Acute Toxicity of Diisoheptyl Phthalate and Diisononyl
Phthalate to Oryzias latipes
Honours Thesis
By: Kira N. MacDougall
Supervisor: V. S. Langlois
April 2014
An undergraduate thesis submitted to the School of Environmental Studies in partial
fulfillment of the requirements for ENSC 502
School of Environmental Studies
Queen’s University, Kingston, ON, Canada
2
ABSTRACT
Phthalates are industrial chemicals used primarily as plasticizers to increase the
flexibility of high molecular-weight polymers. These compounds are not covalently
bound to the polymeric matrix but are instead distributed between the macromolecules of
the polymer. Therefore, it is highly possible for them to leach out, inducing subsequent
environmental contamination. It is important to learn more about how these toxicants are
affecting the environment and vertebrate health. Diisoheptyl phthalate (DIHepP) and
diisononyl phthalate (DINP) are among the list of substances that are being assessed by
Environment Canada’s Chemical Management Plan. This study examines the effect of
DIHepP and DINP on the Asian fish Oryzias latipes because it is subject to phthalate
contamination in its natural habitat when phthalates leach out of plastic substances into
the surrounding environment. At 1 day post hatching, fish were exposed to a range of
environmentally relevant, low concentrations of DIHepP (0.037-0.3 µL/L) and DINP
(0.012-0.1 µL/L) under semi-static conditions for 7 d. None of the treatments
significantly induced mortality or malformation. Gene expression analysis was performed
to assess possible endocrine disruption of these compounds. The transcripts of interest
include deiodinase iodothyronine type 1 (dio1), thyroid hormone receptor β (trβ),
steroidogenic acute regulatory protein (star), steroid 5-alpha reductase type 2 (srd5α2),
and cytochrome P450 aromatase (cyp19β). The exposure did not have a statistically
significant effect on gene expression. Taken together, these results indicate that exposure
to DIHepP and DINP do not significantly affect mortality, malformation, or endocrine
disruption in O. latipes at the tested concentrations.
3
ACKNOWLEDGEMENTS
I would like to thank my supervisor Professor Langlois, for all I have learned
from her and her continual support and guidance during all stages of this thesis. I am
grateful to Professor Hodson for his valuable insight of Medaka health and the blue sac
disease condition. I would like to acknowledge Connor Edington for his help throughout
the duration of the exposure with animal care and sampling, Dr. Jing Zhang for his
guidance during RNA extraction and his assistance with statistical analysis, and Sarah
Wallace for preparing the primers used for this project and for her assistance with cDNA
synthesis. I would like to thank Tash-Lynn Colson, Laura Gibson, Justine MathieuDenoncourt, and the rest of the Langlois lab team for their help during qPCR and
statistically analysis. I would also like to acknowledge the animal care volunteers for
their assistance in caring for the Medaka. In addition, I would like to thank Shane R. de
Solla and Environment Canada for providing funding for this project.
4
TABLE OF CONTENTS
ABSTRACT ................................................................................................................................... 2
LIST OF TABLE ........................................................................................................................... 6
LIST OF COMMON ABBREVIATIONS ................................................................................... 7
1. INTRODUCTION AND LITERATURE REVIEW .............................................................. 8
1.1 Phthalates as Aquatic Contamination ......................................................................................................8
1.2 Effects of Phthalates on the Endocrine System ................................................................................. 10
1.3 Oryzias latipes ............................................................................................................................................. 14
1.4 Hypotheses & Research Objectives ...................................................................................................... 14
2. MATERIALS AND METHODS........................................................................................... 17
2.1 Phthalate Exposure ..................................................................................................................................... 17
2.1.1 Chemicals and reagents .................................................................................................................... 17
2.1.2 Animal Husbandry and Breeding .................................................................................................. 17
2.1.3 Experimental Design of Phthalate Exposures ........................................................................... 18
2.1.4 Sample Collection and Malformation Analysis........................................................................ 18
2.2 Gene Expression .......................................................................................................................................... 19
2.2.1 mRNA Isolation.................................................................................................................................. 19
2.2.2 cDNA Synthesis and Quantitative Real-Time Reverse Transcriptase Polymerase Chain
Reaction (qPCR) ........................................................................................................................................... 20
2.3 Data Analysis ............................................................................................................................................... 21
3. RESULTS ............................................................................................................................... 23
4. DISCUSSION ......................................................................................................................... 28
4.1 Mortality & Malformation ....................................................................................................................... 28
4.2 Gene Expression .......................................................................................................................................... 30
5. SUMMARY............................................................................................................................. 35
REFERENCES ............................................................................................................................ 36
5
