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Jan L. Lyche. 2011. Phthalates. In Reproductive and Developmental Toxicology Editor. Ramesh C. Gupta. 48: 637-655

  • Norwegian University of Life Sciences (NMBU), Oslo
Reproductive and Developmental Toxicology, Edited by Ramesh C. Gupta
ISBN: 978-0-12-382032-7
Copyright © 2011, Elsevier Inc.
Advances in materials sciences and engineering during the
last decades have led to a widespread use of phthalates
(phthalic acid esters) in a wide range of industrial products.
Phthalates are used as plasticizers that impart flexibility and
durability to polyvinylchloride (PVC) products. They are
also used in solvents, lubricating oils, fixatives and as deter-
gents in personal care products. When incorporated into
PVC, phthalates are not covalently bound and are therefore
easily released to the surroundings, leading to contamination
of the external environment. Phthalates are detected in sev-
eral media including food, water, house dust and air, thereby
exposing animals and humans.
Phthalates were first introduced as a plasticizer in the 1920s
and quickly replaced the volatile and odorous camphor. In 1931
polyvinyl chloride became commercially available and the
development of di-2-ethylhexyl phthalate led the boom of the
plasticizer PVC industry starting from the 1950s. Between 1970
and 2006 worldwide production grew from 1.8 to 4.3 million
tons (Habert et al., 2009). In 2006 di-isononyl phthalate (DINP)
had the highest production volume, followed by di-isodecyl
phthalate (DIDP) and di(2-ethylhexyl) phthalate (DEHP), then
butyl benzyl phthalate (BBP), dibutyl phthalate (DBP), di-n-
octyl phthalate (DnOP) and di-isobutyl phthalate (DIBP).
Because of widespread use, ubiquitous and constant envi-
ronmental presence, exposure to humans, domestic animals
and wildlife is virtually unavoidable. In the general popula-
tion the major source of human exposure is through ingestion
of contaminated food and water. Other significant sources
are inhalation of indoor air and dermal uptake via cosmetics
(Kavlock et al., 2006; Koo and Lee, 2007). Humans are also
exposed to high doses of phthalates from medical devices
during medical procedures such as blood transfusions and
hemodialysis (Calafat et al., 2004). Furthermore, increased
human exposure is documented in patients treated with
pharmaceuticals where phthalates are incorporated in the
coatings. The widespread exposure of the general population
is documented in several recent monitoring studies in the
USA and Europe (Koch and Calafat, 2009). Phthalate esters
and their metabolites are detected in human urine (Koch
et al., 2006), breast milk (Lottrup et al., 2006; Main et al., 2006;
Zhu et al., 2006), and amniotic fluid (Silva et al., 2004a,b).
Furthermore, phthalates are also able to cross the placenta,
and fetal exposure is closely correlated with maternal expo-
sure (Latini et al., 2003a).
Phthalate esters are suggested to have endocrine dis-
rupting properties (Latini, 2005) and exposures to high
concentrations were shown to induce fetal death, cancer,
malformations, liver and kidney injury and reproductive tox-
icity in animals (Lovekamp-Swan and Davis, 2003; Hauser
and Calafat, 2005; Latini et al., 2006). Furthermore, phthal-
ates are well-known antiandrogens in experimental animals,
with perinatal exposure as the apparently most sensitive life
stage. Reduced testosterone and adverse male reproductive
system development are extensively documented in rodents
and the adverse effects observed in animals raise concerns
as to whether phthalates represent a potential health risk to
humans (Kavlock et al., 2006). The observed high sensitivity
of the prenatal developmental stage for endocrine disruption
has led to the postulation that increased incidence of human
reproductive deficits may be produced by exposure to envi-
ronmental chemicals during fetal and/or pre-pubertal life
(Sharp and Skakkebaek, 2008). Reports have been published
which also associate exposure to phthalates with increasing
incidence of other human diseases such as obesity, insulin
resistance and type 2 diabetes, asthma and allergies and neu-
rological disorders such as ADHD and autism.
Even though association between adverse human health
effects and exposure to phthalates has been reported, no clear
cause–effect relationships are presently documented. However,
based on the animal data, the Environment Directorate-General
of the European Commission categorized DEHP, DBP and BBP
as “reproductive-toxic”. DEHP and DBP are also anticipated
to be a human carcinogen. Therefore the use of the phthalates
DEHP, DBP and BBP is forbidden in toys, cosmetics and food
contact materials. The use of DINP, DIDP and DNOP (di-n-oc-
tylphthalate) is also forbidden for toys and baby clothes.
Phthalates are diesters of 1,2-benzenedicarboxylic acid
(phthalic acid) containing a benzene ring with two ester func-
tional groups (Figure 48.1). The water solubility is low and
Jan L. Lyche
decreases with increasing length of the side chain (the alco-
hol moiety) or with higher molecular weight (MW). As a con-
sequence, low molecular weight phthalates with short alkyl
groups such as DMP and DBP are more water soluble whereas
the long-chained phthalates are poorly soluble. Volatility at
standard temperature is low, particularly for long-chain
phthalates such as DEHP and BBP (Rusyn et al., 2006). Uses
of the various phthalates mainly depend on their molecular
weight (MW). Higher MW DEHP, DiNP, and DiDP are used in
construction materials and numerous PVC products includ-
ing clothing (footwear, raincoats), flooring and wall coverings,
food packaging, children’s products (toys, grip bumpers), and
medical devices. Manufacturers use low MW phthalates such
as di-methyl phthalate (DMP), di-ethyl phthalate (DEP) and
DBP as solvents in personal care products (perfumes, lotions,
cosmetics), insecticides, lacquers and in coatings including
those used to provide timed releases in some pharmaceuti-
cals. They are also used in PVC (Heudorf et al., 2007).
Humans are exposed to phthalates through ingestion,
inhalation and dermal contact. For the general population the
major route of exposure for most phthalates is ingestion of
food and water. Infants and young children (0.5–4 years of
age) consume more calories per kg body weight and consume
relatively more fatty foods compared to adults. The estimated
total dietary DEHP intake is highest in children followed
by adolescents younger than 19 years of age. In addition to
dietary exposure, oral intake of phthalates also occurs when
children mouth, suck or chew on toys containing phthalates
(Bouma and Schakel, 2002). As a consequence, the European
Union (EU) banned the use of DEHP, DBP, BBP and DiDP in
children’s toys and child-care items for children below 3 years
of age in 1999 (EU Decision, 1999), and in 2005 the EU prohib-
ited DEHP from all toys and child-care products (EU Direc-
tive, 2005). In the USA it is still permitted to use DiNP in toys
and it was shown that children receive considerable levels of
DiNP as a result of mouthing activities (Kavlock et al., 2002).
Humans may also be exposed orally to phthalates incor-
porated in the coatings of commonly used pharmaceuticals.
Herbal preparations and nutritional supplements, including
some intended for use during pregnancy, may also contain
phthalates in the formulation (Schettler, 2006).
Phthalates may permeate into humans via inhalation of
indoor air in rooms with large surfaces of PVC-containing
products (Table 48.1). The most common phthalate found
indoors is DEHP. Other phthalates detected in indoor air
include butyl benzyl phthalate (BBzP) di-n-butyl phthalate
(DnBP), di-isobutyl phthalate (DiBP) and DEP (Bornehag
and Nanberg, 2010). Concentrations of DEHP in indoor air
in Japan were reported to be up to 1,000-fold higher than
outdoor air concentrations (Otake et al., 2004). In Germany
higher levels of DEHP were detected in kindergartens
compared to apartments (Fromme et al., 2004). In Sweden
significantly higher DEHP levels in dust were found in
homes of children with doctor-diagnosed asthma (Bornehag
et al., 2004). This exposure route is suggested to be of limited
importance for adults.
Dermal exposure may also be important routes of expo-
sures to phthalates such as DBP, which are used in many
cosmetics including perfume, hair gels, hair sprays, body
lotion, deodorant and nail polish (Koch et al., 2003a; Koo
and Lee, 2007). In the USA, the urinary levels of DBP
FIGURE 48.1 DEHP and metabolites used to estimate DEHP exposure.
TABLE 48.1 Concentrations of DEHP in various environmental
matrices (Clark et al., 2003b)
Matrix Mean Range
Drinking water (μg/L) 0.55 0.16–170
Sediments (μg/g) 1.4 0.0003–218
Soil (μg/g) 0.03 0.03–1280
Outdoor air (ng/m3) 5 <0.4–65
Indoor air (ng/m3) 109 20–240
Dust (g/kg) 3.24 2.38–4.1
Wastewater (μg/L) 27 0.01–4400
Sludge (g/kg) 0.3 0.0004–58.3
metabolites are significantly higher in women of reproduc-
tive age (20–40 years) compared with concentrations in
males or other age groups (Blount et al., 2000) probably orig-
inating from beauty products containing DBP (Lowekamp-
Swan and Davis, 2003). It may be difficult for people to
avoid exposure from cosmetic products even though they
are subjected to strict labeling requirements. In one study,
72 products, purchased directly from stores, were analyzed
for their phthalate ester content. Despite phthalates not
being identified on any of the labels, they were present in
52 of the products (http://
fall02pretty.html). Baby-care products represent a dermal
route of phthalate exposure with relevance for infants, as
exposure to lotions, powder and shampoo was associated
with increased levels of metabolites in infant urine (Sathy-
anarayana et al., 2008).
Neonates may be exposed to high doses during blood
transfusions and other life-saving procedures (Table 48.2)
because medical devices for administration of medicines or
nutrients may contain high levels of DEHP (20–40%) resulting
in MEHP concentrations several orders of magnitude higher
than the general population (Center for Devices and Radio-
logical Health, 2001). In addition, adults are exposed to DEHP
doses in excess of the TDI when receiving blood transfusions
or during kidney dialysis ( Dine et al., 2000; Buchta et al., 2003,
2005; Koch et al., 2005b).
