Environmental Health Perspectives • VOLUME 115 | SUPPLEMENT 1 | December 2007 69

Monograph

Endocrine disruptors (EDs) are exogenous

compounds that have the potential to interfere with hormonal regulations and the normal endocrine system and consequently

cause health effects in animals and humans

(U.S. Environmental Protection Agency

2000). EDs include environmental persistent

organohalogens, pesticides, and industrial

chemicals such as some plasticizers and surfactants. The involvement of EDs in disruption of development, reproduction, the

immune system, and the neural system has

been supported in a wide range of fish and

animal species, whereas for human beings

the ED hypothesis is still controversial

(Bonefeld-Jorgensen 2004; BonefeldJorgensen and Ayotte 2003; Owens and

Koeter 2003; vom Saal and Hughes 2005).

Most EDs are synthetic compounds, some of

which were designed to act as estrogens (e.g.,

oral contraceptives), whereas many were

designed for other purposes and accidentally

possessed estrogenic activity, such as plasticizers (Krishnan et al. 1993; Soto et al.

1991). Naturally occurring xenoestrogens in

the environment include phytoestrogens

produced by plants reported to have

pleiotropic effects, including antioxidative

and apoptotic activity, inhibitors of kinases,

and suggested anticancer actions on prostate

and breast carcinomas (Basly and Lavier

2005; Mueller 2002; Sirtori et al. 2005).

Other EDs such as the pesticides vinclozolin,

procymidone, and p,p ́-dichlorodiphenyldichloroethene (DDE) possess in vitro and

in vivo androgenic actions and can affect

male reproduction in animals (Gray 1998).

Bisphenol A (BPA) and BPA dimethacrylate (BPA-DM) are monomers used

largely in polycarbonate plastic and polystyrene resins and as dental sealants.

Halogenated derivatives of BPA, such as

tetrabromobisphenol A (TBBPA), are widely

used as flame-retardants for building material, paints, plastic products including epoxy

resin, electronic circuit boards, and other

electronic equipments. Depolymerization of

these products results in BPA and its derivatives, which leach into foods (Brotons et al.

1995), into infant formula from plastic bottles (Biles et al. 1999), into saliva of patients

treated with dental sealants (Olea et al.

1996; Pulgar et al. 2000), and in fresh food

at the microgram to milligram per kilogram

level (Vivacqua et al. 2003). BPA and

TBBPA have been detected in the concentration range of 0.1–10 ppb in human blood,

urine, and fetal tissues, and related BPA levels

in blood and fat tissues have also been

reported (Ikezuki et al. 2002; Schonfelder

et al. 2002; Thomsen et al. 2001; vom Saal

and Hughes 2005).

Alkylphenol ethoxylates (APEs) are

widely used surfactants and detergents in

domestic and industrial products and are

commonly found in wastewater. In sewage

treatment plant effluents, APEs are degraded

to the more resistant alkylphenols such as

4-n-nonylphenol (nNP) and 4-n-octylphenol

(nOP). Data from studies across many

regions of the world have shown significant

levels in samples of every environmental

compartment examined, including fish muscle tissue (Ying et al. 2002), and they are generally ubiquitous in food as well (Guenther

et al. 2002).

The sex steroid receptors such as the

estrogen receptors (ER) α and ERβ and the

androgen receptor (AR) belong to the

nuclear receptor family and are liganddependent transcription factors (Bjornstrom

and Sjoberg 2005; Schwabe and Teichmann

2004; Verrijdt et al. 2003). The genomicmediated pathway of EDs via ER and AR

includes steps such as binding of ligand to

receptor, translocation into nucleus, and

binding of the receptor–ligand complex to a

specific DNA response element causing gene

expression. 

Several years ago BPA, BPA-DM, nNP,

and nOP were reported to elicit estrogenic

activity (Andersen et al. 1999; Krishnan et al.

1993; Soto et al. 1991; Steinmetz et al. 1998;

White et al. 1994). Since then, numerous

studies have been carried out to assess the

endocrine disruption potentials of these

industrial compounds in vitro and in vivo

(Alonso-Magdalena et al. 2006; Choi and

This article is part of the monograph “Endocrine

Disruptors—Exposure Assessment, Novel End

Points, and Low-Dose and Mixture Effects.”

Address correspondence to E.C. BonefeldJorgensen, Unit of Cellular and Molecular

Toxicology, Institute of Public Health, Department

of Environmental and Occupational Medicine,

Vennelyst Blvd 6, Building 1260, University of

Aarhus, 8000 Aarhus C, Denmark. Telephone:

(+45) 89426162. Fax: (+45) 89426199. E-mail:

ebj@mil.au.dk

We thank B.S. Andersen, B.M. Plesning, and

A. Keblovski for excellent technical assistance. 

The study was supported by the European Union

project ENDOMET, QLRT-2001-02637, and the

Danish Research Council grant 2107-04-0006.

The authors declare they have no competing

financial interests.

Received 22 May 2006; accepted 26 September

2006.

