Toxicology Letters 217 (2013) 50–58
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Toxicology Letters
journal homepage: www.elsevier.com/locate/toxlet
Biological impact of phthalates
Rishikesh Mankidy a,∗ , Steve Wiseman a , Hong Ma a , John P. Giesy a,b,c,d
a
Toxicology Center, University of Saskatchewan, Saskatoon, SK, Canada
Department of Veterinary Biomedical Sciences, University of Saskatchewan, Saskatoon, SK, Canada
c
Zoology Department, Center for Integrative Toxicology, Michigan State University, East Lansing, MI, USA
d
Department of Biology and Chemistry and State Key Laboratory for Marine Pollution research, City University of Hong Kong, Kowloon, Hong Kong, China
b
h i g h l i g h t s
Investigated the biological impact of phthalates DEHP, DEP, DBP and BBP.
Phthalates differing in physicochemical properties have similar endpoints.
Phthalates simultaneously affect multiple cellular targets.
Demonstrated the need for the simultaneous assessment of multiple endpoints.
a r t i c l e
i n f o
Article history:
Received 19 October 2012
Received in revised form
27 November 2012
Accepted 28 November 2012
Available online 7 December 2012
Keywords:
Phthalate
Oxidative stress
Caspase
Embryotoxicity
a b s t r a c t
Esters of phthalic acid are chemical agents used to improve the plasticity of industrial polymers. Their
ubiquitous use in multiple commercial products results in extensive exposure to humans and the environment. This study investigated cytotoxicity, endocrine disruption, effects mediated via AhR, lipid
peroxidation and effects on expression of enzymes of xenobiotic metabolism caused by di-(2-ethy hexyl)
phthalate (DEHP), diethyl phthalate (DEP), dibutyl phthalate (DBP) and benzyl butyl phthalate (BBP) in
developing fish embryos. Oxidative stress was identified as the critical mechanism of toxicity (CMTA)
in the case of DEHP and DEP, while the efficient removal of DBP and BBP by phase 1 enzymes resulted
in lesser toxicity. DEHP and DEP did not mimic estradiol (E2 ) in transactivation studies, but at concentrations of 10 mg/L synthesis of sex steroid hormones was affected. Exposure to 10 mg BBP/L resulted
in weak transactivation of the estrogen receptor (ER). All phthalates exhibited weak potency as agonists of the aryl hydrocarbon receptor (AhR). The order of potency of the 4 phthalates studied was;
DEHP > DEP > BBP » DBP. The study highlights the need for simultaneous assessment of: (1) multiple cellular targets affected by phthalates and (2) phthalate mixtures to account for additive effects when multiple
phthalates modulate the same pathway. Such cumulative assessment of multiple biological parameters
is more realistic, and offers the possibility of more accurately identifying the CMTA.
© 2012 Elsevier Ireland Ltd. All rights reserved.
1. Introduction
Phthalates, which are esters of phthalic acid are primarily used
to enhance plasticity of industrial polymers (Sears and Darby,
1982). They are used in a number of consumer end products such
as toys, paints, adhesives, lubricants, packaging and building materials, personal care items, electronics, medical devices, and are an
unavoidable part of modern life (Horn et al., 2004; Shea, 2003). A
recent study estimated that 11 billion pounds of phthalates were
produced worldwide every year (Lowell Center for Sustainable
∗ Corresponding author at: Toxicology Center, University of Saskatchewan, 44
Campus Drive, Saskatoon, S7N 5B3 Canada. Tel.: +1 306 966 8733/4680;
fax: +306 966 4796.
E-mail address: mankidy@gmail.com (R. Mankidy).
0378-4274/$ – see front matter © 2012 Elsevier Ireland Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.toxlet.2012.11.025
Prroduction, 2011). While these plasticizing agents impart beneficial properties to plastics, they are not bound to the polymer by
a covalent linkage which makes them susceptible to leaching from
the matrix (Fromme et al., 2012). Once released into the atmosphere, they have the potential for long-range transport, eventually
entering the food chain (Federal Environmental Agency, 2007).
Structures and physical properties vary among phthalates,
which influences their chemodynamics in the environment
(Staples, 1997). Phthalates with lesser molecular weights, such
as diethyl phthalate (DEP) have greater bioaccumulation factors
(BAFs), while larger phthalates such as di-(2-ethy hexyl) phthalate (DEHP) tend to have lesser BAFs (Staples, 1997). Despite their
greater water–octanol partitioning coefficients (Kow ), phthalate
esters do not have greater biomagnification factors (BMFs) such
that concentrations of phthalates are not greater in higher trophic
levels of aquatic food webs (Gobas F 2003). This is probably due
R. Mankidy et al. / Toxicology Letters 217 (2013) 50–58
to the fact that phthalates have a fairly short half-life in the environment with greater than 50% degradation occurring within 28
days (Staples, 1997), primarily via photo-degradation. Furthermore, phthalates such as DEHP, are readily bio-transformed and
excreted (Barron, 1995), which results in lesser bioaccumulation.