Exposure assessment
Historically, exposure assessment relied on concentrations
found in environmental samples and food. In order to esti-
mate internal levels based on external exposures, not only do
phthalate levels in water and different food products need
to be known, but also ingestion and inhalation rates are
required. In contrast, human biomonitoring assesses expo-
sure by measuring the chemicals in blood or urine (Table
48.3). This method allows the assessment of human exposure
without knowing external exposure. Biomonitoring can be
used to compare exposure levels in the general populations
with specific subpopulations.
For risk assessment biomonitoring/biomarker of expo-
sure measurements are used to determine the exposure level,
which can be compared with toxicological data obtained
from previous research. During the last two decades urinary
concentrations of phthalates have been measured as part of
biomonitoring studies and used to determine exposure levels
of the general population around the world (Koch and Cala-
fat, 2009). This data demonstrated that exposure to DEHP
from food consumption was in the range of the tolerable
daily intake (TDI) and at least some individuals exceed TDI
values on single occasions (Fromme et al., 2007c) suggesting
that background exposure may represent a health risk for the
general population.
Exposure assessment is also a vital component in envi-
ronmental epidemiologic studies. Inadequate exposure esti-
mation can lead to exposure misclassification which in turn
leads to wrong conclusions regarding associations between
exposure and outcome. Because phthalates have short bio-
logical half-lives and are quickly excreted from the body, it
may be difficult to determine exposure over time by mea-
suring a person’s exposure at a single time point. However,
Hauser et al. (2004) documented high day-to-day and month-
to-month variability in phthalate concentrations, but still the
authors showed that a single measurement may adequately
predict average concentrations over a 3-month period for low
TABLE 48.2 The levels of urinary DEHP metabolites measured in patients undergoing medical treatments
Intervention Exposure References
Infants undergoing multiple medical procedures 6 mg/kg bw/day Silva et al. (2004b); Koch et al. (2006)
Parenteral nutrition to preterm neonates 20 mg/day Loff et al. (2000, 2002); Subotic et al. (2007)
Blood transfusion in adults >4 mg/kg bw/day FDA (2004)
Blood transfusion in premature infants 20–fold increase Calafat et al. (2004)
Kidney dialysis patients 0.8 mg/kg bw/treatment Dine et al. (2000)
TABLE 48.3 Markers of DEHP exposure measured in a variety of matrices to assess exposure to DEHP
Marker Marker type Matrices References
DEHP Parent diester Environmental samples, serum McKee (2004)
MEHP Monoester metabolite Serum, urine, amniotic
fluid, saliva, breast milk
Silva et al. (2004)
Calafat et al. (2004)
Kato et al. (2004)
Silva et al. (2005)
5-OH-MEHP Oxidized monoester
Serum, urine, amniotic
fluid, saliva, breast milk
Kato et al. (2004)
Silva et al. (2005)
Barr et al. (2003)
5-oxo-MEHP Oxidized monoester
Serum, urine, amniotic
fluid, saliva, breast milk
Kato et al. (2004)
Silva et al. (2005)
Barr et al. (2003)
2-cx-MEHP Oxidized monoester
Serum, urine, amniotic
fluid, saliva, breast milk
Koch et al. (2005a,b)
Preuss et al. (2005)
5-cx-MEHP Oxidized monoester
Serum, urine, amniotic
fluid, saliva, breast milk
Koch et al. (2005a)
Preuss et al. (2005)
molecular weight phthalates. Because single urinary mea-
sures are less predictive for high molecular weight phthalates
it is suggested that a measurement of a second urine sample
after 30 days be taken (Hauser et al., 2004).
The selection of the proper biomarker which reflects
the actual internal exposure is dependent of the individual
phthalate. In a first rapid step, phthalate diesters are cleaved
to the respective hydrolytic monoesters when passing bio-
logical membranes into the body, followed by a second
step with formation of oxidative metabolites. In a third step
this metabolite can be conjugated with glucuronic acid and
finally excreted (Koch and Calafat, 2009). The rapid transfor-
mation of the intact diester to the hydrolytic monoester indi-
cates that the levels would be very low, rather transient or
artefacts of analytical background contamination suggesting
the diesters as improper biomarkers of exposure (Koch and
Calafat, 2009).
Because the low molecular weight phthalates are more
water soluble they are mostly excreted as the primary metab-
olites (hydrolytic monoesters) while more lipophilic high
molecular weight phthalates need to be transformed to the
secondary oxidative metabolites to increase water solubility
so they can be metabolized via urine. Therefore, using the
hydrolytic primary metabolites may underestimate the inter-
nal exposure of high molecular weight phthalates. For exam-
ple, approximately 70% of an oral dose of DBP (low molecular
weight phthalate) while only 10% of DEHP and 2% of DiNP
(high molecular weight phthalates) are excreted in urine as
the primary hydrolytic monoester (Koch and Calafat, 2009).
Therefore, the proper biomarker should be chosen based on
the individual phthalate to be studied.
Exposure estimates based on DEHP levels in
environmental samples and food
Based on elaborate exposure assessments it was estimated
that more than 90% of the DEHP intake in adult humans is
from food (Clark et al., 2003a, b; Wormuth et al., 2006), whereas
food intake accounts only for 44 and 60% for formula-fed
and breast-fed infants, respectively (Clark et al., 2003a, b).
DEHP concentrations were reported for a great variety of
environmental and food matrices (Table 48.4). Food ingested
by adults from Denmark was calculated to contain less than
0.19 μg/g DEHP resulting in a minimum and maximum daily
intake of 2.7 and 4.3 μg/kg body weight (Petersen and Brein-
dahl, 2000). A more recent estimation based on concentra-
tions in German diets gave comparable results with a range
from 1 to 4.2 μg/kg body weight/day (Fromme et al., 2007b),
which is less than 10% of the TDI of 50 μg/kg body weight/
day. In Japan daily intake up to 11.8 μg DEHP/g was esti-
mated which was attributed to DEHP in the wrappings of
pre-packaged meals (Tsumura et al., 2001a). Heating of food
in a microwave while in contact with PVC significantly ele-
vated phthalate concentrations in packed lunch preparations,
increasing the intake to 92% of TDI (Chen et al., 2008) while
a positive correlation was found between packing date and
DEHP concentrations in curry paste (Kueseng et al., 2007).
Tsumara et al. (2001b) found high levels of DEHP originating
from disposable gloves used in the preparation of the food.
Regulation of these type of gloves by the Japanese Ministry
of Health, Labour and Welfare reduced the estimated daily
intake from 7.4 to 2.3 μg/kg/day after the regulation was
instituted (Tsumura et al., 2003).
Although human DEHP exposure starts in utero as shown
by placental transfer (Shea, 2003; Latini et al., 2003b), it was
suggested that exposure via breast milk to newborns is mark-
edly higher (Calafat et al., 2004; Main et al., 2006). Babies are
also exposed via infant formula containing DEHP (Petersen
and Breindahl, 2000), and it was calculated that the infants
receive comparable doses via formula and breast milk (Clark
et al., 2003a; Latini et al., 2004). The calculated postnatal expo-
sures are well below the European Commission TDI and US
EPA oral reference dose (RfD) (Table 48.5).
Exposure estimates based on biomarkers
The predictive value of a single urine measurement in char-
acterizing exposures to DEHP as high, medium or low over
the course of 3 months was highest for the secondary metab-
olites and lowest for MEHP (Hauser et al., 2004). Urinary
levels for 5-OH-MEHP and 5-oxo-MEHP were 10-fold higher
than MEHP levels (Kato et al., 2004) and the sum of 5-OH-
MEHP, 5-oxo-MEHP, 5-cx-MEPP and 2-cx-MMHP repre-
sents about 70% of the DEHP excretion compared to only 6%
excreted as MEHP (Koch et al., 2006) indicating that DEHP
metabolites in urine provide a more reliable estimate of the
DEHP exposure (Barr et al., 2003; Koch et al., 2003b, 2005a;
Weuve et al., 2006).
TABLE 48.4 Selected food concentrations of DEHP (μg/g)
Food Mean Range References
Baby food 0.12 0.01–0.6 Clark et al. (2003a)
Breast milk 0.062 0.01–0.6 Clark et al. (2003a)
Cereals 0.05 0.02–1.7 Clark et al. (2003a)
Cream 0.2–2.7 Sharman et al. (1994)
Dairy (not milk) 0.96 0.059–16.8 Clark et al. (2003a)
Egg 0.12 <0.01–0.6 Clark et al. (2003a)
Fats and oils 2.4 0.7–11.9 Clark et al. (2003a)
Infant formula 0.12 <0.012–0.98 Clark et al. (2003a)
Meat 0.05 <0.1–0.8 Clark et al. (2003a)
Milk (cow) 0.035 <0.05–1.4 Clark et al. (2003a)
Milk (human) 0.22 0.16–0.4 Zhu et al. (2006)
Poultry 0.9 0.05–2.6 Clark et al. (2003a)
Vegetables 0.048 0.0098–2.2 Clark et al. (2003a)
The ratio of oxidative metabolites to monoester metabo-
lites changes almost linearly with age group but not with
gender or ethnicity with children aged 6–11 years producing
a larger fraction of oxidative metabolites than adolescents
or adults (Silva et al., 2004a, b; Fromme et al., 2007a). Koch
et al. (2003a, b) determined a median DEHP intake of 13.8 μg/
kg/day based on urinary oxidative metabolites of DEHP,
5OH-MEHP and 5oxo-MEHP, in male and female Germans
(n = 85; aged 18–40). Twelve percent of subjects exceeded the
TDI of the EU-CSTEE (50 μg/kg/day) and 31% of the subjects
exceeded the US EPA oral RfD of (20 μg/kg/day).
