, aggression displays and group preference, may
provide an effective assessment of NP in the aquatic
ecosystem but this measurement has some limitations
like biotic and abiotic factors in field conditions
(Spence et al. 2008, Xia et al. 2010). Other studies are
needed to establish more sensitive and specific
markers of APs detrimental activity but, as well
documented, chemicals exist as complex mixtures in
the environment and groups of compounds may cause
mixed toxicity even though each compound is present
below its threshold for effect (Silva et al. 2002).
Conclusions
Recently provided information demonstrated seri-
ous damaging impact of EDCs on aquatic organisms,
in particular. Besides natural/synthetic hormones,
PCBs, dioxins, some insecticides and bisphenol-A,
APs and APEs constitute very potent EDCs as well as
persistent substances with multiple toxicity, deleteri-
ous effect and bioaccumulative features. In spite of
the fact that their use was reduced in Europe, USA
and Japan in last few years, they are still very popular
in some other countries and contamination of the en-
vironment with these chemicals is of substantial con-
cern in the whole world. Not only sex in reversal fish
(Sumpter 1998), but also altered social and reproduc-
tive behaviour in birds (Fox 2001), impaired penile
development in alligators (Guillette et al. 1999),
cryptorchidism and infertility in the Florida panther
(Facemire et al. 1995) are virtual consequences of an
increase in EDCs in the environment. These findings
have implications for the study of human conditions
that are suspected to be caused by environmental es-
trogens, such as undescended testis, breast cancer and
a decrease in sperm counts and quality during last 50
years (Wolf et al. 1993). Furthermore, APs have ap-
peared to be able to interfere with the regulatory sys-
tems of different types of cells by several mechanisms
and thus are not compounds of single harmful effect.
It is an alarming finding that NP is ubiquitous in food-
stuffs at concentrations ranging from 0.1 to 19.4 μg/kg
with an estimated daily intake of 7.5 μg/day for an
adult (Guenther et al. 2002). The ubiquity of this type
of pollutants in the environment supports a need for
greater knowledge of APs bioaccumulation process
and their human or animal health alteration potential.
Options of waste water treatment process improve-
ment and/or also APs/APEs possible substitution in
the industrial use with another type of substances are
still animated issues for further investigation.
Acknowledgements
We thank MVDr. Veronika Simova for manu-
script improvement and English correction.
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