LIST OF FIGURES
Fig. 1. Chemical structure of diisoheptyl phthalate (DIHepP) and diisononyl phthalate
(DINP)....................................................................................................................... 16
Fig. 2. Effect of DIHepP and DINP on mortality and malformation in O. latipes. .......... 24
Fig. 3. Picture of a healthy O. latipes and an O. latipes with BSD taken after the 7 d
exposure .................................................................................................................... 25
Fig. 4. Expression of ef1α, dio1, trβ, star, srd5α2, cyp19β in O. latipes exposed to
DIHepP and DINP .................................................................................................... 27
6
LIST OF TABLE
Table 1. qPCR primers and assay conditions for genes of interest in O. latipes .............. 22
7
LIST OF COMMON ABBREVIATIONS
BBP – Butyl benzyl phthalate
BSD – Blue Sac Disease
cyp19β - Cytochrome P450 aromatase
DBP – Di-n-butyl phthalate
DCHP – Dicyclohexyl phthalate
DEHP – Di-(2-ethylhexyl)-phthalate
DEP – Diethyl phthalate
DIDP – Diisodecyl phthalate
DIHepP – Diisoheptyl phthalate
DINP – Diisononyl phthalate
DMP – Dimethyl phthalate
dio1 - Deiodinase iodothyronine type 1
EDC – Endocrine disrupting compound
MMP – Mono-ethyl phthalate
srd5α2 - Steroid 5-alpha reductase type 2
star - Steroidogenic acute regulatory protein
TH – Thyroid hormones
trβ – Thyroid hormone receptor β
T3 – Triiodothyronine
T4 – Thyroxine
qPCR – quantitative real-time reverse-transcriptase Polymerase Chain Reaction
8
1. INTRODUCTION AND LITERATURE REVIEW
1.1 Phthalates as Aquatic Contamination
Phthalates are industrial chemicals used primarily as plasticizers to increase the
flexibility of high molecular-weight polymers. These compounds are not covalently
bound to the polymeric matrix, but are instead distributed between the macromolecules of
the polymer (Ganning et al., 1984). Therefore, it is highly possible for them to leach out,
inducing subsequent environmental contamination. Phthalate molecules are present in a
variety of products, including medical devices, children’s toys, building materials, food
packaging, and clothing (Martino-Anderson & Chahoud, 2009). Most products contain
approximately 30% di-(2-ethylhexyl) phthalate (DEHP) but some products contain up to
50% (Rank, 2005). Phthalates represent 69% of plasticizer use in the USA, 92% in
Western Europe and 81% in Japan (Stringer et al., 2000). In 2008, global demand for
phthalates exceeded 5 million metric tons (SRI Consulting, Inc. 2009). Given the largescale and widespread use of phthalates and that phthalates are not chemically bound to
the polymer, it is not surprising that many phthalates, such as di-n-butyl phthalate (DBP)
can be detected in air, soil and aquatic ecosystems (Shen et al., 2011). The median DEHP
concentration in Europe, China, and North America are 0.00105, 0.00111, and 0.00027
µL/L, respectively (Bergé et al., 2013). Globally, phthalate contamination in surface
water ranges from several microliters per liter, to several tens of microliters per liter
(Bergé et al., 2013).
The two phthalates used in this study are diisoheptyl phthalate (DIHepP) and
diisononyl phthalate (DINP) (Fig. 1). As of 2013, these two chemical substances were
placed among the list of substances being assessed by Environment Canada’s Chemical
9
Management Plan (Environment Canada, 2013). The goal of this plan is to obtain
information on the effect of chemicals that were used prior to the Canadian
Environmental Protection Act being enacted, and whose effects are unknown. Phthalates
are esters of phthalic acid (an aromatic dicarboxylic acid) that differ from each other by
the length of the carbon side chains. DIHepP has seven carbons, while DINP has nine.
High molecular weight phthalates (six carbon alkyl chain lengths or greater) are ideal
plasticizers due to their fluidity and low volatility (Patyna et al., 2005). Phthalates are
lipophilic compounds with solubilities that range from greater than 5 X 106 mg/L for dimethyl phthalate (DMP) (C1 phthalate) to 0.038 mg/L for diisodecyl phthalate (DIDP)
(C10 phthalate) (Staples et al., 2011). Previous studies have reported increased rates of
malformation and mortality from phthalates. A 96 h acute exposure of X. laevis embryos
investigating the effects of DBP, reported a statistically significant effect on
malformation (Lee et al., 2005). In their experiment, concentrations of 0.1, 0.5, 1, 5, 10,
and 15 µL/L DBP resulted in malformation rates of 7, 9, 15, 37, 51, 53, 90, and 100%
respectively. In the following study conducted by Ghorpade et al. (2002), toxic effects of
diethyl phthalate (DEP) were investigated in the freshwater fish, Cirrhina mrigala. The
fish were treated with 25, 50, 75, and 100 µL/L (w/v) DEP. 100% mortality occurred in
the 100 and 75 µL/L groups within 24 h. In the 25 and 50 µL/L DEP-treated group, there
was 10% and 50% mortality respectively in 72 h. Environment Canada is requesting
more research be performed to learn how DIHepP and DINP are interacting with the
environment and whether they are affecting mortality, malformation, and endocrine
disruption in target species.
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1.2 Effects of Phthalates on the Endocrine System
The endocrine system is a collection of glands that play a vital role in growth,
development, and reproduction. An endocrine-disrupting chemical (EDC) is a compound
that alters the homeostasis of hormonal control in an organism. When hormonal control is
disrupted, signal transduction between the organism and its external environment may be
compromised (Diamanti-Kandarakis et al., 2009). An EDC can be either natural or
synthetic in nature. When the endocrine system is disrupted, growth, developmental,
reproductive, and immune functions of an organism may be affected. In this study, the
effects of DIHepP and DINP will be studied on transcripts of interest including
deiodinase iodothyronine type 1 (dio1), thyroid hormone receptor β (trβ), steroidogenic
acute regulatory protein (star), steroid 5-alpha reductase type 2 (srd5α2), and cytochrome
P450 aromatase (cyp19β). These genes encode for proteins that affect both the thyroid
hormone and the sex steroid pathways.
Thyroxine (T4) and triiodothyronine (T3), collectively known as the thyroid
hormones (THs), are critical to growth, development, and metabolism of vertebrates
(McNabb & King, 1993). More specifically, these effects include regulating growth and
differentiation of tissues and organs; regulating energy homeostasis, body temperature,
and heart rate; and protein, fat, carbohydrate, and vitamin metabolism (Bowen, 2010).