Fromme et al. (2007b) compared the daily DEHP intake
based on concentration in food and on urinary metabolites
in the same study subjects and found almost equal estimates
independent of quantification method, confirming food as
the major source of DEHP exposure (Table 48.5). However, in
the same study, daily DnBP and DiBP intake estimated from
levels in food was lower than TDI calculated by the biomoni-
toring approach suggesting that sources other than food con-
tribute to human phthalates body burden. In a retrospective
human biomonitoring study, Wittassek et al. (2007) extrapo-
lated the daily parent phthalate exposure in Germany based
on the levels of urinary metabolites in samples collected
between 1988 and 2003, and found comparable exposure
doses as for studies using external doses to assess the daily
intake (Table 48.5). In the same study a downward temporal
trend of 40% in the levels of DEHP and DnBP from 1996 to
2003 was observed.
No human in vivo dermal absorption studies are available.
However, in vitro comparison of absorption of phthalate esters
through rat and human skin showed that phthalates were
absorbed faster through rat skin (Scott et al., 1987). Never-
theless dermal absorption in rodents is relatively slow (Elsisi
et al., 1989), but once absorbed phthalates are distributed in the
same manner as orally administered compounds. In guinea-
pigs only 3 and 21% of the applied dermal dose was absorbed
and excreted after 1 and 7 days, respectively (Ng et al., 1992).
There is no study available on absorption of inhaled phthal-
ates in humans. However, indirect evidence for absorption
via the lungs was observed in infants ventilated with PVC
respiratory tubes, and in workers occupationally exposed to
phthalates. In infants and workers significantly higher levels
of MEHP and secondary metabolites were detected in urine
suggesting that intake via the inhalation route occurs (Kav-
lock et al., 2002; Pan et al., 2006). Furthermore, radiolabeled
DEHP was rapidly absorbed in rats exposed via inhalation
(General Motors, 1982).
Phthalates are absorbed from rat intestine in a wide con-
centration-dependent range, mainly in the form of monoes-
ters, due to rapid hydrolysis by gut lipases. DEHP was not
detected in urine of any of the species studied, but was found
in feces in amounts inversely related to the degree of metab-
olites in urine (Kavlock et al., 2002) suggesting little absorp-
tion of the parent compounds from the gastrointestinal tract.
More than 90% of DBP and 40–50% of DEHP incorporated in
feed were detected in urine following oral administration to
rodents, indicating that phthalates in food are well absorbed
(Kluwe, 1982). Koch et al. (2006) estimated oral absorption rate
for DEHP in a healthy Caucasian male volunteer by measur-
ing the levels of metabolites in urine. After 24 h 67% of the dose
was excreted in urine followed by an additional 3.8% on the
second day, indicating that the majority of the ingested DHEP
is systemically absorbed and excreted in urine in humans.
Once absorbed, phthalates and their metabolites are distrib-
uted throughout the body in all tissues (Table 48.3). Several
studies of DEHP distribution in different species indicate
highest concentrations in liver and kidneys. In addition, in
humans, phthalates are also detected in seminal fluid, amni-
otic fluid, breast milk, saliva and placenta (Koch et al., 2003a,b;
Calafat et al., 2004; Silva et al., 2004a,b, 2005; Mortensen et al.,
2005). No significant cumulative accumulation in tissues
TABLE 48.5 Exposure estimates based on DEHP concentration in environmental samples and in (A) food and on urinary metabolites (B)
Life stage DEHP intakeμg/kg/bw/day References
A: Concentration in food
USA 20–70 years 8.2 Clark et al. (2003a)
Denmark adults 2.7–4.3 Petersen and Breindahl (2000)
Germany 14–60 years 2.5–4.3 Fromme et al. (2007b)
USA 12–19 years 10.0 Clark et al. (2003a)
USA 5–11 years 18.9 Clark et al. (2003a)
USA USA 25,8 Clark et al. (2003a)
USA formula fed 0–6 months 5.0 Clark et al. (2003a)
USA breast feed 0–6 months 7.3 Clark et al. (2003a)
UK formula feed (0–3 months) (3–12 months) (13.0) (8.0) Latini et al. (2004)
UK breast feed (0–3 months) (3–12 months) (21.0) (8.0) Latini et al. (2004)
B: Urinary metabolites
USA 20–60 years 0.7–3.6 Wittassek et al. (2007)
Germany 14–60 years 2.2–7.7 Fromme et al. (2007)
German students 20–29 years 2.7–6.4 Wittassek et al. (2007)
German children 2–14 years 4.3–15.2 Wittassek et al. (2007)
USA pregnant women 20–40 years 1.32–9.32 Fromme et al. (2007b)
USA medical exposure neonates 130–6,000 Calafat et al. (2004)
was noted with less than 1% of the dose retained a few days
after administration (Kavlock et al., 2002). In mice, however,
Tomita et al. (1986) found specific sequestration in pancreas.
Phthalates are transported to the fetus via the placenta,
indicating that exposure to these chemicals occurs during
intrauterine life (Latini et al., 2003b, 2006; Mose et al., 2007),
and fetal phthalate levels correlate with maternal concentra-
tions. In addition, phthalates were also detected in human
breast milk, which is the major source of nutrition for infants
(Mortensen et al., 2005; Zhu et al., 2006). Transfer of phthal-
ates to the fetus via the placenta and the newborn via breast
milk was also documented in rodents (Srivastava et al., 1989).
Biotransformation and excretion
Following oral ingestion, diester phthalates undergo a rapid
cleavage into their monoester metabolites by non-specific
esterases and lipases in the gastrointestinal tract. Human
neonates have lower levels of pancreatic lipase compared to
adults suggesting a reduced metabolic capacity in babies. Fol-
lowing absorption, the monoesters are further metabolized
by various oxidation and hydroxylation reactions resulting
in secondary metabolites, which are excreted via urine or
conjugated to glucuronic acid before excretion (Table 48.1). In
children, the glucuronidation pathways are not fully mature
until they are 3 months old, suggesting that this important
clearance mechanism is not fully available to neonates and
young infants, which may increase the internal doses of toxic
metabolites (Cresteil, 1998).
In a human study, two male volunteers received DEHP
and approximately 13% of the dose was excreted in urine
within 24 hours when only the levels of MEHP were analyzed
(Schmid and Schlatter, 1985). In contrast, when the secondary
metabolites (5OH-MEHP), mono(2-ethyl-5-oxohexyl)phthal-
ate (5oxo-MEHP), mono(2-ethyl-5-carboxypentyl)phthalate
(5cx-MEPP) and mono[2-(carboxymethyl)hexyl]phthalate
(2cx-MMHP) were included as markers of exposure, around
70% of the administered dose was detected in urine of a male
volunteer indicating that the previous focus on only mono-
esters may significantly underestimate the level of exposure
(Koch et al., 2003b, 2006; Fromme et al., 2007b).
As discussed above, the low molecular weight phthal-
ates, including among others DEP and DBP, appear to be
less transformed to the oxidative secondary metabolites
indicated by the observation that 70% of an oral dose of DBP
(four carbons in the alkyl chain) is excreted in the urine as
the primary monoester (Anderson et al., 2001). In contrast,
for the high molecular weight phthalate DiNP with 10 carbon
atoms in the alkyl chain, no significant levels of the hydro-
lytic monoester are excreted in urine (Koch and Calafat, 2009)
suggesting that the monoesters are inadequate as biomarkers
of exposure for high molecular weight phthalates. For risk
assessment and epidemiologic studies recent research has
focused on identifying and characterization of suitable oxi-
dative metabolites (Koch and Calafat, 2009).
There is substantial evidence available which shows that
the C4–C6 phthalates, DBP, BBP and DEHP, produce simi-
lar alterations in male reproductive tract functions, and it
was therefore suggested that there may be a similar mode of
action. The underlying mechanisms of these effects are not
clear but have been the focus of numerous investigations. Ini-
tial mechanistic studies focused on phthalates acting as envi-
ronmental estrogens or antiandrogens. However, by using
estrogenic and androgenic screening assays it was demon-
strated that only the unmetabolized phthalates are able to
bind to steroid receptors whereas the monoesters which are
absorbed exert little or no affinity for ER and AR, indicating
a lack of receptor-mediated effects under in vivo condition
(David, 2006).
A key mechanistic step for phthalate toxicity is forma-
tion of the monoester prior to absorption from GIT. In vitro
studies with cultured testicular cells showed that MEHP is a
more potent testicular toxicant compared to DEHP. Another
important concept for phthalate toxicity is that the modes
of action appear to depend upon developmental timing and
dosing. During rodent development, two sets of Leydig cells,
which are the predominant testosterone producing cells in
males, develop successively. The first set, fetal Leydig cells,
differentiate into fully competent steroidogeneic cells at ges-
tation day 12 in rats (Huhtaniemi and Pelliniemi, 1992). Tes-
tosterone and INSL3 produced by these cells are critical for
normal male secondary differentiation (Huhtaniemi and Pel-
liniemi, 1992). Fetal Leydig cells remain in the rodent testis
until after birth followed by a substitution by adult Leydig
cells which start to appear at postnatal day 11. In humans,
fetal Leydig cells differentiate around week 8 of intrauterine
life and persist at least for a few months after birth. At the
beginning of the pubertal period, or following choriogonad-
otropin hormone (hCG) administration during childhood,
cells of mesenchymal origin start to proliferate and differ-
entiate into adult-type Leydig cells (Codesal et al., 1990; Ari-
yaratne et al., 2000; Habert et al., 2001). Exposures of fetal vs.
adult Leydig cells are proposed to produce different adverse
effects. For example, phthalates were found to induce aggre-
gation of fetal Leydig cells (Mylchreest et al., 2000; Barlow
et al., 2004; Ge et al., 2007), whereas such an effect was not
observed in adult Leydig cells (Dostal et al., 1988; Parks
et al., 2000; Ge et al., 2007). Furthermore, when exposed to
high doses, phthalates inhibit testosterone production in
both fetal and adult Leydig cells. In contrast, low doses of
phthalates increase the number of adult Leydig cells and
enhance testosterone production leading to advanced onset
of puberty, whereas high doses reduce testosterone synthesis
and delay puberty (Ge et al., 2007). No such biphasic pat-
tern was documented in fetal Leydig cells. An explanation
for these contradictory results may be that DEHP interferes
with different mechanisms during different life stages (e.g.,
fetal life vs. adulthood). This phenomenon, that effects do
not increase monotonically with dosage, is frequently noted
in toxicological and pharmacological studies (Calabrese and
Baldwin, 2003; Kohn and Melnick, 2007).