In humans, phthalates have been detected in matrices such
as blood, urine, saliva, amniotic fluid, breast milk and cord blood
(Latini et al., 2003b; Main et al., 2006; Silva et al., 2004a,b, 2005).
The major pathway of exposure to phthalates is the oral route,
though inhalation and dermal absorption may play a significant
role in exposure (Adibi et al., 2003; Rudel et al., 2003) Infants
and toddlers are the most vulnerable receptors because: (1) they
exhibit more hand-to-mouth activity, and (2) consume the most
food as a percent of their body weight (wargo et al., 2008). The situation is exaggerated by the fact that ubiquitous phthalates such
as DEHP, which have been classified as endocrine disrupting chemicals (EDCs), exhibit an oral absorption factor of 0.55 (Rhodes et al.,
1986), and affect the most vulnerable receptors at critical stages of
development.
Phthalates have been reported to affect multiple biochemical
processes in humans and wildlife. These include effects on reproduction, damage to sperm (Rozati et al., 2002), early onset of
puberty in females (Wolff et al., 2010), anomalies of reproductive
tract (Desdoits-Lethimonier et al., 2012), infertility (Rozati et al.,
2002; Tranfo et al., 2012) and adverse outcomes of pregnancy
(Latini et al., 2003a; Whyatt et al., 2009), to neurodevelopment
(Engel et al., 2010; Miodovnik et al., 2011) and allergies (Bornehag
et al., 2004; Jaakkola et al., 2000). Because humans and wildlife can
be exposed simultaneously to several phthalates any assessment of
the risks posed by phthalates needs to consider combined effects of
all of the phthalates in mixtures. To do this requires knowledge of
the critical mechanisms of toxic action (CMTA) of each phthalate. It
is only by this knowledge that it can be determined how to aggregate the exposures. While effects of phthalates have been observed
and described, few studies have been conducted to determine the
CMTA. The CMTA is the endpoint that is not only most severe but
that which occurs at the least concentration. For instance, if individual phthalates had similar mechanisms of toxic action, with
different potencies, a toxic units approach could be applied. If the
mechanisms are different, then the effects of the various phthalates
would be better assessed by considering them individually.
In the study presented here, phthalates of different molecular sizes were investigated. A greater molecular weight phthalate,
DEHP, a lesser molecular weight phthalate, DEP, and two phthalates with intermediate molecular weights, dibutyl phthalate (DBP)
and butyl benzyl phthalate (BBP) were studied. The chemical and
physical properties of these phthalates vary and they also have different uses in manufacturing and consumer products. Effects of
these four phthalates were assessed both in vitro and in vivo to
elucidate CMTAs such as endocrine disruption, oxidative stress,
aryl hydrocarbon receptor (AhR) receptor-mediated effects, and
the expression enzymes of phase I xenobiotic metabolism. In conclusion we present a scheme which grades the 4 phthalates on
the overall potential effects on biological systems and allows for
comparison between phthalates with respect to each individual
biological end point.
2. Materials and methods
2.1. Cytotoxicity
MVLN cells were propagated in DMEM/F-12 media containing 10% FBS at 37 ◦ C,
5% CO2 . Cytotoxicities of phthalates were determined by exposing 8 × 104 MVLN
cells to DEHP, DEP, DBP or BBP (Sigma–Adlrich, St. Louis, MO) for a period of 24 h.
WST-1 reagent (Roche Applied Science, Indianapolis, IN) was used to determine
metabolically active cells at the end of the incubation period according to the manufacturer’s recommendations.
51
2.2. Caspase-3 assay
Active caspase-3 was assayed in MVLN cells at concentrations that resulted in
cell death. 1.5 × 106 cells were exposed to phthalates for 3 h in a 6-well plate at concentrations which exhibited cytotoxicity at 24 h. Cells were harvested, lysed, and
active Caspase-3 was quantified using EnzChek Caspase-3 Assay Kit (Life Technologies, Carslbad, CA) according to the manufacturer’s recommendations. Amount of
fluorescence generated was normalized by the amount of protein (␮g) in the extract.