Therefore, disruption of the TH axis can cause severe impairment to the affected
organism (Jugan et al, 2010). T4 and T3 are known to influence gene expression in
virtually every vertebrate tissue (Bianco & Kim, 2006). T4 is the main secretory product
of the thyroid gland and is relatively inactive when it first enters the bloodstream. One
role of the hormone thyroxine 5’-deiodinase, is to convert T4 to T3 (Epstein & Brent,
11
1994). The interaction between T3 and the specific thyroid hormone receptors, TRα and
TRβ, are primarily responsible for action of the THs (Sakurai et al., 1990). The genes
dio1 and trβ were chosen for analysis to determine if the TH pathway was being affected
by the presence of DIHepP and DINP in the water. In the absence and presence of TH,
dio1 acts as a transcription repressor and activator, respectively (Heimeier & Shi, 2010).
Thyroid hormone synthesis and signaling are important targets of endocrine disruptors
such as polychlorinated biphenyls, perchlorates, and brominated flame-retardants, and
phthalates (Jugan et al., 2010). For example, a study performed by Shen and colleagues
(2011) measured the effects of mono-n-butyl phthalate (MBP) and DBP in Xenopus
laevis (African Clawed Frog). X. laevis were exposed to 2, 10 and 15 µL/L for 21 days. It
was concluded that the effects of DBP and MBP induced changes in the expression of
selected thyroid hormone response genes: thyroid hormone receptor-beta (trβ), retinoid X
receptor gamma (rxrγ), and alpha and beta subunits of thyroid-stimulating hormone (tshα
and tshβ). Expression of trβ and rxrγ were down-regulated significantly at all
concentrations of MBP and DBP compared with the control group. The mRNA level of
tshβ increased significantly only in 10 and 15 µL/L MBP treated groups compared with
the control group. These findings suggest the potential disruption of TH signaling by
phthalate contamination.
Kim et al. (2002) assessed the effects of endocrine disruption in Oryzias latipes
(Japanese Medaka) exposed to DEHP. Fish were exposed to 0.001, 0.01, and 0.05 µL/L
DEHP from hatching to 3 months of age. In this study, blood vitellogen was analyzed as
a biomarker of hormonal disruption. Vitellogen is an egg yolk precursor protein
expressed in the female livers of fish species (Robinson, 2008). Blood vitellogen is
12
synthesized in female livers and is stimulated by hormones such as 17-β-estradiol (Chen
et al., 1986). It was concluded that DEHP reduced vitellogen levels in the blood of female
fish: no effects were observed in males. Their research suggests that DEHP is an antiestrogenic compound in female O. latipes (Kim et al., 2002).
In addition to the TH related pathway, some of the genes of interest in this
experiment code for proteins that affect the sex steroid pathway. Sex steroids are steroid
hormones that interact with androgen or estrogen receptors in vertebrates (Guerriero,
2009). Androgens and estrogens are the two main classes of sex steroids and are derived
from dihydrotestosterone and estradiol, respectively. The precursor molecule cholesterol
is converted into testosterone via a series of enzymatic steps within tissues such as the
testis (Waterman & Keeney, 1992). The conversion of cholesterol into an intermediate
molecule, pregnenolone, is largely controlled by star. (Kallen et al., 1998). The
movement of cholesterol from the outer mitochondrial membrane to the inner
mitochondrial membrane is facilitated by the star enzyme, which increases the
availability of substrate for the synthesis of all steroid hormones (Miller, 1988). A study
conducted by Ye and colleagues (2014) showed the upregulation of several genes in
marine medaka (Oryzias melastigma), including star and cyp19α in a 6 month exposure
of DEHP (0.1 and 0.5 µL/L) and monoethylhexyl phthalate MEHP (0.1 and 0.5 µL/L). In
this study, star was chosen as a transcript of interest, as it is crucial in the conversion of
cholesterol to testosterone, of which all sex steroids are derived.
Testosterone acts through two main pathways: androgen (either directly or after
5α-reductase conversion to dihydrotestosterone) and estrogen (via aromatase conversion)
(Pike, 2001). Enzymes coded for by srd5α2 catalyze the synthesis of the potent hormone
13
dihydrotestosterone from testosterone (Thigpen et al., 1993). Dihydrotestosterone is
essential for androgen-mediated growth of tissues and for the formation of the male
phenotype (Andersson et al., 1991). Mutations within the coding region of srd5α2 may
block development of the male phenotype (Waterman and Keeney, 1992) and result in
male pseudohermaphroditism, a condition in which external genitalia are phenotypically
female at birth (Thigpen et al., 1993). The transcript of srd5α2 is expressed at high levels
in androgen-sensitive tissue and was chosen as a gene of interest to study the androgen
axis.
In females, cyp19 encodes an enzyme involved in catalyzing the synthesis of
estrogens and has been identified in many different animal phyla, including vertebrates
(Chiang et al., 2001). In most teleost fish, two cyp 19 genes are present, cyp19α,
expressed in the ovary, and cyp19β, expressed in the brain (Cheshenko et al., 2008). In
the brain, androgens are converted into estrogens via aromatization (Chiang et al., 2001).
It is believed that cyp19β is critical in the regulation of sexual differentiation and the
female reproductive cycle in teleost fish (Cheshenko et al., 2008) and is the reason the
cyp19β gene was chosen as a gene of interest. Finally, diihydrotestosterone and estradiol
diffuse across the cell membrane and bind to androgen and estrogen receptors
respectively, within the cell. Once the sex steroids bind to the receptor, they modulate the
expression of many related genes (Nussey & Whitehead, 2011).