Endocrine-Disrupting Potential of Bisphenol A, Bisphenol A Dimethacrylate,

4-n-Nonylphenol, and 4-n-Octylphenol in Vitro: New Data and a Brief Review

Eva C. Bonefeld-Jørgensen,1 Manhai Long,1 Marlene V. Hofmeister,1 and Anne Marie Vinggaard2

1Unit of Cellular and Molecular Toxicology, Institute of Public Health, Department of Environmental and Occupational Medicine,

University of Aarhus, Aarhus, Denmark; 2Department of Toxicology and Risk Assessment, Danish Institute for Food and Veterinary

Research, Søborg, Denmark

BACKGROUND: An array of environmental compounds is known to possess endocrine disruption

(ED) potentials. Bisphenol A (BPA) and bisphenol A dimethacrylate (BPA-DM) are monomers used

to a high extent in the plastic industry and as dental sealants. Alkylphenols such as 4-n-nonylphenol

(nNP) and 4-n-octylphenol (nOP) are widely used as surfactants.

OBJECTIVES: We investigated the effect in vitro of these four compounds on four key cell mechanisms including transactivation of a) the human estrogen receptor (ER), b) the human androgen

receptor (AR), c) the aryl hydrocarbon receptor (AhR), and d) aromatase activity.

RESULTS: All four compounds inhibited aromatase activity and were agonists and antagonists of ER

and AR, respectively. nNP increased AhR activity concentration-dependently and further increased

the 2,3,7,8-tetrachlorodibenzo-p-dioxin AhR action. nOP caused dual responses with a weak

increased and a decreased AhR activity at lower (10–8 M) and higher concentrations (10–5–10–4 M),

respectively. AhR activity was inhibited with BPA (10–5–10–4 M) and weakly increased with BPADM (10–5 M), respectively. nNP showed the highest relative potency (REP) compared with the

respective controls in the ER, AhR, and aromatase assays, whereas similar REP was observed for the

four chemicals in the AR assay.

CONCLUSION: Our in vitro data clearly indicate that the four industrial compounds have ED potentials and that the effects can be mediated via several cellular pathways, including the two sex steroid

hormone receptors (ER and AR), aromatase activity converting testosterone to estrogen, and AhR;

AhR is involved in syntheses of steroids and metabolism of steroids and xenobiotic compounds.

KEYWORDS: androgenic, aromatase, BPA, BPA-DM, endocrine disruption, estrogenic, nNP, nOP,

nuclear receptors. Environ Health Perspect 115(suppl 1):69–76 (2007). doi:10.1289/ehp.9368

available via http://dx.doi.org/ [Online 8 June 2007]
Jeung 2003; Ghisari and Bonefeld-Jorgensen

2005; Gutendorf and Westendorf 2001;

Kazeto et al. 2004; Mosconi et al. 2002;

Olsen et al. 2003; Rivas et al. 2002; Safe

et al. 2002; Sonnenschein and Soto 1998;

vom Saal and Hughes 2005; Williams et al.

2001). Many studies have focused on the

estrogenic activities of the compounds

in vitro by their potential to affect cell proliferation (E-SCREEN) or ER transactivation

in human or yeast cells (e.g., Andersen et al.

1999; Legler et al. 2002; Van den Belt et al.

2004; Vivacqua et al. 2003; Wilson et al.

2004) or the binding capacity to steroid

receptors (Scippo et al. 2004). In animals,

the rodent uterotrophic bioassays have verified the estrogenic effects of BPA and nNP

(Owens and Koeter 2003), and developmental studies have revealed toxic effects of BPADM on the reproductive system in mice

(Darmani and Al-Hiyasat 2004) and of nOP

in sows (Bogh et al. 2001).

Antagonistic effects on AR in vitro of

BPA, nNP, and nOP have previously been

reported (Lee et al. 2003; Paris et al. 2002;

Roy et al. 2004; Sultan et al. 2001; Xu et al.

2005). Neonatal exposure to BPA and nOP

affected development of the male reproductive system (Nagel et al. 1999) and plasma

testosterone in infant rats (Williams et al.

2001), whereas BPA was reported to have no

antiandrogenic effects on adult rats in the

Hersberger assay (Nishino et al. 2006).

An androgen:estrogen balance disturbed

by estrogenic compounds was suggested to

influence premature activation of spermatogenesis in humans (Kula et al. 1996), being

consistent with the ability of BPA and nOP

to advance the onset of pubertal spermatogenesis in rats (Atanassova et al. 2000). The

androgen:estrogen ratio is among other

things determined by aromatase (CYP19)

activity that is responsible for the irreversible

estrogen biosynthesis from androgens (Jones

et al. 2006; Seralini and Moslemi 2001;

Simpson et al. 2002). Depressed ovarian aromatase activity in the red mullet was suggested

to be caused by nNP and nOP (MartinSkilton et al. 2006), whereas increased CYP19

gene expression was reported in nNP-exposed

zebrafish (Kazeto et al. 2004), and nNP- or

BPA-exposed medaka fish liver (Min et al.

2003). In rats, a decreased serum 17β-estradiol

(E2) and aromatase mRNA level in Leydig

cells was interpreted to play a role in inhibited

testicular steroidogenesis by BPA (Akingbemi

et al. 2004). Interestingly, although no effect

of BPA was observed on CYP19 mRNA levels

in human placental JEG-3 cells, a time- and

concentration-dependent modulation of the

aromatase activity was reported suggesting an

interaction between the enzyme and BPA

(Nativelle-Serpentini et al. 2003). In summary, effects on aromatase activity caused by

the alkylphenols and BPA have been reported

in fish, rodent, and human cell studies.

Many EDs elicit multiple mechanisms of

action; and apart from their cell and tissuespecific ER and AR agonist or antagonist

activities, the involvement of other receptors

such as the aryl hydrocarbon receptor (AhR)

must be considered as well (Safe et al. 2002).