2.3. Endocrine disruption
2.3.1. H295R steroidogenesis assay
The H295R Steroidogenesis assay has been validated and is an established system routinely used as a tier-I screen for steroidogenic effects of test chemicals
(Hilscherova et al., 2004; Sanderson et al., 2000; Zhang et al., 2005). This cell line
has the full complement of enzymes required sex steroid biosynthesis (Gracia et al.,
2006; Hecker et al., 2006, 2007, 2011), and hence is a good model to study disruptions in steroidogenesis. H295R cells, purchased from ATCC (Manassas, VA), were
propagated in DMEM/Hams F-12 medium containing 10%FBS at 37 ◦ C, 5% CO2 . Cells
were exposed to phthalates under conditions previously described (Hecker et al.,
2006). Following exposure to phthalates for 48 h, conditioned media was collected
and concentrations of 17-␤ estradiol (E2 ) and testosterone (T) in culture media were
determined by use of ELISA (Cayman Chemical, Ann Arbor, MI) according to the
manufacturer’s recommendations.
2.3.2. Estrogen receptor transactivation assay
MVLN cells, derived from the MCF-7 breast cancer cell line, are engineered to
express luciferase under the control of estrogen responsive elements (Demirpence
et al., 1993; Pons et al., 1990). Transactivation of the estrogen receptor (ER) by phthalates was determined by exposing 3 × 104 cells to individual phthalates for 48 h in
a 96-well plate. Following exposure, cells were lysed and luminescence quantified
by use of SteadylitePlus reagent (Perkin-Elmer, Waltham, MA). Potencies of individual phthalates as agonists of the ER were determined by lumniscence caused by
standard concentrations of E2 (Sigma–Aldrich, St. Louis, MO)
2.4. Aryl hydrocarbon receptor transactivation assay
The assay to determine the potential for phthalates to activate the AhR was
conducted as described (Garrison et al., 1996) with a few modifications. H4IIE cells
were propagated in DMEM containing 10%FBS at 37 ◦ C, 5% CO2 . 5 × 104 H4IIE cells
were exposed to the phthalates for 24 h. Cells were harvested and luminescence
quantified using SteadylitePlus reagent (Perkin-Elmer, Waltham, MA) according
to the manufacturer’s recommendations. Luminescence derived from the exposure to phthalates was compared with that obtained from a standard curve for
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD; Wellington Laboratory, Guelph, ON) to
determine TCDD equivalents.
2.5. Fathead minnow experiments
2.5.1. Embryotoxicity
Fathead minnows (Pimephales promelis) were cultured in 200 L tanks in the
Aquatic Toxicology Research Facility (ATRF) in the Toxicology Center at the University of Saskatchewan. Tanks for breeding fish were maintained as described
previously (He et al., 2012). Fertilized embryos were collected within 1 h post fertilization (hpf) and were pooled in a Petri-dish containing control water. Eggs were
rinsed 3 times and unfertilized eggs discarded. One exposure replicate consisted of
10–15 eggs in each well of a 6-well plate containing 2 mL of control water or containing phthalates. 50% of the volume (1 mL) was replaced daily with fresh test solutions.
Exposures were performed at 25 ± 1 ◦ C with a 16/8 h light/dark photoperiod. Daily
enumeration of live and dead embryos was made prior to the 50% water renewal
and dead eggs were discarded. Exposures were terminated 96 hpf and cumulative
percent mortality determined. Live embryos were collected, flash frozen in liquid
nitrogen, and stored at −80 ◦ C until needed for determination of lipid peroxidation
and abundances of transcripts of target genes. All exposures were conducted on 5
separate batches of eggs.
2.5.2. Lipid peroxidation
Fertilized eggs exposed to phthalates for 96 h were used to determine the degree
of lipid peroxidation by use of the Lipid hydroperoxide assay kit (Cayman Chemical,
Ann Arbor, MI). Wet mass of pooled embryos from each replicate was determined
prior to extraction of lipids. Lipids were extracted with 500 ␮L chloroform containing 1% triton X-100 and directly used in the assay according to recommendations
of the manufacturer. Amount of lipid peroxide (nmol) was quantified by reading
the absorbance at 500 nm. Data were normalized by the wet weight (mg) of tissue.
2.5.3. Molecular studies
Total RNA was extracted from 5 embryos in each treatment group by use of
the Qiagen RNeasy Plus Mini Kit according to the manufacturer’s protocol (Qiagen, Mississauga, ON, Canada). The concentration of RNA was determined with a
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R. Mankidy et al. / Toxicology Letters 217 (2013) 50–58
Table 1
List of fathead minnow genes with sequences of primers.