Previous studies have investigated the effects of phthalates on genes involved in
the sex steroid pathways. In utero exposure to DBP leads to a reduction in testosterone
production by the fetal testis. Thompson et al. (2004) gavaged 12 pregnant rats daily with
500 mg/kg DBP. A significant decrease in testosterone production and mRNA expression
14
of scavenger receptor B1, side chain cleavage enzyme (P450scc), star, and cytochrome
p450c17 were observed. Each of the genes demonstrated a similar relative expression
level at both gestational day 17 and 18, but the mean percentage of expression of the four
genes relative to the control was 46.4% and 15.4% in the gestational day 17 and 18 DBPexposed fetuses respectively (Thompson et al., 2004). This study shows that DBP is
coincident with decreased transcription of several genes in the cholesterol transport and
steroidogenesis pathways.
1.3 Oryzias latipes
O. latipes are unique among common laboratory teleosts. The presence of highly
polymorphic inbred lines, high fecundity, small adult size, ease of husbandry, and a
relatively small sized (~800 Mb), completely sequenced genome, are reasons why O.
latipes are extensively used in biological research and ecotoxicology studies (Takeda &
Shimada, 2010). Its genome is approximately one-third the size of the human genome
and less than half the size of the Danio rerio (zebrafish) genome. O. latipes is readily
bred under laboratory conditions and reaches sexual maturity four to six weeks after
hatching. O. latipes was chosen as the species of interest because it is subject to phthalate
contamination in its natural habitat when phthalates leach out of plastic substances into
the surrounding environment.
1.4 Hypotheses & Research Objectives
O. latipes were exposed to a range of environmentally relevant, low
concentrations of DIHepP (0.037-0.3 µL/L) and DINP (0.012-0.1 µL/L). These
concentrations were chosen to reflect environmental conditions. It is hypothesized that
15
DIHepP and DINP will have a significant affect on mortality and endocrine disruption
due to evidence from past environmental studies. However, after reviewing previous
literature, it is hypothesized that there will not be a statistically significant effect on
malformation. In a 96 h acute exposure of X. laevis embryos, there was only a 7% and
9% malformation rate at concentrations of 100 and 500 DBP respectively. The
concentrations of phthalate in this study are higher than the concentrations being tested in
the present study, and therefore it is hypothesized there will not be a statistically
significant change between treatments. To test this hypothesis, gene expression analysis
was performed in O. latipes. The transcripts of interest include dio1, trβ, star, srd5α2,
and cyp19β.
16
A.
B.
Fig. 1. Chemical structure of A. diisoheptyl phthalate (DIHepP) and B. diisononyl
phthalate (DINP). DIHepP and DINP are aromatic dicarboxylic acids that differ from
each other by the length of their carbon side chains.
17
2. MATERIALS AND METHODS
2.1 Phthalate Exposure
2.1.1 Chemicals and reagents
DIHepP, DINP, formaldehyde and ethyl 3-aminobenzoate methanesulfonic acid
(MS-222) were obtained from Sigma-Aldrich, (Oakville, ON, CA).
2.1.2 Animal Husbandry and Breeding
The adult fish were fed three times at 09:00, 12:00, and 18:00, once with brine
shrimp (Pets and Ponds, Orillia, ON, CA) and twice with tropical fish flake (Pets and
Ponds, Orillia, ON, CA). Water temperature and pH were kept between 24.8-25.6 °C and
7.0-7.5, respectively. A 16-h light/8-h dark photoperiod was used to maintain health of
fish. Males and females were kept together in the same tanks to allow for daily egg
production. Eggs were collected twice per day directly from the females by capturing the
fish with a small mesh net and running the first finger and thumb gently along the length
of the fish. Eggs were collected over a three-day timespan and placed in an embryorearing medium containing 1000 mg/L NaCl, 40 mg/L CaCl2, 30 mg/L KCl, and 160
mg/L MgSO4 to prevent bacterial infection. The eggs were placed in 19 X 65 mm glass
vials (Thermofisher, Ottawa, ON, CA) and checked for hatchlings every 24 h. The water
was fully replaced (100% renewal) every 24 h. The hatchlings were moved to a holding
tank with dechlorinated water until 24 h post hatching, when the exposure began. O.
latipes were kept at the Queen’s University Animal Care Facility (Kingston, ON, CA).
Animal care and experimentation was performed according to the Canadian Council on
Animal Care Guide to the Care and Use of Experimental Animals (1993) as well as the
18
policies and procedures established by the Queen’s University Animal Care Committee
and the Canadian Council of Animal Care.
2.1.3 Experimental Design of Phthalate Exposures
The highest concentration of DIHepP and DINP tested in this experiment was
determined by performing a solubility test. Acetone (0.01%) is recommended by the
Standard Guide for Conducting Early Life-Stage Toxicity Tests with Fishes
(Environmental Protection Agency, 1996) and was chosen for this experiment. Larval
fish were exposed to a geometric series of concentrations of DIHepP (0.037, 0.075, 0.15,
and 0.3 µL/L) and DINP (0.012, 0.025, 0.05, and 0.1 µL/L). Stock solutions were
prepared for DINP and DIHepP by mixing the total daily amount of phthalate required
into 300 mL of distilled water. Both water and acetone were used as negative controls.
At 1 d post hatching, fish were exposed to the solutions under semi-static
conditions for 7 d. Each treatment was run in quadruplicate. Each jar contained 5 fish
held in 70 mL of solution. Water and solution renewal were completed every 24 h. Dead
fish were removed from each jar daily, counted, and placed in a 10% formalin solution
for further analysis of malformation.
2.1.4 Sample Collection and Malformation Analysis
After one week of exposure, the remaining fish were anaesthetized in a 100 mg/L
MS-222 solution and scored for malformations under a Stereomaster dissecting
microscope (Thermofisher, New Jersey, USA). Blue Sac Disease (BSD), was used for
analysis of malformation, as BSD can be result from exposure to environmental
contaminants, such as DIHepP and DINP. A modified BSD scoring method was used
19
(Lin, 2014) in this experiment. Blind analysis was conducted to remove observer bias.