The mode of action underlying phthalate toxicity on Ley-
dig cells remains unclear. However, it was proposed that
peroxisome proliferator-activated receptors (PPAR) may be
involved in testicular toxicity following phthalate exposure
because many phthalate monoesters including MEHP and
MDP were found to induce PPAR in vitro (Hurst and Wax-
mann, 2003; Corton and Lapinskas, 2005; Ge et al., 2007) and
that these receptors may be expressed differently in Leydig
cells and other testicular cells at different life stages. The PPAR
family contains three subtypes, PPARα, PPARβ and PPARγ,
encoded by different genes, and the activation of PPAR
regulates genes which are generally involved in metabolism,
cell growth and stress responses (Lemberger et al., 1996). In
rats, Schultz et al. (1999) showed that both PPARα and PPARγ
are strongly expressed in fetal Leydig cells whereas only
PPARα is expressed in Leydig cells of adult rats. Although
apparently not expressed in adult Leydig cells, testis hom-
eogenates from juvenile male rats exposed from week 4 to
8 of age to DEHP demonstrated a concentration-dependent
increase in PPARγ protein, indicating that other testicular
cells may contain PPARγ. Further, in the same study, the
levels of RXRalpha and several apoptotic proteins were also
upregulated in testes suggesting that DEHP exposure may
induce the expression of apoptosis-related proteins in tes-
tes through induction of PPARγ (Ryu et al., 2007a). Juvenile
exposure to DBP was also found to raise the expression of
the PPARγ gene in testes of Sprague-Dawley rats (Ryu et al.,
2007b). A possible involvement of PPAR in male reproductive
toxicity is supported by the observation that DEHP exposure
did not affect testosterone production in PPARα-null mice to
the same extent as in wild-type mice (Corton and Lapinskas,
2005). Maloney and Waxman (1999) also showed that MEHP
activates human PPARα and PPARγ in COS-1 cells. How-
ever, there are no clear data to support or reject the suggested
involvement of PPARs in the dysgenesis of the male repro-
ductive tract following fetal exposure (David, 2006).
DEHP or DBP was also found to dysregulate genes
involved in cholesterol transport across mitochondrial
membrane and steroidogenic enzyme activities in Leydig
cells with subsequent decrease in testicular testosterone
production, which may adversely affect the differentiation
of androgen-dependent tissues (Akingbemi et al., 2001;
Shultz et al., 2001; Barlow et al., 2003). Ryu et al. (2007b)
exposed juvenile Sprague-Dawley rats to DBP for 30 days
and observed that genes involved in steroidogenensis (SR-
B1, StAr, P450scc, CYP17, CYP19) were expressed differently
compared to control.
Reduced testosterone levels were also observed in male
rats exposed prenatally to 50 mg/kg DBP (Lehmann et al.,
2004) and 10 mg/kg DEHP (Akingbemi et al., 2001). How-
deshell et al. (2008) exposed fetal rats to a mixture of phthal-
ates containing DEHP, BBP, DBP and di-isobutyl (DiBP) and
co-administration of these phthalates reduced testosterone
production in a dose additive fashion. Evidence indicated
that the individual phthalates induce cumulative, dose addi-
tive effects on fetal testosterone production when adminis-
tered as a mixture due to the similar mechanisms of action.
Furthermore, Borch et al. (2006) showed that DiBP exerted
similar effects as DEHP, DBP and DiNP on rat fetal testicu-
lar testosterone production and testicular histopathology
following exposure in utero. The proteins StAR and P450scc
involved in steroid synthesis in testes were also decreased by
DiBP as shown for other phthalates. These findings contra-
dict the conclusion in an EFSA (2005b) report which stated
that no group-TDI could be allocated for phthalates because
different mechanisms are involved.
Elevation of glucocorticoid has been associated with sup-
pression of reproductive functions, obesity and type 2 dia-
betes. Glucocorticoids are controlled by 11β-hydroxysteroid
dehydrogenase (11β-HSD) which catalyzes conversion of
active cortisol to inert steroids. Using human and rat kid-
ney microsomes and murine gonadotrope LβT2 cells, it was
demonstrated that MEHP, di-n-butyl-phthalate (DBP), dipro-
pyl phthalate (DPrP) and di-cyclohexyl phthalate (DCHP)
downregulated 11β-HSD leading to increased cortisol
concentration (Hong et al., 2009). Downregulation of 11β-
HSD in Leydig cells accompanied by increased cortisol is
shown to inhibit Leydig cell testosterone production (Hu
et al., 2008), suggesting a new mechanism for testicular
toxicity induced by phthalates.
In addition to the adverse effects of androgens on steroido-
genesis, phthalates were also found to reduce the expression
of insulin-like factor 3 (insl3) gene (Wilson et al., 2004; McKin-
nell et al., 2005, Ryu et al., 2007b). The insl3 protein is involved
in the initial stages of testicular descent and together with
androgens controls normal testicular descent into the scro-
tum, while failure of this process results in cryptorchidism
(Sharpe, 2006). Decreased insl3 expression observed after
fetal exposure to DEHP, DBP and BBP may be related to
increased incidence of cryptorchidism (David, 2006). Knock-
outs of this gene in mice show complete cryptorchidism (Nef
and Parada, 2000).
The observation that phthalate exposure reduces Sertoli
cell proliferation and triggers formation of multinucleated
gonocytes may be an indirect effect due to reduced androgen
production by Leydig cells (David, 2006). However, direct
effects of phthalates on Sertoli cells have also been proposed.
DEHP exposure of neonatal Sertoli cells decreased the cell
cycle gene cyclin D2 and reduced proliferation. Furthermore,
Kang et al. (2002) demonstrated that DEHP (500 microM)
inhibited apoptosis in TM5 Sertoli cells, preceded by the
downregulation of gap junctional intercellular communica-
tion. Sertoli cell communication is essential for proliferation
of gonocytes, and a reduction of Sertoli cells may result in
dysgenesis of gonocytes (David, 2006) similar to changes
observed when rats are exposed to phthalates (Mylchreest
et al., 2002; Barlow et al., 2003; Fisher et al., 2003).
DNA methylation controls gene expression and epigen-
etic modulation of DNA, which may represent an important
newly discovered mechanism for endocrine disruption of
gene expression during development leading to permanent
effects (Gore, 2008). Treatment of MCF7 cells with BBP or
DBP at 10−5 M led to the demethylation of ERalpha promoter-
associated CpG islands suggesting that phthalates may dis-
rupt endocrine function by epigenetic mechanisms (Kang
and Lee, 2005).
Mechanisms of action in females
As primary effects of phthalates on female rats appear to
be associated with reduction in plasma estradiol concentra-
tions, different phthalates were tested for their effects on
estradiol levels. To study whether reduced production and/
or increased metabolism contributed to the lower estra-
diol levels, the concentrations of estradiol and estrone, the
primary metabolite, were determined and both DEHP and
DBP decreased the estradiol to estrone ratio indicative of an
increased estrogen metabolism following exposure to these
compounds (Lovekamp-Swan and Davis, 2003). In contrast,
MEHP, the active metabolite of DEHP, was the only phthal-
ate monoester that significantly decreased estradiol pro-
duction in rat granulosa cells in vitro suggesting that DEHP
lowered estradiol levels by interfering with both production
and metabolism whereas DBP affected only the metabolic
Mechanistic studies with primary rat granulosa cell cul-
tures demonstrated that DEHP decreased estradiol produc-
tion by reducing the levels of aromatase, the enzyme that
converts testosterone to estradiol (Lovekamp and Davis,
2001). This dysregulation of aromatase may be due to DEHP-
induced activation of PPARα and PPARγ in rat granulosa
cells. By comparing known PPAR ligands with DEHP, Love-
kamp and Davis (2001) showed that phthalates inhibit aro-
matase expression and activity in rat granulosa cells in the
same manner. By activating PPARγ, DEHP disrupted the
critical timing and growth of the ovarian follicle. Sufficient
estradiol production prior to ovulation is thus critical for
induction of the ovulatory LH surge. After ovulation aroma-
tase activity is rapidly decreased both by increased degrada-
tion of mRNA and inhibition of transcription (Fitzpatrick
et al., 1997). PPARγ activation also diminishes the activation
of the aromatase gene and increases the turnover of mRNA
(Mu et al., 2001). This is probably part of the normal program
involved in LH-induced luteinization. Consequently, DEHP
triggers granulosa cell differentiation without ovulation by
activation of PPARγ (Lovekamp-Swan and Davis, 2003).
Based on this information from studies in rats it is likely that
DEHP also affects human female reproductive function since
(1) PPARγ is expressed in human and rodent ovaries and (2)
DEHP stimulates transcriptional activity of both human and
rodent PPARγ (Lovekamp-Swan and Davis, 2003).
In animal studies phthalates produced a variety of adverse
effects including liver tumors in rats and mice as well as Ley-
dig cell and pancreatic cell tumors in rats (Kavlock et al., 2006).