Symbol
Gene name
Forward primer
Reverse primer
Primer efficiency
GST
SOD
CAT
CYP1A
CYP3A
CYP2K19
AR
ER␣
ER␤
RPL8
Glutathione-S-transferase
Superoxide dismutase
Catalase
Cytochrome p450 1A
Cytochrome p450 3A
Cytochrome p450 1A
Androgen receptor
Estrogen receptor alpha
Estrogen receptor beta
Ribosomal protein L8
CCGGCAAGAGCTTCACCAT
CCAGACATGTCGGAGACCTT
TTATCAGGGATGCGCTTCTGT
CCTGCAGGGAGAACTGAG
CGACGAGACCTTCCCAAAT
CACAAGGTCCCTCCCTTACA
CAACGCGTCTAAATCCCATT
CGGTGTGCAGTGACTATGCT
CGTTTTGGCATAACCATGTG
CTCCGTCTTCAAAGCCCATGT
AGTGAAGTCGTGGGAAATAGGC
ATGGAATGTTGCCCTGAGAG
TTCACATGAGTCTGCGGATTTC
TCGACGTACAGTGAGGGA
GTTTTCTTGCAGACCCGTT
CAAAACCAGAAGCAACAGCA
TGTTCGAACTGACACGAAGC
CTCTTCCTGCGGTTTCTGTC
TGCTGTCAGACTTCCGAATG
TCCTTCACGATCCCCTTGATG
1.9
1.96
1.91
1.85
1.89
1.85
2.04
2.06
1.99
2.01
NanoDrop ND-1000 Spectrophotometer (Nanodrop Technologies, Welmington, DE,
USA) and was stored at −80 ◦ C until required. First-strand cDNA was synthesized
from 1 ␮g of each RNA sample by use of a QuantiTect Reverse Transcription Kit
(Qiagen) according to the manufacturer’s protocol. Primers were designed using
Primer 3 software and based on sequences obtained by Illumina RNA sequencing
(unpublished data) or sequences available through the NCBI database. Nucleotide
sequences of primers and the biological functions of the target transcripts are given
(Table 1).Quantitative real-time PCR (qPCR) was performed on an ABI 7300 RealTime PCR System in 96-well PCR plates (Applied Biosystems, Foster City, CA, USA). A
PCR reaction mixture for 1 reaction contained 10 ␮L of Quantitect SYBR Green PCR
reagent (Qiagen) and optimized volumes of cDNA and primers (Invitrogen, Carlsbad, CA, USA). The PCR reaction mix was denatured at 95 ◦ C for 10 min before the
first PCR cycle. The thermal cycle profile was: denaturizing for 15 s at 95 ◦ C and
annealing and extension for 1 min at 60 ◦ C for a total of 40 PCR cycles. The qPCR
cycle was followed by a melt curve analysis to confirm homogeneity of the PCR
product. Changes in abundances of transcripts of target genes were quantified by
normalizing to the expression of housekeeping gene ribosomal protein L8 (RPL8).
The efficiencies of PCR reactions were determined using serial dilutions of cDNA
(Table 1).
2.6. Statistical analyses
Data was analyzed using GraphPad Prism 4.0. Differences in the means were
determined by ANOVA followed by a post hoc Dunnett’s t test.
3. Results
3.1. Cellular toxicity
Of the four phthalates, BBP exhibited the greatest toxic potency
toward MVLN cells with significant toxicity observed at 1.0 mg
BBP/L (P < 0.01) (Fig. 1). This was followed by DEHP that caused
cytotoxicity at 10 mg/L (P < 0.01). DEP and DBP caused significant
toxicity only at concentrations of 100 mg/L or greater (P < 0.01).
The cause of the observed toxicity was investigated to determine
if it was the result of initiation of an apoptotic response. Active
Fig. 1. Cytotoxicity of phthalates. MVLN cells were exposed to phthalates (1.0 × 10−4
to 1.0 × 103 mg/L) or to solvent control (DMSO) for 24 h. Metabolically active cells
were enumerated using WST-1 reagent. Asterisk indicates significant mortality
(P < 0.01) compared to control conditions. Results represent the mean of 4 independent experiments; error bars indicate standard errors of the mean (SEM).
caspase-3, which is an indicator of commitment of cells to programmed cell death, was greater in MVLN cells exposed to 10 mg
DEHP/L, 100 mg DEP/L, or 1.0 mg BBP/L (P < 0.01); concentrations
of active caspase-3 were 7-, 4.7- and 5.7-fold greater than that in
the control cells, respectively (Fig. 2). Exposure of cells to 100 mg
DBP/L did not initiate the apoptotic cascade. Camptothecin served
as a positive control for caspase-3 activation. A 5-fold activation
over un-induced cells was observed at a concentration of 17 mg/L.