The AhR is a transcription factor that mediates the effects of polyaromatic hydrocarbons,

dioxins such as 2,3,7,8-tetrachlorodibenzo-pdioxin (TCDD), polychlorinated biphenyls

(PCBs) and other EDs such as certain pesticides (Fujii-Kuriyama and Mimura 2005;

Long et al. 2003, 2006). As a heterodimer

with AhR-nuclear-translocator (Arnt), AhR

regulates the inducible expression of CYP1

and other CYP families and plays a crucial

role in xenobiotic metabolism, teratogenesis

(Thomae et al. 2006) and immune suppression (Novosad et al. 2002). In addition, studies with AhR-null female mice have shown

that AhR plays a key role in female reproduction by activating ovarian CYP19 gene transcription (Baba et al. 2005). The potential of

BPA and alkylphenols to affect the role of

AhR is demonstrated by the following studies:

Low-dose in utero exposure of mice embryos

showed increased AhR mRNA expression in

brain, testes, and ovaries (Nishizawa et al.

2005b). BPA also up-regulated the mRNA

level of the AhR repressor (AhRR) and Arnt

in mid- and late-stage mice embryos, disrupting the expression of AhR and related factors

and xenobiotic metabolizing enzymes

(Nishizawa et al. 2005a). In mice Hepa-1c1c7

cells, nNP suppressed CYP1A1 expression by

antagonizing the dioxin-responsive element

(DRE) binding of nuclear AhR (Jeong et al.

2001). In parallel with E2, an estrogenic effect

of nNP was observed in marine fish, Gobious

niger, and the P4501A1 inhibition by nNP

was mediated through activation of the AhRR

(Maradonna et al. 2004). Also, in Atlantic

salmon nNP was suggested to have an impact

on the metabolism of endogenous and exogenous substrates by modulation of hepatic

CYP1A1 via AhR (Meucci and Arukwe

2006). In summary, BPA and nNP are both

able to affect AhR action in cell cultures, in

murine fetuses, and/or in fish.

The aim of the present study was to

analyze the relative ED potencies of BPA,

BPA-DM, nNP, and nOP (Figure 1) in four

key in vitro bioassays for ED effects including ER, AR, and AhR transactivation and

aromatase activity. 

Materials and Methods

Materials. BPA was obtained from SigmaAldrich Co. (Birmingham, UK). BPA-DM

was purchased from Aldrich Chemical Co.

(St. Louis, MO, USA). nNP and nOP were

purchased from Lancaster Synthesis Ltd.

(Birmingham, UK) and Aldrich Chemical

Co., respectively. The four chemicals (purity

≥ 98%), and TCDD, 98% (Cambridge

Isotopes Laboratories Inc., Andover, MA,

USA) were dissolved in dimethyl sulfoxide

(DMSO) (BDH Laboratory Supplies, Pool,

UK). The E2 (Sigma, St. Louis, MO, USA)

was dissolved in 96% ethanol (extra pure;

Merck, Darmstadt Germany). Luciferin

and fluorescamine were from Amersham

Pharmacia Biotech (Piscataway, NJ, USA)

and Molecular Probes (Eugene, OR, USA),

Bonefeld-Jørgensen et al.

70 VOLUME 115 | SUPPLEMENT 1 | December 2007 • Environmental Health Perspectives

BPA

nOP

nNP

BPA-DM

4-Androsten-3,17-dione

(androstenedione)

4-Androsten-4-ol-3,17-dione

(4-hydroxyandrostenedione)

E2

R1881 (methyltrienolone)

TCDD

Figure 1. Structures of the chemicals used in the present study. 
respectively, and bovine serum albumin

from Promega (Madison WI, USA); [1β- 3H]Androst-4-ene-3,17-dione was from

PerkinElmer (Boston, MA, USA), and both

4-androstene-3,17-dione (4-AD) and 4-

androsten-4-ol-3,17-dione (4-AOD) were

from Sigma Aldrich (Milwaukee, WI, USA).

Methyltrienolone (R1881) was purchased

from Mikromol Gmbh (Luckenwalde,

Germany). The structures of the test chemicals are shown in Figure 1.

ER-activated luciferase expression assay.

We used the stable transfected MVLN cell

line, derived from the human breast adenocarcinoma MCF-7 cell line (Pons et al.

1990), to assess effects on ER–luciferase

transactivation as described by BonefeldJorgensen et al. (2005). The luciferase data

were corrected to cell protein, and the results

are given as relative light unit (RLU)/microgram protein. Each compound was tested in

triplicate in at least three independent assays

alone and as co-treatment with 25 μM E2

[40% of the effect concentration (EC40) of

E2]. An E2 dose–response control (0.05–500

pM E2) was performed in parallel each analysis day as described (Bonefeld-Jorgensen et al.

2005). The maximal (EC100) and half maximal (EC50) effective concentrations of E2

were 150 pM and 33 pM, respectively. 

Aromatase activity. We performed the

assay protocol as described by Drenth et al.