Based on those observations the US Environmental Protec-
tion Agency (EPA) classified the risk for DEHP carcinogenic-
ity as B2 (probable human carcinogen) in 1993. However, in
2000, the International Agency for Research on Cancer (IARC)
downgraded the level of potential health risks of DEHP from
2B (possibly carcinogenic to humans) to 3 (not classifiable
as to carcinogenicity to humans) based on an expert panel’s
acceptance that DEHP induces liver tumors in rodents by a
mechanism dependent on PPARα activation that is not rele-
vant in humans (Guyton et al., 2009). This decision has been
variously argued by several scientists based on the lack of
experimental studies that support the PPARα hypothesis to
dismiss the human relevance of effects observed in labora-
tory animals (Guyton et al., 2009). Furthermore, two recent
studies showed that DEHP induces liver tumors in PPARα
knockout mice suggesting that DEHP may produce liver
tumors by alternative mechanisms (Ito et al., 2007). Based on
those uncertainties, in contrast to IARC’s classification, the
Japan Society for Occupational Health has maintained the 2B
class of DEHP carcinogenicity because of the obvious rodent
carcinogenicity (Japan Society for Occupational Health 2007,
More recently phthalates have attracted special atten-
tion because of their possible endocrine disrupting poten-
tial that may disrupt biological functions including sexual
development and reproductive functions in adults. The first
finding of phthalate induced reproductive toxicity were tes-
ticular injury in experimental animals (Schaffer et al., 1945).
Later male reproductive and developmental effects that are
observed in animal models include pathological changes in
testes and male reproductive accessory glands, hypospadias,
cryptorchidism, retention of nipples, reduced ano-genital dis-
tance (AGD) and reduced sperm production. These postnatal
changes were preceded by an impairment of fetal Leydig cell
function associated with reduced testosterone production
(Mylchreest et al., 2000; Parks et al., 2000; Fisher et al., 2003;
Foster, 2006; Latini et al., 2006; Sharpe, 2006) and insulin-like
factor 3 (insl3) mRNA levels (Wilson et al., 2004). Further-
more, fetal exposure to DBP also resulted in testicular Leydig
cell adenomas in adult life (Mylchreest et al., 2000; Barlow
and Foster, 2003; Barlow et al., 2003). At present the general
consensus is that DEHP, DBP, BBP and DINP have potential
to disrupt normal development and reproduction (Fabjan
et al., 2006). However, other phthalates need further evaluation
before definite conclusions can be drawn with respect to their
developmental and reproductive toxicity (Fabjan et al., 2006).
The adverse effects induced by phthalates depend upon
the dose and timing. Previous standard teratology studies in
rats using exposure on gestation day (GD) 6–15 were only
able to demonstrate effects following exposure to high doses
(Ema et al., 1993; Hellwig et al., 1997; Waterman et al., 1999).
However, those studies did not expose pregnant dams during
the period when the reproductive system differentiates (GD
12–20). More recent studies showed that all of the adverse
responses on reproductive development were induced at
lower doses when exposure occurred in late gestation (Myl-
chereest et al., 2000; Gray et al., 2000; Foster et al., 2001, 2006;
Tyl et al., 2004; Ryu et al., 2007a). The most frequently inves-
tigated phthalates are DEHP, DBP and BBP and they were
found to produce almost identical responses with a relative
toxic potency of DEHP > DBP > BBP (Foster, 2005). However,
despite limited data regarding the reproductive toxic poten-
tial of other phthalates, there are reports suggesting that other
phthalates such as di-n-hexyl phthalate (DnHP), di-isobutyl
phthalate (DiBP) and di-isononyl phthalate (DiNP) may also
induce reproductive toxicity (Fabjan et al., 2006; Borch et al.,
2006; Howdeshell et al., 2006).
The effects observed in rodent studies resemble testicular
dysgenesis syndrome (TDS) in humans (Sharpe and Skak-
kebaek, 2008). TDS represents a number of reproductive dis-
orders in humans, hypothesized to be induced by exposure
to endocrine disruptive chemicals during development. This
syndrome was proposed to explain the reported increase in
male reproductive deficits such as decline in sperm count,
elevated incidence of testicular and prostate cancers and
higher incidence of cryptorchidism and hypospadias (Sharpe
and Skakkebaek, 2008). Although the present data suggest a
link between phthalate exposure and adverse human repro-
ductive health effects, the evidence is too limited to estab-
lish cause–effect relationships between human exposure and
these effects. However, the attempt to relate toxic events dur-
ing fetal and neonatal life to subsequent adult diseases is an
exceedingly difficult challenge for epidemiology (Foster, 2005;
Latini, 2006; Skakkebaek et al., 2006; Matsumoto et al., 2008).
Although far less investigated, there are reports that
suggest that phthalates may adversely affect other func-
tions and systems such as thyroid signaling, immune
functions, metabolic homeostasis, behavior and neuro-
nal development and functioning (Román, 2007; Stahlhut
et al., 2007; Jaakkola and Knight, 2008; Meeker et al., 2009;
Bornehag and Nanberg, 2010). Due to the observation that
low testosterone in adult males is associated with obesity
and increased risk of type 2 diabetes it has also been sug-
gested that phthalates may represent one of the etiologic
factors for the increased prevalence of these diseases
reported in developed countries (Stahlhut et al., 2007; Latini
et al., 2009 ).
Male human studies
In humans a few studies are available assessing the associa-
tion between developmental phthalate exposure and adult
male reproductive and developmental endpoints such as
hormone levels, adverse semen parameters, time to preg-
nancy and infertility diagnosis (Table 48.6). Various studies
suggested a possible association between phthalate expo-
sure and disturbance of normal sperm function (Matsumoto
et al., 2008) such as fewer motile sperm (Jönsson et al., 2005),
low sperm concentration and motility (Hauser et al., 2006),
sperm malformations (Rozati et al., 2002; Zhang et al., 2006)
and increased DNA damage (Rozati et al., 2002; Hauser et al.,
Two studies reported an association between DEP expo-
sure and increased DNA damage in sperm (Duty et al., 2003;
Hauser, 2008). Urinary MEP levels have also been associated
with enlarged testes, fewer motile sperm and lower serum
luteinizing hormone (LH) levels (Hauser et al., 2005; Jönsson
et al., 2005). Interestingly, no adverse effects of DEP exposure
have thus far been noted in animal studies.
Significant dose–response relationships were found
between DBP exposure and low sperm concentration and
motility in male partners of subfertile couples, whereas in
this study no relationships were observed between DEP
and DEHP and semen quality (Hauser et al., 2006). In
another study both DBP and DEHP were associated with
lower plasma testosterone levels in occupationally exposed
workers (Pan et al., 2006). DEHP was also correlated with
increased DNA damage in a group of men exposed to doses
comparable to those reported for the US general population,
suggesting that exposure to DEHP may affect the popula-
tion distribution of sperm DNA damage (Hauser et al., 2007).
Higher levels of total phthalates (DMP, DEP, DBP, BBP, DOP,
DEHP) were found in a group of infertile men compared
to controls (Rozati et al., 2002). In this group, a significant
correlation was found between sum phthalates and normal
sperm morphology and percent single stranded DNA in the
sperm (Rozati et al., 2002). In a study of phthalate concen-
tration in breast milk, significant negative correlation was
observed between DBP and free testosterone, positive cor-
relation between DEP, DMP, DBP and DiNP and LH and
between DEP and DBP and sex hormone binding globulin
(Main et al., 2006).
Female human studies
In females data are limited to a few studies. Two studies sug-
gested a possible association between phthalates and endo-
metriosis (Cobellis et al., 2003; Reddy et al., 2006). In the first
study higher plasma concentrations of DEHP and MEHP
were found in women with endometriosis (study group)
compared to controls. In the second study higher levels of
DBP, BBP, DOP and DEHP were detected in the study group.
Latini et al. (2003b) found an association between cord blood
concentrations of DEHP and a lower gestational age, and
it was suggested that the increased number of premature
babies might be due to induction of intrauterine inflamma-
tion by DEHP. Disturbance of immune functions may also
trigger the development of endometriosis (Matsumoto et al.,
2008). In Puerto Rico, premature breast development was
associated with high blood levels of DMP, DEP, DBP and
DEHP (Colón et al., 2000). These results were challenged by
McKee (2004) because the levels detected were unusually
high and may reflect contamination of the samples during
analysis, and because no such effects are observed in animal
studies. Recently, Meeker et al. (2009) found significant asso-
ciation between phthalate exposure (DEHP, DBP, DOP, BBzP)
and increased risk for preterm birth among a group of Mexi-
can women.
Human infant studies
Anogenital distance (AGD) is a commonly used endpoint for
hormonally regulated sex differentiation in rodents. AGD in
male rats is normally twice that in females, and a similar sex
difference occurs in humans (Salazar-Martinez et al., 2004). A
negative association between anogenital index (AGI; AGD/
weight) in boys of 2–36 months of age and the levels of MEP,
MBP, MBzP and MiBP detected in their mothers’ urine was
reported (Swan et al., 2005). Furthermore, in a recent study,
Huang et al. (2009) found negative correlations between AGI
in newborns and levels of MEHP and MBP in amniotic fluid.
Main et al. (2006) found no association between phthalate
exposure via breast milk and testicular descent. However, in
the same study, positive correlations were noted between (1)
DEP and DBP and sex hormone binding protein, (2) DBP and
the ratio of LH/free testosterone, and (3) negative correlation
TABLE 48.6 Male reproductive effects in human populations
Compounds Study subjects Significant associated effects References
MEP n = 168 DNA damage in sperm Duty et al. (2003)
n = 234 large testis sperm motility, LH 113 mg/kg bw/day Jönsson et al. (2005)
n = 379 DNA damage in sperm Hauser et al. (2007)
DBP n = 37 semen volume Zhang et al. (2006)
MBP n = 168 sperm concentration, sperm motility Duty et al. (2003)
n = 463 sperm concentration, sperm motility Hauser et al. (2006)
n = 74 plasma free testosterone Pan et al. (2006)
DEHP n = 37 semen volume, rate of sperm malformation Zhang et al. (2006)
MEHP n = 187 straight-line velocity and curvilinear velocity of sperm Duty et al. (2004)
n = 74 plasma free testosterone Pan et al. (2006)
n = 379 sperm DNA damage Hauser et al. (2007)
Sum phthalates n = 21 sperm normal morphology, percent of single-stranded
DNA in sperm
Rozati et al. (2002)
= decrease, = increase
between DBP and free testosterone (Main et al., 2006). The
study did not analyze the secondary metabolites. Further-
more, the samples might have been contaminated because
they used breast pumps containing phthalates. Although
these studies suggest a possible association between expo-
sures to phthalates and adverse effects in humans, no effects
were observed in adolescents exposed as neonates to high
levels of DEHP from medical devices (Rais-Bahrami et al.,
2004). However, in this study, controls were not included, the
levels of phthalates were not measured and the small size of
the study group limits the power to elucidate statistical sig-
nificance (Kavlock et al., 2006).