3.2. Activation of AhR signaling by phthalates
All phthalates investigated in this study exhibited weak potency
as agonists of the AhR, which corresponded to (1.5–2) × 10−7 mg/L
TCDD equivalents (Fig. 3). DEHP, DEP, and BBP were more
potent agonists exhibiting ∼4–5 fold increase in expression of
the reporter gene over control conditions. Effects were significant even at the least concentration of 0.01 mg/L (P < 0.01), with
no further increase in the signal despite a 1000-fold increase
in concentrations (0.01–10 mg/L) of phthalates. DBP exhibited
concentration-dependent production of AhR mediated light units;
while 0.01 mg DBP/L resulted in expression of luciferase activity 2fold greater than that of the control (P < 0.01), a maximum response
which was 7-fold greater than control was observed at 10 mg DBP/L,
(P < 0.01).
3.3. Phthalates as endocrine disruptors
DEHP and DEP were the only phthalates that significantly
affected concentrations of E2 in media. Exposure to 10 mg DEHP/L
resulted in 4-fold greater concentration of E2 (P < 0.01), while exposure to 10 mg DEP/L resulted in 2.3-fold greater concentration of E2
compared to solvent control (P < 0.01), (Fig. 4). DBP or BBP did not
alter concentrations of E2 in media. DEHP had the greatest effect on
concentration of T in media. Exposure to 0.1 mg DEHP/L resulted
Fig. 2. Apoptosis triggered by phthalates. Caspase-3 activation in MVLN cells
exposed to the lowest phthalate concentration which resulted in cell death or to
control conditions (C); 10 mg/L DEHP, 100 mg/L DEP, 100 mg/L DBP and 1 mg/L BBP.
Camptothecin (17 mg/L) exposure served as a positive control for apoptosis. Asterisk
indicates a value significantly different than control (P < 0.01). Data are an average
of 3 independent experiments; error bars indicate standard deviation.
R. Mankidy et al. / Toxicology Letters 217 (2013) 50–58
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Fig. 3. TCDD-like activity in rat hepatoma H4IIE-luc cells exposed to phthalates. Data represent TCDD equivalents (mg/L) observed following exposure to DEHP, DEP, DBP,
BBP or DMSO solvent control (C) and are an average of four replicates. All concentrations tested were significantly higher than control cells denoted by an asterisk (P < 0.01).
Error bars indicate SEM.
Fig. 4. Concentration of E2 in conditioned media of H295R cells. Values represent the average concentration of E2 (mg/L) in the conditioned medium following exposure to
phthalates (DEHP, DEP, DBP or BBP) or to the solvent control (C). Forkoslin (F) and prochloraz (P) served as controls which up regulate or down regulate steroidogenesis
pathway, respectively. * Significantly greater than control (P < 0.01), ** Significantly less than control (P < 0.01). Data are representative of 3 independent experiments; error
bars indicate the SEM.
in 70% less T (P < 0.01) in the medium relative to that of controls
(Fig. 5). Similarly, exposure to 0.1 mg DEP/L resulted in a 60% lesser
concentration of T (P < 0.01) and exposure to 0.1 mg DBP/L resulted
in 50% lesser concentration of T (P < 0.01) than that in the control. BBP did not affect synthesis of T at any of the concentrations
tested. Exposure of H295R cells to 4.1 mg forskolin/L up-regulated
steroidogenesis, which resulted in a 3-fold greater mean concentration of E2 (Fig. 4) and 1.5-fold greater mean concentration of T
(Fig. 5) in conditioned medium. Cells were exposed to 1 × 10−5 mg
prochloraz L as a negative control which antagonized synthesis of
E2 and T; concentrations of these two steroid hormones were 47%
and 41% less than those of controls, respectively.
Fig. 5. Concentration of T in conditioned media of H295R cells. Values represent the average concentration of T (mg/L) in the conditioned medium following exposure to
phthalates (DEHP, DEP, DBP or BBP) or to the solvent control (C). Forkoslin (F) and prochloraz (P) served as controls which up regulate or down regulate steroidogenesis
pathway respectively. * Significantly greater than control (P < 0.01), ** Significantly less than control (P < 0.01). Data are representative of 3 independent experiments; error
bars indicate the SEM.
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Fig. 6. ER-mediated effects of phthalates. E2 equivalents (mg/L) observed in MVLN cells exposed to DEHP, DEP, DBP, BBP, or to DMSO control (C). Asterisk indicates a value
significantly different than control (P < 0.001). Data are an average of 4 independent experiments; error bars represent the SEM.