(1998), with minor modifications. Human

JEG-3 choriocarcinoma cells (no. HTB-36;

ATCC, Manassas, VA, USA) were maintained in minimum essential medium

(Invitrogen, Life Technologies, Glasgow, UK)

supplemented with 10% fetal calf serum,

2 mM L-glutamine, 64 mg/L garamycin, and

1 mM sodium pyruvate (Gibco BRL, Life

Technology, Gaithersburg, MD, USA). Cells

were seeded in 24-well culture plates (Nunc) at

4 × 104 cells/well in 1 mL culture medium for

2 days (∼ 50% confluence). Serum-free

medium ± test compound and/or control were

added (maximum 0.1% DMSO) and incubated (37°C, 5% CO2/95% O2) for 18 hr

(∼ 90% confluence), and then medium was

removed, cells were washed with phosphatebuffered saline (PBS), and 0.5 mL serum-free

medium containing 0.2 μCi [1β3H]androst4-ene-3,17-dione and 10 nM unlabeled 4-AD,

corresponding to the KM value of the enzyme,

was added. After 2 hr of incubation, the aromatization was terminated by placing the 24-

well plates on ice. We extracted 200 μL of the

culture medium with CHCL3, and treated

100 μL of the aqueous phase with 100 μL dextran-charcoal in PBS (5%) (Sigma Aldrich).

Finally, we mixed 150 μL of the water phase

with 4 mL Hionic Fluor (Packard BioScience,

Groningen, the Netherlands) in a 6-mL vial

for scintillation (Packard BioScience) and

assayed for radioactivity (Packard Liquid

Scintillation Analyzer, model Tri-carb 2200;

Packard Instrument, Meriden, CT, USA).

The determined aromatase activity was subtracted from background level (data from

wells with medium only), corrected to cell

protein concentration, and related to the solvent control (0.1% DMSO). The protein

was determined on cell lyses with 0.5 mL

0.1 M NaOH by the modified Lowry protein

assay reagent (Pierce, Rockford, IL, USA)

according to the manufacturer’s manual. In

parallel, the 4-AOD was used as an aromatase

inhibitor control at 10 μM (EC100) and

6 nM (EC50). Each compound was tested at

10–9, 10–8, 10–7, 10–6, 10–5, and 10–4 M in

triplicate in at least three independent assays.

In each assay, all data were related to the positive control (substrate; 4-AD), which were

set to 100%. 

AhR-CALUX. The stable transfected

mouse hepatoma cell line Hepa1.12cR carrying the pGudLuc1.1 AhR-luciferase reporter

gene was kindly provided by M.S. Denison

(University of California, Davis, CA, USA).

The AhR transactivation of luciferase (AhRCALUX) was carried out as described by

Long et al. (2003, 2006). The data are given

as RLU/microgram cell protein. Each compound was tested in at least three independent assays in quadruplicate. The mean of the

solvent control (0.1% DMSO) (basal activity) was set to 1, to which the activity of test

compounds was related. Results are reported

as mean ± SD. 

Cell cytotoxicity. Cell cytotoxicity tests for

the ER, AhR-CALUX, and aromatase assays

were performed according to the CellTiter 96

Non-Radioactive Cell Proliferation assay

from Promega (Madison, WI, USA) and

cytotoxicity detection kit (LDH) from Roche

(Mannheim, Germany) as described by

Bonefeld-Jorgensen et al. (2005) and Long

et al. (2003).

Statistical analyses for aromatase-, ER-,

and AhR-transactivation analyses. The statistical analysis was performed in SPSS 10.0

(SPSS Inc., Chicago, IL). Because of inequality of variance and relatively few data points

per concentration, nonparametric statistics

was performed. We used the Kruskal-Wallis

test to compare differences between different

concentrations, and the Jonckheere-Tepstra

test to analyze for a linear trend between concentrations and response. If one or both tests

showed a significant difference (p ≤ 0.05), we

used the Mann-Whitney test to compare the

difference between each test concentration

and the respective control.

We performed dose–response analysis in

Sigma Plot 8.0 (SPSS) by fitting the curves

to the four (ER) and three (aromatase, AhR)

“parameter sigmoid curve” fit, and calculated

EC100, EC50, IC100, or IC50 for each test

compound.

AR-reporter gene assay. We tested AR

transactivation in a luciferase reporter assay.

Chinese hamster ovary (CHO) cells were

transiently transfected with the expression

vector pSVAR0 (human AR) and the

MMTV-LUC reporter plasmid as described

by Vinggaard et al. (2002), except that most

pipetting procedures were performed using a

Biomek2000 laboratory robot (Beckman

Coulter, Fullerton, CA, USA) and that 0.1

nM R1881 was used as the AR agonist. Each

test compound was tested at concentrations

of 0.15, 0.3, 1.3, 2.5, 5, 10, 20, and 40 ×

10–6 M in three independent assays as triplicates, and within each assay all data were

related to 0.1 nM R1881, which was set to

100%. Using the SigmaStat program, we

performed analysis of variance, and if data

were statistically significant followed with a

Dunnett’s test (p < 0.05). Dose–response

analysis was performed in Sigma Plot fitting

the curves to the “four parameter logistic

curve” fit, and IC50 (half maximal inhibitory

concentration) was calculated for each test

compound. We determined cytotoxicity in

parallel by transfecting cells with a plasmid

(pSVAR13) encoding for a constitutively

active AR, which lacks the ligand-binding

domain (Vinggaard et al. 2002). These

experiments were designed exactly as was the

AR reporter gene assay except that the ratio

between pSVAR13 and MMTV-LUC was

2:100.