Animal studies
The existing human data are insufficient to evaluate the
reproductive effects of phthalate exposure in humans. In ani-
mals, however, there are strong indications that phthalates
have the potential to adversely affect normal development
and disrupt reproductive functions (Kavlock et al., 2006). In
a number of studies reproductive and endocrine endpoints
were investigated in rodents exposed during development
(Table 48.7).
Some studies used multiple doses giving the opportu-
nity to establish dose–response relationships. The remaining
investigations are single dose studies focusing on modes of
action for developmental reproductive toxicity. In one study
Sprague-Dawley rats were orally dosed with DEHP at 0, 375,
750 or 1,500 mg/kg body weight/day from GD 3 to PND 21
and endpoints related to sexual development were studied
through puberty and adulthood in male and female off-
spring. The two highest doses reduced the prenatal maternal
weight gain. In the high dose group the number of pups was
reduced and the postnatal mortality was increased at the two
highest dose levels. However, at all doses increased aerolae
or nipple sizes in the male offspring were observed and this
effect was persistent until adulthood. Furthermore, a range
of male accessory reproductive organ developmental effects
was observed at the two highest doses and the changes also
persisted until adulthood. Although not statistically signifi-
cant, similar effects were also found at the lowest dose levels
and these findings were regarded as biologically significant
because of the rarity of such effects. In the female offspring
no effects were associated with the DEHP treatment. Because
effects were observed in all treatment groups the study could
not identify a no-observable-adverse-effect level (NOAEL)
but a lowest-observable-adverse-effect level (LOAEL) at
375 mg/kg body weight/day was identified based on a sig-
nificant decrease in anterior prostate weight and an increase
in permanent nipple retention (Moore et al., 2001). Retained
nipples were also found in Wistar rats orally treated with
1,088 mg/kg body weight/day DEHP during gestation and
lactation. In the same study, focal tubular atrophy was shown
at all doses giving an LOAEL at 113 mg/kg body weight/day
for reproductive toxicity. At this dose no systemic toxicity
was observed (Schilling et al., 2001). In another study using
Sprague-Dawley rats, exposure to 100–500 mg/kg body
weight DEHP reduced Sertoli cell proliferation and increased
the number of large gonocytes containing 2–4 nuclei in their
pups in a dose-dependent manner (Li et al., 2000). Increased
levels of multinucleated germ cells at 125 mg/kg body
weight/day and interstitial hyperplasia at 250 or 500 mg/
kg body weight/day were also observed in offspring of rats
exposed orally to DEHP from GD 7 to 18 (Shirota et al., 2005).
Cammack et al. (2003) studied reproductive development
of Sprague-Dawley rats exposed i.v. or orally for 3 weeks
postnatally to multiple doses of DEHP. Depletion of germi-
nal tubule and decreased seminiferous tubule diameter were
observed in the groups treated with 300 or 600 mg/kg body
weight by both exposure routes. Furthermore, decreased
seminiferous tubule diameter persisted until adult life in the
groups receiving 300 mg/kg body weight DEHP. Repro-
ductive effects of DEHP exposure in feed were investigated
in a comprehensive multiple dose multigenerational breed-
ing study in Sprague-Dawley rats sponsored by the US
National Toxicology Programme (NTP). Adverse reproduc-
tive effects such as seminiferous tubuli atrophy, decrease in
pregnancy index, number of litters per pair, male reproduc-
tive organ weights, sperm counts and sperm motility were
observed in all generations. The NTP-CEHR reviewed the
data from this study, and based on the occurrence of small
reproductive organ weights in F1 and F2 generation the
panel considered 14–23 mg/kg body weight/day to be the
LOAEL giving an NOAEL of 3–5 mg/kg body weight/day
(Kavlock et al., 2006).
TABLE 48.7 Summary of male reproductive toxicity data from studies in rats
Species and dosing Most sensitive outcome Effect level References
Sprague-Dawley: DEHP orally at 0, 375,
750, 1,500 mg/kg bw/day. GD 3–PND 21
Increased aerolae or nipples sizes in the
male offspring
375 mg/kg bw/day
Moore et al.
Wistar rats: DEHP orally at 0, 113, 340,
1,088 mg/kg bw/day
Gestation and lactation
Focal tubular atrophy LOAEL
113 mg/kg bw/day
Schilling et al.
Sprague-Dawley: DEHP orally at 0, 100,
250, 500 mg/kg bw/day. GD 7–18
Reduced Sertoli cell proliferation. Increased
large gonocytes with 2–4 nuclei
100 mg/kg bw/day
Shirota et al.
Sprague-Dawley: DEHP orally at 0, 150,
300, 600 mg/kg bw/day. PND 0–21
Persistent decreased seminiferous tubule
diameter until adulthood
300 mg/kg bw/day
Cammack et al.
Sprague-Dawley; DEHP orally at 1.5, 10,
30, 100, 300, 1,000, 7,500 or 10,000 ppm.
Continously 3 generations
F2 percent mobile sperm NOAEL 100 ppm
(3–5 mg/kg bw/day)
Kavlock et al.
Rat, Long-Evans: DEHP orally at 0, 1, 10,
100 or 200 mg/kg bw/day 14 or 28 days
Decreased 17α-hydroxylase in testis, altered
ex vivo Leydig cell testosterone synthesis
1 mg/kg bw/day
Akingbemi et al.
Sprague-Dawley: DBP injected s.c.
PND 5–14
Reduced weights of testes and accessory
glands. Affected tubules
20 mg/animal Kim et al.
Long-Evans rats orally dosed with DEHP from postna-
tal day (PND) 21 (weaning) to PND 120 showed no signs
of overt toxicity. However, exposure to a dose as low as
10 mg/kg body weight/day reduced Leydig cell testos-
terone production ex vivo, increased serum LH, testoster-
one and 17 β-estradiol and led to Leydig cell hypoplasia,
suggesting an NOAEL of 1 mg/kg body weight/day
(Akingbemi et al., 2004). The increased steroid levels were
probably produced by increased number of Leydig cells
and chronically increased LH levels induced by DEHP
exposure. The rise in testosterone levels found in this
study contrasts with the decrease in testosterone amounts
observed in studies where exposure to DEHP occurred
In a multi-dose study, rats were injected subcutaneously
from PND 5 to PND 14 with corn oil (control) or DBP (5, 10 or
20 mg/animal). DBP exposure (20 mg/animal) significantly
reduced the weights of testes and accessory sex organs in
the same manner as DEHP. These adverse effects persisted
through puberty at PND 42. Histomorphological examina-
tion showed mild diffuse Leydig cell hyperplasia in the inter-
stitium of severely affected tubules on PND 31. Furthermore,
DBP (20 mg/animal) significantly decreased the expression
of androgen receptor (AR), whereas estrogen receptor (ER)
expression and steroidogenic factor 1 (SF-1) expression were
increased in a dose-dependent manner on PND 31 in the rat
testes (Kim et al., 2004).
Effects observed in single dose studies are similar to those
found in the multiple dose level studies. For example, treat-
ment of dams with a single dose of DEHP at 750 mg/kg body
weight in late gestation or early lactation decreased testicu-
lar weights and increased testicular lesions in the offspring.
Additional effects observed in male rats exposed to a single
dose of DEHP at 750 mg/kg body weight included retained
nipples, reduced AGD, lack of testicular descent, agenesis of
accessory reproductive organs and incomplete preputial sep-
aration (Gray et al., 2000; Parks et al., 2000; Borch et al., 2004).
In a 65-week oral toxicity study in marmosets, DEHP was
administered from prior to puberty until young adulthood.
Although a dose-dependent delay in the onset of puberty
was found in males, no testicular toxicity was observed at
the highest dose level (2,500 mg/kg body weight/day) sug-
gesting that male marmosets are less sensitive compared
to rodents. These conflicting results may be explained by a
lower absorption of DEHP from the intestine in marmosets
(Tomonari et al., 2006).
Female animal studies
Although less studied, phthalates were also reported to
induce adverse reproductive effects in females (Table 48.8).
In the 65-week oral DEHP toxicity study in marmosets
(Tomonari et al., 2006), DEHP at 500 mg/kg body weight/
day elevated serum estradiol levels, advanced the onset of
puberty and increased ovarian and uterine weights. In the
same study no differences were observed in weights of the
male accessory sex organs. In female rats gestational and
lactational exposure to DBP induced uterine malformations
and reduced fecundity at the same doses (500 or 750 mg/kg/
day), which produced malformations in males (Mylcherest
et al., 1998). In a multiple dose, two-generation reproductive
study in rats, DEHP exposure reduced the weight and weight
gain in F2 pups in the high dose group (1,088 mg/kg body
weight/day; Schilling et al., 2001). Furthermore, in the high
dose females deficits in growing follicles and corpora lutea
were noted. Exposure of adult Sprague-Dawley rats with
2 g/kg DEHP decreased serum estradiol levels, prolonged
estrus cycles and there was an absence of ovulation. Missing
ovulation led to an absence of the development of corpora
lutea and follicles became cystic. Histological analysis of the
preovulatory follicles showed significantly smaller granulosa
cells in the DEHP-exposed rats (Davis et al., 1994). Perina-
tal exposure to DEHP at 5, 15, 45, 135 or 405 mg/kg body
weight/day was studied by Grande et al. (2006, 2007). There
was delayed puberty at 15 mg/kg body weight/day in female
Wistar rats in addition to inducing a significant increase in
tertiary atretic follicles in adult female offspring exposed to
405 mg/kg body weight/day. An NOAEL of 5 mg/kg body
weight/day was estimated for female reproductive toxicity
(Grande et al., 2006, 2007).