Potency of phthalates as agonists of the estrogen receptor (ER) as
determined by use of the MVLN transactivation reporter assay was
observed at a concentration range of 0.01–10 mg/L, a concentration
range which did not exhibit any cellular toxicity. In this assay, BBP
was the only chemical that was a weak ER agonist; 10 mg BBP/L
generated light units equivalent to 6.8 × 10−6 mg/L E2 (P < 0.001)
(Fig. 6). DEHP, DEP or DBP did not elicit ER-mediated responses
that were greater than those in the control.
Effects of the four phthalates on survival of fertilized eggs of fathead minnows were determined at concentrations which did not
affect proliferation of cells in vitro. Survival and development of
fathead minnow embryos were monitored during a 96-h exposure
to each of the phthalates (Fig. 7). DEHP and DEP were the most
potent based on mortality. Exposure to 1 mg DEHP/L resulted in
30% mortality while exposure to 10 mg DEP/L caused 52% mortality compared to 3% mortality in the control population (P < 0.05).
Exposure to DBP or BBP at 1 mg/L did not cause significantly greater
mortality of embryos compared to the controls. Oxidative stress
was assessed in developing embryos following exposure to the
phthalates until 96 hpf. Exposure to 1 mg DEHP/L, 10 mg DEP/L, and
1 mg BBP/L caused a 2-fold greater lipid peroxidation in membranes
of developing embryos (P < 0.05), while 1 mg DBP/L exposure failed
to generate lipid peroxide levels different from those observed in
the controls (Fig. 8).
Molecular mechanisms of toxic action of phthalates were investigated by assessing transcriptional changes in developing embryos
following exposure to phthalates until 96 hpf. Expression of transcripts of genes that encode enzymes involved in metabolism
of xenobiotics, including CYP1A-aryl hydroxylase, CYP2K19monooxygenase, and CYP3A-monooxygenase were assessed. In
addition, abundances of transcripts of markers of oxidative stress
such as glutathione-S-transferase (GST), superoxide dismutase
(SOD), catalase (CAT), and indicators of endocrine-related effects,
such as androgen receptor (AR), estrogen receptors (ER-␣ and
ER-␤) were quantified. No differences in expression of aryl hydroxylase CYP1A following exposure to any of the phthalates were
observed (Fig. 9A). Only a small effect on expression of the
monooxygenase enzyme CYP3A was observed after exposure to
10 mg DEP/L (1.5-fold elevation), though it was significantly elevated (2.5-fold elevation) with exposure to 1 mg BBP/L (P < 0.05).
The expression of monooxygenase enzyme CYP2K19 was 2-fold
greater than control after exposure to 10 mg DEP/L and 1 mg BBP/L,
and 8-fold greater than the control after exposure to 1 mg DBP/L
(Fig. 9A). Exposure to 1 mg DEHP/L had no effect on expression
of any of the phase 1 enzymes in the embryos. Neither 1 mg
DEHP/L nor 10 mg DEP/L affected expression of mRNA of SOD,
GST or CAT genes, which code for enzymes responsible for mitigating oxidative stress in embryos (Fig. 9B). Alternatively, 1 mg
DBP/L caused a 2.5-fold greater expression of mRNA for SOD and
14-fold greater expression of mRNA for CAT relative to those
of controls, while exposure to 1 mg BBP/L resulted in 1.5- and
Fig. 7. Fish embryo toxicity. Embryo survival following exposure to phthalates (1 mg
DEHP/L, 10 mg DEP/L, 1 mg DBP/L or 1 mg BBP/L) for 96 hpf presented as percentage
of control (C). Each bar represents the average of five independent exposures; each
exposure contained ∼15 embryos. Asterisk indicates values that are significantly
different from solvent controls (P < 0.05). Error bars represent the SEM.
Fig. 8. Lipid peroxidation in fish embryos. Extent of lipid peroxidation observed
in developing fish embryos as a percentage of solvent control (C). Embryos were
exposed to 1 mg DEHP/L, 10 mg DEP/L, 1 mg DBP/L, 1 mg BBP/L or to DMSO solvent. Values are an average of two independent exposures; each exposure contained
∼15 embryos. Asterisk indicates values that are significantly different from solvent
controls (P < 0.05). Error bars represent the SEM.
3.4. Effects of phthalates on fathead minnow embryos
R. Mankidy et al. / Toxicology Letters 217 (2013) 50–58
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Fig. 9. Transcription profile of genes involved in xenobiotic metabolism (panel A), mitigation of oxidative stress (panel B) and sex steroid receptors (panel C) in developing
fathead minnow embryos exposed to phthalates (1 mg DEHP/L, 10 mg DEP/L, 1 mg DBP/L, 1 mg BBP/L)for 96 h or to control conditions (C). Data represent the average mRNA
expression in 5 individual embryos, and are normalized to the expression of housekeeping gene RPL8. Asterisk indicates values significantly different from DMSO controls
(P < 0.05). Error bars represent the SEM of 5 independent replicate exposures.