Results

Effects on ER transactivity. BPA and BPA-DM

elicited ER-mediated dose–response luciferase

activity in the concentration range of 10–8 to 5

× 10–5 M, with the maximum response being

approximately 75% of the natural ligand E2-

induced maximum response. The EC50 values

were 3.9 μM for BPA and 4.8 μM for BPADM. (Figure 2, Table 1). Co-treatment of

MVLN cells with 25 pM E2 (EC40) and each

of the two compounds showed at 5 μM and 10

μM a further increased estrogenic effect above

the E2 EC40 control, whereas the significant

decreased response at 50 μM might be the

beginning of cell cytotoxicity (Figure 2B).

Although lower maximum induction was

observed, the alkylphenols nNP and nOP

alone also elicited a clearly significant estrogenic response in the concentration range of

10–8 M to 10–5 M and 10–6 M to 10–5 M,

respectively, with EC50 values of 8.9 μM and

4.9 μM, respectively (Figure 2, Table 1). We

observed further increase of the 25 pM E2-

induced activity level for nNP and nOP at 5 ×

10–6 and 5 × 10–6–10–5 M, respectively

(Figure 2B); cytotoxicity was observed at 25

and 50 μM, respectively (Table 1).

Effects on AR transactivity. BPA, BPADM, nNP, and nOP elicited a significantly

antiandrogenic effect on the R1881-induced

Endocrine-disruption potential of BPA, BPA-DM, nNP, and nOP

Environmental Health Perspectives • VOLUME 115 | SUPPLEMENT 1 | December 2007 71
AR activity in the range of 0.60–20 μM, with

IC50 values of 1.0, 2.3, 14.1, and 1.1 μM,

respectively, and a maximum inhibition (MI)

of 90, 89, 56, and 92%, respectively. We

determined cytotoxicity for BPA, BPA-DM,

and nNP at concentrations > 40 μM and for

nOP at concentrations > 20 μM (Figure 3

and Table 1). 

Aromatase activity. The inhibitor control

4-AOD caused an aromatase inhibition in the

concentration range of 10–9 M to 10–3 M, with

an IC50 of 6 × 10–9 M and 100% inhibition

at 10–4 M (Table 1). BPA and BPA-DM

decreased the aromatase activity significantly

at 10–4 M, with MI of 59% and 40%, respectively. nNP and nOP decreased the aromatase activity in the range 10–9–10–5 M and

10–7–10–5 M, respectively, with MI at 10–5 M

of 71% and 47%, respectively. Cytotoxicity

was observed at 10–4 M for both nNP and

nOP (Table 1).

Effects on AhR transactivation. Effects on

AhR transactivation data are shown in Table

1 and Table 2. Decreased AhR activity was

observed for BPA alone and on co-treatment

with TCDD-EC50 at 5 × 10–5 to 10–4 M and

10–4 M, respectively. BPA-DM alone weakly

but significantly increased the CALUX activity at 10–5 M. Although weak compared with

TCDD, nNP alone elicited significant

increased (∼ 4-fold) dose–response AhRCALUX activity in the range of 5 × 10–8 M

to 10–4 M, with an EC50 of 24 μM.

Moreover, nNP further increased and

decreased the TCDD-induced AhR activity

at 2.5–5 × 10–5 M and 10–4 M, respectively.

For nOP alone, weak increased AhR-CALUX

activity was observed at 10–8–2.5 × 10–8 M,

whereas a decrease was observed at 5 × 10–5

M and 10–4 M. At 10–4 M, nOP decreased

TCDD-induced activity by 60%. None of

the compounds elicited any toxicity in

Hepa1.1 2cR cells at the tested concentration

range. Rather, each of the compounds caused

a significantly increased cell proliferation at

10–4 M. However, we cannot exclude that

the decreased CALUX activity observed at

10–4 M for all compounds, except nNP

alone, might be a response due to beginning

cytotoxicity.

Discussion

In the present study we demonstrated that

BPA, BPA-DM, nNP, and nOP elicited an

impact on most of the selected end points:

ER, AR, AhR, and aromatase activity.

Estrogenicity was observed for the four compounds where BPA and nNP showed the

highest relative potency compared with E2,

supporting previously reported in vitro and

in vivo data. The four chemicals also antagonized AR transactivation in a concentrationdependent manner, with IC50 values in the

order BPA ≤ nOP ≤ BPA-DM < nNP. In

addition, all four chemicals inhibited the aromatase activity in JEG-3 cells. Furthermore,

nNP activated AhR in a concentrationdependent manner; BPA-DM activated AhR

at 10–5 M only, whereas BPA inhibited the

AhR action at 5 × 10–5 to 10–4 M, and nOP

elicited a weakly induced and a decreased

AhR activity at lower and higher concentrations, respectively. Thus our in vitro data

indicate that the four industrial chemicals

have the potential to affect several cellular

pathway systems, including gene expressions

regulated via the steroid receptors ER and

AR, the conversion of testosterone into estrogen by aromatase, and the function of AhR,

involved in syntheses of steroids such as estrogens and metabolism of steroids and xenobiotic compounds. 

MVLN cells, derived from MCF-7 cells

(Pons et al. 1990), express both ERα and

ERβ (Gaido et al. 1998; Grunfeld and

Bonefeld-Jorgensen 2004; Hofmeister and

Bonefeld-Jorgensen 2004) that can bind the

ER response element vitellogenin, vit-tk, in

Bonefeld-Jørgensen et al.