The fish were removed from the MS-222 and immediately placed in 1.5 mL tubes
(Thermofisher, Ottawa, ON, CA) on dry ice, and stored at -80 °C until gene expression
analysis was conducted. For water analysis purposes, 200 mL water samples were
collected during the experiment to validate absolute and constant concentrations of the
phthalates. Water samples were collected prior to the start of the exposure (0 h), after 24
h, and stored at -20 °C until further analysis. The water samples were not the scope of
this honours thesis project, and thus data is not included.
2.2 Gene Expression
2.2.1 mRNA Isolation
Three fish were pooled together to make one sample weighing approximately
0.0034 g. There were 6 replicates for each phthalate treatment and 8 replicates for each of
the water and acetone controls. A commercially available kit, QIAGEN RNeasy Micro
kit (Ottawa, ON, CA), was used to isolate RNA. The procedure was performed quickly to
prevent RNA degradation. All steps were performed at room temperature. Samples were
homogenized by sonication (Homogenizer MDL 150T 1/8 MT, Thermofisher, Ottawa,
ON, CA) and centrifuged using the Sorvall Legend 21 Centrifuge (Thermofisher, Ottawa,
ON, CA). A series of buffers and wash solutions were subsequently passed through the
membrane of the column and discarded. RNA was quantified using the NanoDrop-2000
Spectrophotometer (Thermofisher, Ottawa, ON, CA) and stored at -80 °C in 1.5 mL
tubes.
20
2.2.2 cDNA Synthesis and Quantitative Real-Time Reverse Transcriptase
Polymerase Chain Reaction (qPCR)
Specific Primers (Table 1) used in this study were developed using the National
Center for Biotechnology Information (NCBI) Primer Blast. RNA samples were used as a
template to synthesize complementary DNA (cDNA) using a commercially available kit
(QuantiTect Reverse Transcripation Kit, QIAGEN, Ottawa, ON, CA). Primers and
nuclease-free water were added to 1 µg RNA according to the manufacturer’s protocol.
Samples were incubated for 2 min at 42 °C then placed on ice. Master Mix containing
RNase-free water, gDNA Wipeout Buffer (Quantiscript), Reverse Transcriptase,
Quantiscript RT Buffer, RT Primer Mix, and nuclease-free water was added to the
solution. The samples were then placed in the thermocycler to incubate for 15 min at 42
°C and inactivated for 3 min at 95 °C. For each sample, 0.5 µg of complimentary (cDNA)
was made. Finally, the samples were stored at -20 °C.
cDNA was amplified and quantified by quantitative real-time reverse transcriptase
polymerase chain reaction (qPCR). A BioRad CFX96 Real Time System qPCR
(Mississauga, ON, CA) was used to assess the levels of gene expression in the samples. A
serial dilution (1:4) was used to create the standard curves. Every plate contained a no
template control (NTC) and a negative reverse transcriptase control (NoRT) to ensure the
samples were not contaminated. A Master Mix was created with Promega GoTaq qPCR
MasterMix (Madison, Wisconsin, USA), forward primer, reverse primer and nuclease
free water and was added to each sample. Both the standard curves and the samples were
run in duplicate. The thermocycler program included heating the samples to 95 °C for 3
min, cycles at 95 °C for 15 s, and finally cycles at 58 °C or 60 °C, depending on the gene
21
being analyzed. The genes of interest included srd5α2, cyp19β, trβ, dio1, and star. A 1:40
dilution was performed for trβ, dio1, and star and a 1:80 dilution was performed for
srd5α2 and cyp19β. Gene expression was reported as fold changes relative to the water
control. The results were normalized with elongation factor 1 (ef1α). This gene was
chosen as the normalizing gene, as it did not change significantly between treatments.
2.3 Data Analysis
After one week, the percentage of mortality was calculated for each replicate and
the average calculated for each treatment. Each fish was examined for BSD at the end of
the exposure. The malformations were calculated for each replicate by taking the sum of
the malformations and dividing by the total number of individuals per treatment; the
average was calculated for each treatment. A Fisher exact test was used to determine if
there were statistical significance and differences among treatments in mortality and
malformations. For gene expression data analysis, normality of the results was examined
using a Goodness of Fit test. Homoscedasticity was verified using the Levene’s test. The
Kruskal-Wallis test was used for genes trβ, dio1, and star, because the data were
nonparametric. A one-way analysis of variance (ANOVA) was performed for each gene.
The level of statistical significance was set at p < 0.05.
22
Table 1. qPCR primers and assay conditions for genes of interest in O. latipes. Complete
list of target genes, primer sequences (5’-3’), annealing temperature (°C), amplicon size
(bp) and optimized primer concentration (µM). Primer sequences were custom developed
from NCBI Primer Blast. Legend: F, forward primer; R, reverse primer; srd5α2, steroid
5-alpha reductase type 2; cyp19β, cytochrome P450 aromatase; trβ, thyroid hormone
receptor β; dio1, deiodinase, iodothyronine, type 1; star steroidogenic acute regulatory
protein; ef1α, elongation factor 1. The primers were custom developed.