9-fold greater expression of mRNA for SOD and CAT, respectively
(P < 0.05). Expression of mRNA for GST was not affected by any of
the phthalates. None of the concentrations of any of the phthalates altered in vivo expression of mRNA of the ER ␣ or ␤, while
expression of mRNA of AR was slightly greater in cells exposed to
1 mg DEHP/L (2.7-fold), 1 mg DBP/L (1.5 fold), but reached statistical significance (P < 0.05) only for 1 mg BBP/L treatment (2-fold)
(Fig. 9c).
4. Discussion
Soon after the advent of their use in manufacturing, questions were raised about potential effects of phthalates on health
of humans and effects on wildlife. While a number of studies have
been conducted, these were often ectopic, involving assessments of
a single phthalate. Also, most studies have focused on specific target
processes in vivo such as cell death, sex differentiation, reproductive defects, oxidative stress, etc. (Gray, Jr. et al., 2000; Howdeshell
et al., 2007; Sun et al., 2012; Zhao et al., 2012) ignoring the fact
that phthalates can affect multiple targets within the cell simultaneously. In this report a comparison of effects and potencies are
presented for four phthalates, DEHP, DEP, DBP, and BBP which differ
in molecular weight, physical, and chemical properties. An initial
assessment of in vitro toxicity of the phthalates was conducted,
and the data generated used to guide the exposure concentrations
in subsequent cellular bioassays.
DEHP and DEP, despite their differences in molecular weights
and physico-chemical properties, were only moderately toxic
to cells in vitro and in embryos of fathead minnows. Mortality
of embryos could be attributed in part to oxidative stress, since
significantly greater products of lipid peroxidation were observed
in membranes of these developing embryos. Additional evidence
supporting this mode of toxic action is provided by the observation
that DEHP and DEP failed to up-regulate expression of mRNA of the
enzymes SOD, GST or CAT, which are induced as counter measures
to the adverse effects of oxidative stress. These data corroborate
recent reports in the literature (Erkekoglu et al., 2011, 2012; Kang
et al., 2010; Rosado-Berrios et al., 2011) and suggest that oxidative
stress is a CMAT for DEHP and DEP. Although activity of caspase-3
was not measured directly in embryos, the results of in vitro experiments suggested that apoptosis was occurring and was a plausible
mechanism for the observed toxicity of DEHP and DEP to fathead
minnow embryos. These observations are consistent with results of
previous studies that suggested a link between oxidative stress and
programmed cell death (Carvour et al., 2008) and places apoptosis
downstream of events generating oxidative stress in cells.
Despite the 100-fold difference with respect to in vitro toxicity,
DBP and BBP exhibited no differences in potency for mortality of
embryos of fathead minnows. The cellular and molecular investigations revealed that exposure to either phthalate elevated transcript
abundances of SOD and CAT. In the case of exposure to DBP, the
antioxidant counter measure was sufficient to mitigate oxidative
stress in embryos since the magnitude of lipid peroxidation was
indistinguishable from that of controls. However, with exposure
to BBP, the antioxidant response was not sufficient to prevent
oxidative damage in the developing embryos resulting in lipid peroxidation greater than that of the controls. Despite the greater lipid
peroxidation, mortality was not affected. This result suggests the
presence of additional pathways responsible for the death of the
embryos.
56
R. Mankidy et al. / Toxicology Letters 217 (2013) 50–58
The H4IIE-luc transactivation reporter assay uses rat hepatoma
cells stably transfected with a reporter for chemicals which act
via the AhR (El-Fouly et al., 1994; Hilscherova et al., 2000). The
H4IIE-luc assay has been extensively used for the detection of
polychlorinated dibenzo-p-dioxins, polychlorinated biphenyls and
polycyclic aromatic hydrocarbons (Garrison et al., 1996; Murk et al.,
1996; Sanderson et al., 1996), and it is also responsive to chemical
stressors which are distinctly different from the above mentioned
chemicals (Denison and Nagy, 2003). Exposure of H4IIE-luc cells to
phthalates resulted in similar responses in the transactivation assay
consistent with reported effects mediated by the AhR (Kruger et al.,
2008). These effects corresponded to 2 × 10−7 mg/L TCDD equivalents. Despite the similar magnitude of transactivation observed
in vitro, differences were evident at the molecular level. DEHP and
DEP failed to affect expression of mRNA of phase I enzymes in
embryos. Failure to up regulate these enzymes could explain the
lesser viability observed with DEHP and DEP exposure. Exposures to
either DBP or BBP affected xenobiotic metabolism pathways though
the effects were mediated via expression of different key phase 1
enzymes of xenobiotic metabolism. Despite these differences, both
pathways appear to be equally efficient in processing/clearing the
xenobiotic phthalate; this resulted in mortality similar to controls.