72 VOLUME 115 | SUPPLEMENT 1 | December 2007 • Environmental Health Perspectives

Figure 2. Dose–response ER transactivation of E2 and the four test chemicals. (A) The MVLN cells were

exposed to E2 in the concentration range of 0.05–500 pM and to the test chemicals at 10–8–10–4 M for 24 hr.

Solvent control was set to 100%. E2 EC100 and EC50 as well as EC50 for each of the test chemicals were

determined by Sigma Plot 8.0. (B) Agonistic and antagonistic ER activity of test chemicals BPA, BPA-DM,

nNP, and nOP. The chemicals were tested alone or on co-exposure with 25 pM E2, which was set to 1.

Mean values are shown (n ≥ 3). 

*Significantly different from the respective solvent controls (cells + 0.1% DMSO; 25 pM E2 + 0.1% DMSO). 

1.6

1.4

1.2

1.0

0.8

0.6

0.4

0.2

0

1.6

1.4

1.2

1.0

0.8

0.6

0.4

0.2

0

400

350

300

250

200

150

100

50

10–13

Concentration (M)

RLU/protein; solvent = 100%

10–12 10–11 10–10 10–9 10–8 10–7 10–6 10–5 10–4

EC50

BPA; 3.9 μM

BPA-DM; 4.8 μM

nNP; 8.9 μM

nOP; 4.9 μM

MVLN cells; EC100 = 150 pM

EC50 = 33 pM

A

B

1× 10–8 0.1% 1× 10–7 1× 10–6 5× 10–6 1× 10–5 5× 10–5

DMSO (M)

RLU/μg protein (25 pM E2 = 1)

*

*

*

*

*

*

BPA

BPA + 25 pM E2

1× 10–8 0.1% 1× 10–7 1× 10–6 5× 10–6 1× 10–5

RLU/μg protein (25 pM E2 = 1)

*

*

*

nNP

nNP + 25 pM E2

BPA nNP

* * * *

1× 10–8 0.1% 1× 10–7 1× 10–6 5× 10–6 1× 10–5 5× 10–5

RLU/μg protein (25 pM E2 = 1)

*

*

*

*

*

BPA-DM

BPA-DM + 25 pM E2

1× 10–8 0.1% 1× 10–7 1× 10–6 5× 10–6 1× 10–5

RLU/μg protein (25 pM E2 = 1)

*

* *

nOP

nOP + 25 pM E2

BPA-DM nOP

1.6

1.4

1.2

1.0

0.8

0.6

0.4

0.2

0

1.6

1.4

1.2

1.0

0.8

0.6

0.4

0.2

0

*

DMSO (M)

DMSO (M)

DMSO (M)
front of the luciferase gene (Gruber et al.

2004). Thus the estrogenicity demonstrated

for the four tested compounds in our analyses

is mediated by transactivation of ERα and/or

ERβ. Moreover, based on previous reports

including E-SCREEN, ER transactivation,

binding assays to the two ER subtypes, and

effects observed with the ER antagonists ICI

182,780 and tamoxifen (Gutendorf and

Westendorf 2001; Scippo et al. 2004;

Vivacqua et al. 2003; Wilson et al. 2004), we

can conclude that the estrogenic potential of

these four compounds is accomplished

through the ERs. However, we cannot

exclude that other cellular mechanisms are

involved, such as activation of the transcription factor cAMP-responsive element binding

protein via binding to nonclassical membrane

ERs (Quesada et al. 2002). In addition, BPA

and nNP effects were exerted predominantly

by the parent compounds and not by their

metabolites (Legler et al. 2002). 

We observed an antagonistic effect on

R1881-induced AR transactivition of the four

tested chemicals. To our knowledge, the AR

antagonism of BPA-DM has not previously

been reported. As we performed the present

study, another study using transiently transfected [human AR (hAR) and MMTV-CAT]

Africa monkey kidney CV-1 cells reported an

antagonizing effect on AR by BPA, nNP, and

nOP with similar IC50 values, except that

nOP had an IC50 approximately 100 times

higher (97.1 μM) (Xu et al. 2005) than ours.

In contrast, in a recently established stably

transfected CHO-AR-LUC cell line, only

BPA but not nNP and nOP elicited anti-AR

effects (Roy et al. 2004); similarly, BPA but

not nNP antagonized AR action in stable

transfected (hAR and MMTV-luciferase) PC3 prostate cells (Paris et al. 2002; Sultan et al.

2001). Our AR antagonistic data for BPA and

nNP are consistent with results obtained in

yeast (Lee et al. 2003) and transient transfected (hAR and ARE2-luciferase) NIH3T3

cells (Kitamura et al. 2005). In contrast, BPA

showed no effect on AR action in transient

transfected (hAR and MMTV-luciferase)

human hepatoma HepG2 cells (Gaido et al.

2000). Whole cell binding assays showed that

BPA has the potential of binding to the AR

(Paris et al. 2002). We suggest that the antiAR action observed in the present study for

BPA, BPA-DM, nNP, and nOP indicates that

the chemicals have the ability to bind the AR

and thus compete with endogenous androgens

for binding and regulating AR-dependent gene

expression. However, indirect interference

with other transcriptional factors involved in

AR transactivation may also play a role. The

discrepancy between the reported AR effects of

the compounds might be caused by different

sensitivities of the various assays, differences in

co-factors in the different cell lines, and/or the

use of different AR response elements

(Sommer and Haendler 2003; Verrijdt et al.