Target
gene
Primer
Sequence (5’ – 3’)
Annealing
temperature (°C)
Amplicon size (bp)
Primer Concentration
ef1α
F
R
F
R
F
R
F
R
F
R
F
R
AGTGGCTTTTGTCCCCATCT
CACTGGCATTTCCGTCCTTG
GACTTCGTGAATGTGCGTTGT
GCTCTAATGATGCCCTTCTGCT
ATCGTTCCTGTGCTCGTGG
TGAAGAGGTTGGTGGGGTCT
TGACCTTAGACTGCCTCCCA
AGATCCCCCTCTGTCTGGGT
AGTCTGGGTCTCGTTCTGTT
CCCCAAGAGAAACACCGCTC
CAAAGAAGCCGTCACCAACA
ATGCTGGTCCTCTCCGTCTC
58
113
58
107
60
97
58
83
60
118
60
108
0.35
0.35
0.35
0.35
0.35
0.35
0.30
0.30
0.35
0.35
0.30
0.30
star
srd5α2
cyp19
trβ
dio1
23
3. RESULTS
There was no statistically significant effect of DIHepP or DINP on malformation
or mortality (Fig. 2). Malformation and morality were equal to or below 12.5% for each
treatment. Results were not statistically significant for either malformation or mortality (p
> 0.05). The malformation used in this experiment was BSD (Fig. 3).
qPCR was used to determine the expression level of six genes, ef1α, dio1, trβ,
star, srd5α2, and cyp19β (Fig. 4). Dissociation curves displayed a single peak for each
gene, indicating there was no DNA contamination in the wells. All the NTC and NoRT
wells showed no amplification. The calibration curve efficiency ranged from 90.4 to
113.8% and R2 was between 0.991 and 0.997. The control gene, ef1α, did not vary
significantly among treatments (p > 0.05) and was therefore used as the normalizing
gene. For the srd5α2 gene, mRNA expression levels were transformed by log10. Once
normalized, the mean relative gene expression did not vary significantly between
treatments for any gene (p > 0.05). There were no trends in the data. The water analyses
were not a major purpose of this project and were not included.
24
Fig. 2. Effect of DIHepP and DINP on A. malformation and B. mortality in O. latipes.
Data are expressed as mean + SD. Data were analyzed using a one-way ANOVA (n = 48); p > 0.05).
25
A
B
1.
2.
Fig. 3. Pictures of A. normal healthy O latipes from the water treatment and B. O. latipes
with BSD from the 0.3 µL/L treatment, taken after the 7 d exposure. Legend: 1. A
cranial-facial deformity; 2. Yolk sac edema.
26
A ef1α
A.
6
Relative mRNA Levels
1
4
2
0
CC. trβ
0
0
0
0
0
0
0.
D star
D.
3
Relative mRNA Levels
6
4
2
0
2
1
0
e 7 5 0 0 2 5 0 0
er
at eton .03 .07 .15 .30 .01 .02 .05 .10
0 0 0 0 0 0 0 0
W c
A
DIHepP
DINP
e 7 5 0 0 2 5 0 0
er
at eton .03 .07 .15 .30 .01 .02 .05 .10
0 0 0 0 0 0 0 0
W c
A
DIHepP
DINP
Treatments (ppb)
Treatments (ppb)
F.
F cyp19β
E.
E srd5α2
4
Relative mRNA Levels
2.50
Relative mRNA Levels
0
0
e 7 7 5 5 0 12 2 5 5 0 0
er
1 3
at to .
. 0. 0.1
0 0. 0. 0.
A
DIHepP
DINP
Treatments (ppb)
n
W
.
.
0
.
0
.
75
.
.
.
A
c
t
t
37
2 5 0 0
er one
a
0 0 15 30 01 02 05 10
W e 0 0 0 0 0 0 0 0.
DIHepP
DINP
Treatments (ppb)
03
0
ce
Relative mRNA Levels
2
Relative mRNA Levels
27
B.B dio1
1.25
0.00
2
0
e 7 5 0 0 2 5 0 0
er
at ton 03 07 15 30 01 02 05 10
W ce 0. 0. 0. 0. 0. 0. 0. 0.
A
DIHepP
DINP
Treatments (ppb)
W
e 7 5 0 0 2 5 0 0
er
at eton .03 .07 .15 .30 .01 .02 .05 .10
0 0 0 0 0 0 0 0
c
A
DIHepP
DINP
Treatments (ppb)
Fig. 4. Expression of A. ef1α, B. dio1, C. trβ, D. star, E. srd5α2, F. cyp19β in O. latipes
exposed to DIHepP and DINP. Expression levels were measured after the 7 d exposure.
Relative mRNA levels for each sample was analyzed using qPCR. Bars represent the
mean fold change + SE. The scales on the y-axis vary between genes. Data were analyzed
using a one-way ANOVA (n = 5-6); p > 0.05).
28
4. DISCUSSION
4.1 Mortality & Malformation
According to the ASTM (1998), in order for this experiment to be valid, mortality
in the controls must be below 10%. A mortality rate of 5% and 0% for the water control
and acetone control, respectively, was found. In our study, DIHepP and DINP did not
have a statistically significant effect on mortality at the tested concentrations. A
comparable aquatic toxicology study performed by Adams et al. (1995) on freshwater
and marine species, reported similar findings. In their study, 14 different phthalates,
including DINP, were tested at concentrations ranging from 0.21 to 377 µL/L. It was
concluded that phthalates with alkyl chain lengths of six carbon atoms or more, such is
the case with DIHepP and DINP, were not acutely toxic at concentrations approaching
their respective aqueous solubility. Furthermore, Chikae et al. (2004) investigated the
effects of DEHP (0.00001, 0.0001, 0.001, and 0.01 µL/L) in O. latipes after a 3 wk
period. These concentrations are comparable to the concentrations used in our experiment
and despite a longer exposure period, there was still no effect on mortality. In contrast,
Mankidy et al. (2013), studied the biological impacts of DEHP, DEP, di-n-butyl phthalate
(DBP), and butyl benzyl phthalate (BBP) in developing Pimephales promelas (Fathead
Minnow) embryos. Exposure to 1 µL/L DEHP resulted in 30% mortality, while exposure
to 10 µL/L DEP caused 52% mortality. Exposure to DBP or BBP at 1 µL/L did not cause
significantly greater mortality compared to the controls. The embryos were exposed to
the phthalate until 96 h post hatching. These concentrations were 10 and 100 times
greater than the concentrations tested in our experiment and may explain why mortality
was observed in this experiment compared to the present study. In the following study
29
conducted by Ghorpade et al. (2002), toxic effects of DEP were investigated in the
freshwater fish, Cirrhina mrigala. The fish were treated with 25, 50, 75, and 100 µL/L
(w/v) DEP. 100% mortality occurred in the 100 and 75 µL/L groups within 24 h. In the
25 and 50 µL/L DEP-treated group, there was 10% and 50% mortality respectively in 72
h. These concentrations are 250 to 1000 times higher than the concentrations used in our
experiment and may account for the difference in toxicity. In conclusion, studies that
tested phthalates at concentrations similar to our tested concentrations, no effect on
mortality was observed. However, in studies that tested phthalates at concentrations
greater than 0.1 µL/L, mortality was significantly elevated. Therefore, if our study
investigated the effect of DIHepP and DINP at concentrations higher than 0.1 µL/L,
mortality might have occurred.