These data collectively suggest that the CMTA for the phthalates in developing fish embryos were different among phthalates.
Oxidative stress and/or failure to efficiently metabolize the xenobiotic toxin appear to be the modes of toxic action in fish larvae
following exposure to DEHP and DEP. Failure to observe any discernible difference in viability following exposure to DBP and BBP
exposure could be explained by the fact that the phase 1 monooxygenases CYP2K19 and CYP3A efficiently removed DBP and BBP,
respectively.
Several studies with phthalates have identified endocrine disruption as cause of major concern (Desdoits-Lethimonier et al.,
2012; Giribabu et al., 2012; Gray, Jr. et al., 2000), though often these
studies have not elucidated the mechanism of action with respect to
endocrine disruption. The studies outlined in this report show varied modes of endocrine disruption amongst the phthalates. DEHP
and DEP targeted steroid biosynthesis pathways resulting in greater
production of E2 with a concurrent reduction in concentration
of T. BBP on the other hand displayed weak estrogen mimicking
potency in the MVLN transactivation assays similar to previous
reports (Harris et al., 1997). This was accompanied by a small,
yet significant elevation in the expression of AR in the developing
embryos which could be explained as compensatory up regulation
in the case of anti-androgenic compounds, effects similar to those
reported (Sohoni and Sumpter, 1998). DBP was the only phthalate that showed no effects with respect to any of the endocrine
disrupting end points in the study.
In conclusion, a grading scheme for assessing the overall biological effects of phthalates is presented here. This process gives
a crude quantitative measure of the overall biological effects of a
phthalate relative to other compounds in the study, which could
be expanded and used for comparing multiple related or unrelated
chemicals. According to the data presented here, DEHP exhibited
the greatest biological effects with a score of 25, followed by DEP
(21), BBP (17), and lastly DBP (6) (Table 2).
Risk assessments with individual phthalates have concluded
that exposure to receptors is often several orders of magnitude
below the toxicological threshold (Kamrin, 2009). However, in reality a receptor is exposed to multiple phthalates simultaneously. The
fact that multiple phthalates can act through the same biological
pathway in vivo further compounds their impact on the organism
and highlights the importance of investigating the effects of exposure to mixtures of phthalates (Howdeshell et al., 2008). Lastly, the
heterogeneity of response to phthalates in the study adds another
layer of complexity demonstrating the need to investigate multiple
Table 2
Compilation of effects of phthalates in bioassays.
Assay
DEHP
DEP
DBP
BBP
Cytotoxicity
Caspase 3 activation
Fish survival
Toxicity
E2 steroidogenesis
T steroidogenesis
E2 (receptor-mediated)
Endocrine disruption
AhR mediated
Lipid peroxidation
Anti-oxidant defense
Oxidative stress
Cumulative effect
3
4
3
10
4
4
1
9
3
4
−1
3
25
2
2
4
8
3
3
1
7
3
4
−1
3
21
2
1
1
4
1
1
1
3
2
1
−4
−3
6
4
3
1
8
1
1
2
4
4
4
−3
1
17
The extent of the effects observed with respect to the corresponding end-point of
each assay was graded on a scale from 1–4; 1-having no effect in the assay to 4showing maximum effect. A negative value was used in the case of assays where the
greatest impact was a favorable response, as in the case of elevation of antioxidant
status.
endpoints simultaneously while assessing human health and ecological health. A number of studies label phthalates as innocuous
agents that do not persist in the environment (Staples, 1997) and
exhibit very little tendency to bioaccumulate (Gobas F 2003). However, it is important to note that the massive rate of synthesis of
phthalates could supersede the natural rate of removal, leading to
an eventual accumulation of these chemicals in the environment
(aquatic sediments, particulate material) and ultimately in humans.
Conflict of interest statement
None declared.
Acknowledgements
The research was supported by a Discovery Grant from the Natural Science and Engineering Research Council of Canada (Project #
406497). The authors wish to acknowledge the support of an instrumentation grant from the Canada Foundation for Infrastructure.
Prof. Giesy was supported by the Canada Research Chair program,
an at large Chair Professorship at the Department of Biology and
Chemistry and State Key Laboratory in Marine Pollution, City University of Hong Kong, The Einstein Professor Program of the Chinese
Academy of Sciences.
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