2003). However, it is important to clarify the

cause(s) of the discrepancies in the use of ARreporter analyses to assess the potential actions

of chemicals on androgenic processes in vivo.

In the synthesis of steroid hormones from

cholesterol, the aromatase enzyme is pivotal

by its irreversible conversion of androgens to

estrogens (Simpson et al. 2002). Aromatase,

which is located in the endoplasmic reticulum membrane, is expressed in several tissues

and cell types in humans, including adipose

tissue, various sites of the brain, and testicular

Leydig cells in males (Jones et al. 2006).

Aromatase activity and endogenous estrogens

(and ERs) are important in male reproduction (Carreau et al. 2006). In rat Leydig cells,

BPA inhibited testicular steroidogenesis by a

decrease in 17α-hydroxylase/17-20 lyase and

CYP19 expression causing decreased testosterone and E2 synthesis (Akingbemi et al.

2004). After 18-hr exposure of JEG-3 cells to

BPA, BPA-DM, nNP, or nOP, we observed

a significant inhibition of aromatase activity.

To our knowledge, this is the first report

showing BPA-DM, nNP, and nOP effects on

aromatase activity in mammalian cells. nNP

and, to lesser extent nOP, caused a dosedependent inhibition of aromatase activity in

the range of 10–9–10–5M. Our data are supported by an earlier report examining BPA in

JEG-3 cells, in which a significant effect was

observed on aromatase activity but not on

gene expression (Nativelle-Serpentini et al.

2003). Given the observation that short

(2 hr) exposure increased and long (18 hr)

exposure decreased aromatase activity, and

that parallel transfections with CYP19 cDNA

elicited similar effects, Nativelle-Serpentini

et al. (2003) concluded that the xenobiotics

acted at the aromatase protein level. In summary, the tested plasticizers seem to have the

potential to affect the aromatase activity and

thus the synthesis of estrogens in mammals as

well as in fish (Kazeto et al. 2004; MartinSkilton et al. 2006; Min et al. 2003).

The present study suggests that BPA,

BPA-DM, nNP, and nOP have the potential

for indirect AhR-mediated actions on xenobiotic metabolism, steroid synthesis, and

metabolism. This is the first report, to our

knowledge, showing the ability of BPA-DM

Endocrine-disruption potential of BPA, BPA-DM, nNP, and nOP

Environmental Health Perspectives • VOLUME 115 | SUPPLEMENT 1 | December 2007 73

Table 1. ER, AR, AhR, and aromatase characteristics of the four test chemicals alone. 

Assay LOEC MOEC Maximum %a REP EC50b/IC50c Cytotox (M)

ER

E2 1 × 10–11 1.5 × 10–10 — 1 3.3 × 10–11b —

BPA 1 × 10–7 1 × 10–5 — 1 × 10–4 3.9 × 10–6b —

BPA-DM 5 × 10–6 1 × 10–5 — 2 × 10–6 4.8 × 10–6b —

nNP 1 × 10–8 1 × 10–5 — 1 × 10–3 8.9 × 10–6b 2.5 × 10–5

nOP 5 × 10–6 1 × 10–5 — 2 × 10–6 4.9 × 10–6b 5.0 × 10–5

ARd

BPA 0.6 × 10–6 2 × 10–5 90 ND 1.0 × 10–6c > 4 × 10–5

BPA-DM 2.5 × 10–6 2 × 10–5 89 ND 2.3 × 10–6c > 4 × 10–5

nNP 2.5 × 10–6 2 × 10–5 56 ND 1.4 × 10–5c > 4 × 10–5

nOP 0.6 × 10–6 1 × 10–5 92 ND 1.1 × 10–6c > 2 × 10–5

Aromatased

4-AOD 1 × 10–9 1 × 10–4 100 1 6 × 10–9c —

BPA 1 × 10–4 1 × 10–4 59 1 × 10–5 — — 

BPA DM 1 × 10–4 1 × 10–4 40 1 × 10–5 — —

nNP 1 × 10–9 1 × 10–5 71 1 — 1 × 10–4

nOP 1 × 10–7 1 × 10–5 47 1 × 10–2 — 1 × 10–4

AhR

TCDD 2 × 10–12 1 × 10–8 — 1 6.4 × 10–11b —

BPAd 5 × 10–5 d 1 × 10–4d 54 — ND > 10–4

BPA-DM 1 × 10–5 1 × 10–5 — 1 × 10–7 ND > 10–4

nNP 5 × 10–8 1 × 10–4 — 4 × 10–5 2.4 × 10–5 > 10–4

nOP 1 × 10–8 2.5 × 10–8 — 1 × 10–4 ND > 10–4

nOPd 5 × 10–5 d 1 × 10–4d 46 — ND > 10–4

Abbreviations: —, no data; LOEC, lowest effect concentration in molar (M); MOEC, maximal effect concentration in molar

(M); ND, not determined; REP, relative potency. 

aMaximum down-regulation of the control inducer, which was set to 100% [0.1 nM R1881 (AR), 10 nM 4-AD (aromatase),

60 pM TCDD (AhR)]; REP-ER = LOECE2/LOECtest chemical; REP-aromatase = LOEC4-AOD/LOECtest compound. b EC50/

c IC50: Molar

concentration which exert 50% increase/50% inhibition compared to the max response of their respective control,

respectively. d Inhibited activity. 