In addition to mortality, malformation was used as an endpoint in this study. The
maximum acceptable rate of malformation for the controls was set at 10% according to
the American Society for Testing and Materials (ASTM). In our experiment, the water
control induced 2.5% malformation, which is acceptable according to the ASTM
guidelines. The rate of malformation in the acetone control (0.01%) was 12.5%, above
the acceptable range. However, 0.01% acetone was used in each of the treatments, and
this was the only treatment that exceeded the acceptable limit. The purpose of the acetone
control was to ensure that malformations seen in our experiment were induced by the
presence of DIHepP and DINP in the water, and not the acetone itself. Our results
indicate there was no statistically significant effect on malformation between treatments.
In contrast, a 96 h acute exposure of X. laevis embryos investigating the effects of DBP,
reported a statistically significant effect on malformation (Lee et al., 2005). In their
30
experiment, concentrations of 0.1, 0.5, 1, 5, 10, and 15 µL/L DBP resulted in
malformation rates of 7, 9, 15, 37, 51, 53, 90, and 100% respectively. The concentrations
used in this study, with the exception of 0.1 µL/L, are much higher than the
concentrations studied in our experiment. Therefore, if the concentrations of phthalates
used in this study had been 0.5 µL/L or higher, a statistically significant change in rate or
malformation may have been observed.
4.2 Gene Expression
Phthalates are known endocrine disruptors and have been shown to have an effect
on thyroid hormone-related genes in aquatic species (Shen et al. 2011). The gene dio1
was chosen to determine if the conversion of T4 to T3, which is critical to growth,
development, and metabolism (McNabb & King, 1993), is being affected by the presence
of DIHepP and DINP. In this experiment, the transcript level of dio1 did not vary
significantly among treatments. Another study investigating the effects of mono-methyl
phthalate (MMP), DMP, and dicyclohexyl phthalate (DCHP) in Silurana tropicalis
(Western Clawed Frog) embryos exposed from Nieuwkoop and Faber (NF) 12 to NF 46,
suggested that the chemical structure of the phthalate is important in whether or not the
transcript level of dio1 is significantly affected. In this acute exposure conducted by
Mathieu-Denoncourt et al. (in preparation), MMP and DMP did not have a statistically
significant effect on dio1 transcript levels at concentrations up to 4.1 µL/L. However,
dio1 was upregulated significantly at concentrations of 1.5 and 4.1 µL/L DCHP. DCHP
has a cyclic ring at the end of each alkyl side chain, while MMP and DMP do not, which
may account for the endocrine disrupting effect of this compound. Like MMP and DMP,
DIHepP and DINP lack this cyclic functional group. Furthermore, this study tested
31
concentrations of phthalate that were more than 15 times higher than the concentrations
tested in the present study and was conducted in frogs. A second study of DIHepP and
DINP in S. tropicalis embryos showed that dio1 transcript levels did not vary
significantly. In an acute exposure of S. tropicalis embryos, concentrations of 0.03, 0.3,
3, 30 µL/L DIHepP and 0.01, 0.1, 1, 10 µL/L DINP were tested from NF 12 to NF46. As
in our experiment, the transcript level of dio1 was not statistically different among
treatment groups (Lam et al., in preparation). These concentrations are comparable to the
ones used in our experiment and show that in two different aquatic species, S. tropicalis
and O. latipes, DIHepP and DINP do not affect transcript levels of dio1. As in dio1, no
significant trend or change was observed in trβ transcript levels. This gene is a thyroid
hormone receptor that acts as a transcription repressor and activator in the absence and
presence of TH, respectively (Heimeier & Shi, 2010). In contrast to our results, a 2011
study by Shen et al., assessed the effects of exposing X. laevis to concentrations of 2, 10
and 15 µL/L DBP separately for 21 days. The trβ mRNA levels were downregulated in
all treatments relative to the control. Shen and colleagues used concentrations that were
more than 20 times higher than the concentrations used in our experiment, and may
explain why the transcript level of trβ changed significantly in their experiment.
Furthermore, the duration of this experiment was three times longer than ours, which may
also account for this difference. However, two other studies of the effect of phthalates on
the transcript level of trβ did not report statistically significant changes. In two 72 h study
of S. tropicalis embryos (NF 12 to NF 46), the transcript level of trβ did not change
significantly at concentrations of up to 4.1 µL/L MMP, 4.1 µL/L DMP, 4.1 µL/L DCHP,
10 µL/L DINP, and 30 µL/L DIHepP (Mathieu-Denoncourt et al. (in preparation)); Lam
32
et al. (in preparation)). Concentrations must often be greater than 2 µL/L for the TH
pathway to be significantly impacted by most phthalates. This threshold may vary due to
factors such as species, duration of exposure, and chemical structure of the phthalate, and
therefore more research is needed in this area.