Developmental effects in other organs and
Normal thyroid hormone function was shown to be impor-
tant for reproductive system development and function in
both males and females (Poppe et al., 2007). Furthermore, defi-
ciency of thyroid hormones during critical periods of brain
development both pre- and postnatally is a well- recognized
etiologic factor for brain damage leading to various neu-
rologic disorders (Román, 2007) Serum thyroid hormone
(TH) and thyroid stimulating hormone (TSH) levels were
TABLE 48.8 Summary of female reproductive toxicity data from animal studies
Species and dosing Most sensitive outcome Effect level References
Marmosets: DEHP orally at 0, 100, 500 or
2,500 mg/kg daily 3–18 months
Increased serum estradiol levels, advanced
onset of puberty, increased ovarian and
uterine weights
NOAEL 250 mg/kg bw/day Tomonari et al.
Sprague-Dawley: DBP orally at 0, 250,
500 or 750 mg/kg/day. GD 3–PND 21
Uterine malformations, reduced fecundity NOAEL 100 mg/kg bw/day Mylcherest et al.
Wistar rats: DEHP orally at 0, 113, 340,
1,088 mg/kg bw/day
Gestation and lactation
Deficits in growing follicles and corpora
NOAEL 340 mg/kg bw/day Schilling et al.
Sprague–Dawley: Adults DEHP orally at
2 g/kg DEHP for 12 days
Decreased estradiol, prolonged estrus cycles,
significantly smaller granulosa cells
LOAEL 2,000 mg/kg bw/day Davis et al.
Wistar rats: DEHP orally at 5, 15, 45, 135
and 405 mg/kg bw/day. GD 3–PND 21
Delayed puberty NOAEL 5 mg/kg bw/day Grande et al.
(2006, 2007)
inversely correlated with urinary MEHP concentrations com-
parable with levels reported for the general US population
(Meeker et al., 2007, 2009). In Taiwanese pregnant women,
urinary MBP levels showed significant negative associations
with thyroxine and free thyroxine (Huang et al., 2009). In ani-
mal studies, rats receiving diets with added DEHP had lower
plasma thyroxin (T4) concentrations compared with controls
(Hinton et al., 1986; Poon et al., 1997; Howarth et al., 2001). A
dose-dependent inverse association between DBP and both
triiodothyronine (T3) and T4 has also been reported in male
rats (O’Connor et al., 2002). Furthermore, in a recent in vitro
study the TH disrupting potential of various phthalates was
determined by the effect on the TH-dependent rat pituitary
GH3 cell proliferation (T-screen) and BBP, DBP, DOP, DIDP,
DINP and DEHP was shown to significantly affect the GH3
cell proliferation (Ghisari and Bonefeld-Jorgensen, 2009).
Association between severe hypothyroxemia and neuro-
logical disorders such as reduced IQ and motor deficiency
is well described (Ohara et al., 2004). However, even mild
subclinical maternal hypothyroxemia has been documented
to affect fetal neurodevelopment (Haddow et al., 1999; Pop
et al., 1999). Engel et al. (2009) reported significant association
between prenatal exposure and performance on the Neonatal
Behavior Assessment Scale in a multiethnic cohort from New
York. Based on the reported effect of phthalates on thyroxin
function, the authors suggested maternal hypothyroxemia as
a possible mechanism for the observed effect. Another study
reported association between DEHP and attention-deficit/
hyperactivity disorder (ADHD) in school-age children (Kim
et al., 2009a). Associations between phthalates and hyperac-
tivity are also reported in animal studies (Ishido et al., 2004).
Furthermore, behaviors of young rats were more seriously
affected when exposure occurred during differentiation and
synaptogenesis compared to later life stages suggesting that
the level of vulnerability of the developing brain is depen-
dent on the life stage at which the exposure occurs (Kim
et al., 2009a). Many studies have documented an association
between the midbrain dopaminergic system and the patho-
genesis for ADHD. Tyrosine hydroxylase, the rate-limiting
enzyme of dopamine production, was downregulated in mid-
brain dopaminergic nuclei in mice exposed to doses of DEHP
below the NOAEL (Tanida et al., 2009). Decreased tyrosine
hydroxylase expression was also observed in rats exposed
to dicyclohexylphthalate (DCHP). Furthermore, microarray
data revealed that DEHP and DBP changed the gene expres-
sion of other genes, dopamine receptor D4 (DRD4) and dopa-
mine transporter in the midbrain of rats (Masuo et al., 2004).
Modulation of DRD4 and dopamine transporter can lead to
changes in extracellular dopamine and neuronal dopamine
sensitivity resulting in hyperactivity and impulsivity in rats.
Accordingly, overexpression of dopamine transporter was
one of the consistent findings in ADHD children (Dough-
erty et al., 1999; Kim et al., 2009b). Because thyroid hormones
play crucial roles in brain development of both humans and
experimental animals and since phthalates are suggested to
affect thyroid functions it is likely that prenatal modulation
of dopamine production and sensitivity may be due to dis-
ruption of thyroid hormone signaling. Thus, the reduced lev-
els of tyrosine hydroxylase in midbrain dopaminergic nuclei
observed in rodents exposed to phthalates may be caused by
disruption of thyroid hormone-dependent gene expression
(Tanida et al., 2009).
Because of the documented antiandrogenic effects of cer-
tain phthalates in animal models and the suggestive results
from human epidemiological studies, it has been questioned
whether low testosterone in adult males may be associated
with an increased prevalence of obesity and type 2 diabetes
(Ding et al., 2006). Stahlhut et al. (2007) compared urinary
phthalate metabolites with increased waist circumference
and insulin resistance in adult American men and found
significant positive correlation between those parameters.
Based on their results, it was suggested that phthalates might
contribute to the ongoing increase in prevalence of obesity,
insulin resistance and related clinical disorders (metabolic
syndrome) observed in US humans via endocrine disrup-
tive mechanisms. In rats, prenatal exposure to DiBP reduced
plasma protein levels of insulin in male and female offspring
and plasma testosterone and liver and testis mRNA levels
of PPARα in males (Boberg et al., 2008). The recent environ-
mental obesogen hypothesis suggests that environmental
chemicals may contribute to the development of obesity and
metabolic disorders such as type 2 diabetes by chemical inter-
action with nuclear receptors including steroid receptors,
retionid x-receptor (RXR) and PPARs (Newbold et al., 2007).
Furthermore, reduced leptin levels in gestation lead to adi-
posity in adult offspring (Vickers, 2007). The existing human
and animal data provide preliminary evidence of a potential
contributing role of phthalates in the increasing prevalence of
obesity diabetes and related clinical condition, but additional
studies are needed (Meeker et al., 2009).
Exposure to phthalates during fetal and neonatal life may
also affect the normal development of the immune system
leading to immune system disorders such as asthma and
allergies. In Oslo, Norway, indoor PVC surface materials in
the home were associated with increased risk of bronchial
obstruction in small children (Jaakkola and Knight, 2008).
In Swedish children (aged 3–8 years), significantly higher
concentrations of DEHP and butyl benzyl phthalate (BBzP)
were associated with eczematous and asthmatic symptoms
(Bornehag et al., 2004). Similar effects were observed in Bul-
garian children where high levels of DEHP were associated
with increased asthma/wheezing in a dose-dependent man-
ner (Kolarik et al., 2008). These studies are of cross-sectional
design that reduces the possibility of conclusive results.
However, positive association between PVC flooring and
increased incidence of asthma was found in a Swedish study
with longitudinal design (Larsson et al., 2009). Other short-
comings in these studies include that none of the studies
focus on early-life exposure and that PVC materials instead
of phthalates were used as exposure data (Bornehag and
Nanberg, 2010).
Animal studies using mice models demonstrated that
DEHP as well as DiNP, DnBP and DnOP at low doses admin-
istered subcutaneously (s.c.) or by intraperitoneal injection
and in long-term inhalation protocols produced Th2 response
with induction of IgE and IgG1 (Larsen et al., 2001, 2007; Lee
et al., 2004 ). Furthermore, increased lymphocyte-dependent
production of Th2 cytokines including IL-4, IL-5 and IL-10
following DEHP exposure was also shown in mice (Larsen
et al., 2001, 2007; Lee et al., 2004). In contrast topical applica-
tion or s.c. injection of high doses of DEHP did not induce
IgE and IgG1 (Kimber and Dearman, 2010). Even though the
individual studies differ in several critical aspects such as
strain, antigen, method for sensitization, route of administra-
tion, time of exposure and doses of the phthalates studied,
the overall picture of the results suggests that several phthal-
ates affect Th2 differentiation and Th2 promoted antigen pro-
duction of IgG1 and IgE (Bornehag and Nanberg, 2010).