RLU: 0.1 nM R1881 control = 100%

Concentration (μM)

0.1 1 10 100

120

100

80

60

40

20

0

nOP

BPA

BPA-DM

nNP

Figure 3. AR antagonism of the test chemicals on

co-exposure with R1881 in transient transfected

CHO-K1 cells. Cells were transiently transfected

with the AR expression pSVAR0 and the reporter

pMMTV-LUC vector. 
Bonefeld-Jørgensen et al.

74 VOLUME 115 | SUPPLEMENT 1 | December 2007 • Environmental Health Perspectives

and nOP to affect the transactivation of AhR.

BPA and BPA-DM elicited in the 10–5 M

range weak antagonistic and agonistic AhR

effects, respectively. We observed a clear dosedependent AhR activation for nNP, although

it was weak compared with that for TCDD.

In previous in vitro studies, BPA and nNP

had no effects on their own, but antagonized

TCDD-induced CYP1A1 mRNA and

7-ethoxyresorufin-O-deethylase (EROD)

activity levels in Hepa-1c1c7 cells via interference with AhR:DRE binding and/or transport

into the nucleus and/or the action of co-transcription factors (Jeong et al. 2000, 2001).

Similarly to BPA and nNP, estradiol was

shown to antagonize AhR:TCDD-induced

CYP1A1 expression and EROD activity, and

weakly induce EROD activity via AhR not

involving ER (Jeong and Lee 1998). In vivo,

the endocrine disruption potential of BPA on

mice embryonic development was demonstrated at very low doses (0.02 μg/kg/day)

(Nishizawa et al. 2005a, 2005b). 

We observed weak increased and

decreased AhR activity at lower (10–8 M)

and higher (10–6 M) nOP concentrations,

respectively, and nOP also induced ER activity at ∼ 10–6 M. We wonder whether this

estrogenic mimic of nOP can be explained

by the fact that the bifocal effects of estradiol

on AhR can either induce or antagonize AhR

function (Jeong and Lee 1998; Kharat and

Saatcioglu 1996). That nOP elicited AhR

agonism at lower concentrations and reduced

AhR transactivation at higher concentrations

might be attributed to inhibitory cross-talk

of AhR and ER. However, the specific mechanisms involved must be further studied.

The demonstration in AhR knock-out

mice that AhR is an important factor in

female reproduction by regulating the

expression of ovarian aromatase elucidates

the physiologic role of AhR and also suggests

AhR as a mediator of endocrine disruption

(Baba et al. 2005). Antiestrogenic actions

mediated via AhR are well described, in

which AhR:ligand inhibits ER binding to

ERE (Pocar et al. 2005) and increases proteasomal degradation of ER (Wormke et al.

2003). A further perspective of AhR as a factor in ED is given by AhR’s estrogenic

potential in absence of estrogen via the agonist-activated AhR:Arnt interaction, with

ERα and/or ERβ leading to transcriptional

activation of ER-regulated genes (Ohtake

et al. 2003). In addition, antiandrogenic

actions of TCDD have pointed out two possible mechanisms: blocking of AR-induced

gene expression and AhR–AR cross-talk possibly involving competition of co-regulators

(Barnes-Ellerbe et al. 2004; Jana et al. 1999).

Hypothetically, AhR might play a role in

regulating the cell ratio of androgens:estrogens via activation of CYP19, thereby affecting male and female reproduction, affecting

estrogenic actions via ERs, inducing cell

metabolism via, for example, CYP1A, and

inhibiting AR functions.

In summary, our in vitro data demonstrated that the four tested chemicals have the

potential to affect central endocrine pathways

through their capacity to affect the function of

the nuclear receptors ER, AR, AhR, and aromatase activity. BPA-related compounds and

alkylphenols have been found in human fluids

in 0.1–10 nM (vom Saal and Hughes 2005)

and 0.4–13.9 ng/mL urine (Kawaguchi et al.

2004; Kuklenyik et al. 2003), respectively. In

this study, the lowest observed effect concentrations for the BPA phenols and the

alkylphenols were in the 100–1,000 nM and

10–1,000 nM ranges, respectively. Although

the effective concentrations in vitro for the

tested compounds were ≥ 1,000 times higher

than the level found in humans, their ability

to act via more than one mechanism might

enhance the biologic effect in the intact

organism, because the final response will likely

be determined as a sum of the interactions of

all pathways implicated. Furthermore, because

most humans are exposed to several chemicals

simultaneously and some EDs have been

shown to act additively in vitro and in vivo

(Birkhoj et al. 2004; Nellemann et al. 2003;

Rajapakse et al. 2002), potential mixture

effects should also be taken into consideration

in the risk assessment. 

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Concentration BPA BPA-DM nNP nOP

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Compounds alone

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1.0 × 10–4 M 0.46 ± 0.15* 0.97 ± 0.25 3.68 ± 0.20* 0.54 ± 0.19*

Compounds + TCDD

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5.0 × 10–5 M 0.93 ± 0.16 1.23 ± 0.21 1.41 ± 0.36* 1.10 ± 0.39

1.0 × 10–4 M 0.44 ± 0.13* 0.56 ± 0.29 0.61 ± 0.22* 0.41 ± 0.11*

Concentrations up to 10–4 M were not cytotoxic to the cells, whereas concentrations > 10–4 M elicited cytotoxicity to the

Hepa1.1 2cR cell. The results given refer to effects observed at concentrations eliciting no cytotoxicity (n ≥ 3). 

*p < 0.05, versus the respective control, which was set to 1. 
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