In addition to the TH pathway, genes involved in the sex steroid pathway were
also investigated due to previous studies showing the effects of phthalates on endocrine
disruption in this pathway (Ye et al. 2014; Lovekamp & Davis, 2003). In this experiment,
there was no significant change observed in star, a gene responsible for increasing the
movement of cholesterol from the outer mitochondrial membrane to the inner
mitochondrial membrane (Miller, 1988). This movement increases the availability of
substrate for the synthesis of all steroid hormones and is therefore an important gene in
the sex steroid pathway. In contrast to our results, Ye et al. (2014) observed an
upregulation of star in male fish (O. melastigma) exposed to DEHP (0.1 and 0.5 µL/L)
and MEHP (0.1 and 0.5 µL/L) in for 6 months. In females, the transcript level of star did
not change significantly among treatments for either phthalate. These concentrations are
comparable to those used in our study, suggesting that if the duration of the exposure was
increased, star transcript levels may have changed significantly between treatments.
Another study by Aoki et al. (2011), studied the effects of DBP (0, 0.015, and 0.035
µL/L) in three-spined sticklebacks (Gasterosteus aculetaus) for 22 d. Their results
indicate the expression of star did not appear to be affected by DBP exposure. Despite
the longer exposure period, these results are consistent with the results from our study
and tested similar concentrations. After cholesterol is converted to testosterone,
testosterone can be metabolically activated into the more potent androgen,
33
dihydrotestosterone, by enzymes coded for by srd5α2 (Ye et al., 2011). Therefore, in this
study, the expression of the transcript for the rate-limiting enzyme involved in androgen
biosynthesis was investigated. Transcripts coded for by srd5α2, did not significantly
change among treatments. While there was no significant change in the mRNA level of
srd5α2 at concentrations as high as 30 µL/L DIHepP and 10 µL/L DINP (Lam et al., in
preparation), there was after exposure to 0.3 µL/L DCHP (Mathieu-Denoncourt et al., in
preparation). As mentioned previously, the structural differences of DCHP compared to
DIHepP and DINP may account for this difference in gene expression.
Along with dihydrotestosterone, estradiol can be synthesized from testosterone.
During the final step of estrogen biosynthesis, cytochrome P450 aromatase, encoded by
cyp19β, converts androgens into estrogens (Cheshenko et al., 2008). As a previous study
has shown, phthalates can cause endocrine disruption in the female axis in O. melastigma
(Ye et al., 2014). Therefore, cyp19β was chosen as a transcript of interest. In this
experiment, the transcript level of cyp19β did not vary significantly between treatments.
Two other studies, investigated the effects of DIHepP and DINP in S. tropicalis from NF
12 to NF 46. Concentrations as high as 30 µL/L DIHepP and 10 µL/L DINP did not
result in a significant change in the transcript level of cyp19β, exposure to 0.3 and 1.5
µL/L DCHP did. As mentioned previously, this may be due to the cyclic functional
groups on the DCMP alkyl side chains. Finally, a study conducted by Ye et al. (2014),
investigated the effects of DEHP (0.1 and 0.5 µL/L) and MEHP (0.1 and 0.5 µL/L) in O.
melastigma for 6 months. They observed an upregulation of cyp19β in females exposed
to 0.1 µL/L DEHP, whereas 0.5 µL/L DEHP and 0.1 µL/L MEHP significantly
downregulated cyp19β. In males, cyp19β was significantly downregulated after 0.5 µL/L
34
DEHP and 0.1 µL/L MEHP. In our study, sex was not investigated. However, this would
provide an interesting avenue for future research.
According to the Health Canada guideline, a concentration of up to 1,000 µL/L
DINP is allowed in the soft vinyl of toys and childcare articles (Health Canada, 2012).
Given this standard, future research should focus on studying DIHepP and DINP at
higher concentrations. Furthermore, DIHepP and DINP could be studied in more fish
species to gain a more holistic view of how these two phthalates are affecting aquatic
ecosystems. Experimental variables such as duration of exposure could also be altered to
gain more insight into the effects of these phthalates. However, in our study, DIHepP and
DINP did not have a significant effect on mortality, malformation or gene expression of
ef1α, dio1, trβ, star, srd5α2, and cyp19β in O. latipes at the environmentally relevant
concentrations tested over a 7 d exposure.
35
5. SUMMARY
Phthalates are industrial chemicals used primarily to increase the flexibility of
high molecular-weight polymers. The phthalates tested in this study, DIHepP and DINP,
are among the list of substances that are being assessed by Environment Canada’s
Chemical Management Plan. The effects of DIHepP and DINP on mortality,
malformation, and endocrine disruption were assessed in an acute exposure of O. latipes.
At one day post hatching, fish were exposed to concentrations of DIHepP (0.037-0.3
µL/L) and DINP (0.012-0.1 µL/L) under semi-static conditions for 7 d. The transcripts of
interest included dio1, trβ, star, srd5α2, and cyp19β. These genes encode for proteins that
affect both the thyroid hormone and the sex steroid pathways. Contradictory to the initial
hypothesis, it was concluded that there was no statistically significant effect on mortality,
malformation, or gene expression for this experiment despite the environmentally
relevant concentrations tested. This brings considerable optimism with respect to the
future of aquatic ecosystem health and phthalate contamination. The data from this
experiment may prove useful to Environment Canada in making future policy decisions
regarding phthalate contamination in O. latipes.
36
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