There is widespread recognition that humans, fish and wild-
life are simultaneously exposed to multiple phthalates as well
as other contaminants on a continuous basis. The chemicals
frequently detected in biological samples include pesticides,
industrial chemicals, pharmaceuticals and hormones. To
date, risk assessments are typically conducted on a chemical-
by-chemical basis and regulatory efforts have not accounted
for realistic environmental exposure to mixture of phthal-
ates and other contaminants. However, recent studies have
documented that individual phthalates induce cumulative,
dose additive effects by the same mechanism of action when
administered as a mixture (Borch et al., 2006; Howdeshell
et al., 2008). These findings contradict the conclusion in an
EFSA (2005b) report which stated that no group-TDI could
be allocated for phthalates because different mechanisms are
involved. Furthermore, exposure to phthalates in combina-
tion with other groups of chemicals including phenols, diox-
ins, pesticides and pharmaceuticals are reported to induce
antagonistic, additive or synergistic effects depending on
the different combinations (Christiansen et al., 2009; Ghisari
and Bonefeld-Jorgensen, 2009; Tanida et al., 2009). These
results indicate that compounds that act by different modes
of action interact when present in combination. The results
also suggest that a modification of the approach for cumula-
tive risk assessments should be considered, from one based
upon “common mechanism of toxicity” to one that includes
the cumulative assessment of chemicals that disrupt devel-
opment of the same tissues or biological systems resulting
in a target organ- and timing-based approach rather than on
a narrow mechanism of toxicity. The cumulative risk assess-
ment could then potentially include all chemicals that target
one system during the same critical developmental period
(Rider et al., 2009).
The five main phthalate plasticizers, DINP, DIDP,
DEHP, BBP and DBP, have all undergone comprehen-
sive European Union Risk Assessments conducted under
European Union Regulation 793/93, by the Scientific
Committee on Toxicity, Ecotoxicity and the Environment
The risk assessments stated that di-isononyl phthalate
(DINP) and di-isodecyl phthalate (DIDP) pose no risk to
either human health or the environment from any current
use. However, in Europe, based on the precautionary princi-
ple, DINP can no longer be used in toys and child-care items
that can be put in the mouth even though the EU scientific
risk assessment concluded that its use in toys does not pose a
risk to human health or the environment. DIDP, however, is
still allowed in toys and child-care items.
For BBP, the conclusion of the assessment of the risks
to human health is that there is at present no need for fur-
ther information and/or testing or for risk reduction mea-
sures beyond those which are being applied. It is further
concluded that there is a need for better information to
adequately characterize the risks to aquatic ecosystem and
terrestrial ecosystems. For example, a long-term fish study
on reproductive and endocrine effects was mentioned as a
specific requirement.
The assessment of the risks of DBP exposure to humans
and aquatic and terrestrial ecosystems stated that there is
no need for further information or testing or risk reduction
measures beyond those which are being applied already.
However, for workers more information is required because
of concerns for general systemic toxicity as a consequence
of repeated dermal exposure arising from aerosol forming
activities and adverse local effects in the respiratory tract as
a consequence of repeated inhalation exposure in all occupa-
tional exposure scenarios.
The risk assessment demonstrates that DEHP poses no
risk to the general population and that no further measures
need to be taken to manage the substance in any of its key
end-use applications. The areas of possible risk identified in
the assessment relate to the use of DEHP in children’s toys
and DEHP is no longer permitted in toys and child-care
articles in the EU. Some localized environmental exposures
near to factories are documented and the European Union
has initiated measures relating to emission controls from
converters’ plants. An EU Scientific Review was requested
to determine whether there may be any risk from the use
of DEHP in certain medical applications (children and neo-
nates undergoing long-term blood transfusion and adults
undergoing long-term hemodialysis). In February 2008, the
EU Scientific Committee on Emerging and Newly Identi-
fied Health Risks (SCENIHR) published an Opinion in
which they said there is reason for some concern for prema-
turely born male neonates although follow-up studies after
high DEHP exposures in neonates do not indicate there is
an effect of DEHP on the development of the human male
reproductive system.
In an EU risk assessment on cumulative effects of mix-
tures of phthalates it was concluded that no group-TDI could
be allocated for phthalates because different mechanisms are
involved (EFSA, 2005b). This conclusion contradicts recent
data which show that different phthalates exert toxicity in a
dose-additive fashion by the same mode of actions.
A major concern regarding the risk assessments con-
ducted by the EU is that experimental and epidemiological
data published after 2000 are not considered. Furthermore,
at present the EU has not initiated any specific action plan on
phthalates (http://
In a review on DEHP in 2006, the National Toxicology
Program, Centre for the Evaluation of Risks to Human
Reproduction (NTP-CERHR) expressed a reduced level of
concern for the effects of DEHP on male offspring exposed
to general population levels during pregnancy and lactation
due to greater confidence in exposure levels in humans and
in effect levels in experimental animals. The expert panel
underlines that it has concern for possible effects on male
children of women undergoing medical treatment during
pregnancy and lactation leading to additional exposure to
DEHP during development (Kavlock et al., 2006). In 2002,
the NTP concluded in its review on the potential human
impact of DBP exposure that, “Based upon recent estimated
DBP exposures among women of reproductive age, the NTP
has some concern for DBP causing adverse effects to human
development, particularly development of the reproductive
A major concern regarding the risk assessments con-
ducted by the EU is that experimental and epidemiological
data published after 2000 are not considered. Furthermore,
at present the EU has not initiated any specific action plan
on phthalates. In contrast the US Environmental Protec-
tion Agency (EPA) recently published their phthalates
management plan that includes eight phthalates (DIBP,
DINP, DIDP, BBP, DBP, DEHP, DnPP, DnOP). In develop-
ing this plan, the EPA considered the toxicity of phthalates,
their prevalence in the environment and their widespread
use and human exposure (
The EPA is concerned about phthalates because of their
toxicity and the evidence of pervasive human and environ-
mental exposure to them. Thus, the EPA intends to initiate
action to address the manufacturing, processing, distribu-
tion in commerce and use of these eight phthalates. The EPA
intends to take action as part of a coordinated approach with
the Consumer Product Safety Commission (CPSC) and the
Food and Drug Administration (FDA).
Because of the reported cumulative effects of mixtures the
EPA has scheduled a major cumulative hazard assessment con-
duced by the CPSC with planned date of completion in 2012.
Because of the reported adverse effects observed in test
animals the Consumer Product Safety Improvement Act of
2008 (CPSIA) banned the use of six phthalates in toys and
child-care items at concentrations greater than 0.1%: DEHP,
DBP, BBP, DINP, DIDP and DnOP. The Food and Drug
Administration (FDA) regulates phthalates in food contact
substances (such as plastic wrap), cosmetics, pharmaceuticals
and medical devices. The FDA announced in June 2008 that it
was reviewing available use and toxicology information asso-
ciated with phthalate exposure from FDA regulated products
to better characterize any potential risk from these uses.
The widespread use of phthalates in consumer products
leads to ubiquitous exposure of humans to these compounds
from fetal life to adulthood. At present, exposure data are
inconsistent because of the different methods used to assess
exposure levels. Oxidized metabolites of DEHP were recently
recognized as the major urinary metabolites in humans, sug-
gesting an underestimation of exposure when using only the
primary metabolite MEHP as a biomarker of DEHP expo-
sure. Furthermore, there are little data available on human
TK and metabolism of other phthalates.
Estimated DEHP concentrations in the general popula-
tion were found to be highest in children and decreased with
age. The human levels approximate currently accepted TDI
(EFSA, 2005a) suggesting that at least in some individuals
TDI are exceeded (Fromme et al., 2007c). However, DEHP
concentrations measured in neonates receiving medical treat-
ments are several orders of magnitude higher than in the
general population. Thus, neonates constitute a population at
particular risk. Other groups, which may be exposed to high
levels, are occupationally exposed workers, adults undergo-
ing medical treatments with medical devices or pharmaceuti-
cal drugs containing phthalates.
The current human toxicological data are insufficient to
evaluate the prenatal and childhood effects following phthal-
ate exposure. Animal data are, however, sufficient to conclude
that DEHP, DBP and BBP are potential reproductive toxicants
in rats. The critical period for effects on male reproductive
development appears to be late gestation and into the imme-
diate postnatal period. Although most of the animal studies
focused on male toxicity there are both human and animal
data available, which indicate that exposure to phthalates
may also affect female reproductive functions.
Recent data demonstrated that individual phthalates
with a similar mechanism of action elicit cumulative, dose-
additive effects on fetal testosterone production and tes-
ticular histopathology when administered as a mixture
(Howdeshell et al., 2008). These findings emphasize the need
for testing combinations of phthalates to better assess the
health risks of known human exposure to multiple sources
of these chemicals. Additional TK and toxicodynamic (TD)
data on less well-studied phthalates are needed. Further,
new endpoints need to be included in the study of putative
phthalate toxicity on humans.
Due to the fact that there are substantial gaps in knowl-
edge in both phthalate levels of exposure and consequent
health effects in humans, additional research is warranted.
1. It is of key importance to improve the knowledge of
human TK and toxicity, specifically during pregnancy and
the nursing period, because in utero and early postnatal
exposure appears to be the most vulnerable period during
2. Well-designed follow-up studies of reproductive system
development and functions in the most heavily exposed
and most vulnerable human populations may address the
question of whether phthalates produce adverse human
reproductive effects. Reproductive developmental toxicity
is well studied in male animals. However, data on female
reproductive toxicity are scarce and need further research.
Further in vitro and in vivo studies are also warranted
to improve the understanding of the modes of action of
phthalates in humans.
3. Most studies focused on adverse reproductive and devel-
opmental effects associated with exposure to single
phthalates. However, because humans are exposed to mix-
tures of phthalates both concurrently and sequentially, and
available experimental evidence suggests that mixtures of
phthalates may induce endocrine disruption in a cumula-
tive fashion, it is necessary to initiate studies which focus
on mixture effects.
4. Phthalates with shorter and longer C backbones may need
further evaluation before definite conclusions are drawn
correlating physicochemical properties to toxicity.
5. It is also important to identify the most reliable biomark-
ers of exposure and the biologic media best suited for bio-
marker analysis. Phthalates occur as mixtures in nature
and it needs to be considered whether a summary of the
different phthalates and their metabolites might be a more
appropriate measurement for biomarkers of exposure.
6. Focus should be extended to endpoints other than repro-
ductive toxicity.
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