NOAA Technical Memorandum NOS NCCOS CCMA 149 

Endocrine Disruption in Fish

An Assessment of Recent

Research and Results

U.S. Department of Commerce

NOAA National Oceanic and Atmospheric Administration

Center for Coastal Monitoring and Assessment

National Centers for Coastal Ocean Science

National Ocean Service

June 2002
Center for Coastal Monitoring and Assessment 

National Centers for Coastal Ocean Science 

National Ocean Service 

National Oceanic and Atmospheric Administration 

U.S. Department of Commerce 

N/SCI-1, SSMC4 

1305 East/West Highway 

Silver Spring, MD 20910 

Notice 

This report has been reviewed by the National Ocean Service of the National Oceanic and Atmospheric 

Administration (NOAA) and approved for publication. Such approval does not signify that the contents of 

this report necessarily represents the official position of NOAA or of the Government of the United States, nor 

does mention of trade names or commercial products constitute endorsement or recommendation for their 

use. 

V1 
NOAA Technical Memorandum NOS NCCOS CCMA 149 

Endocrine Disruption in Fish

An Assessment of Recent Research and Results

Anthony S. Pait 

National Oceanic and Atmospheric Administration 

Judd O. Nelson 

University of Maryland 

Silver Spring, Maryland 

June 2002 

United States National Oceanic and National Ocean Service 

Department of Commerce Atmospheric Administration 

Donald L. Evans Conrad C. Lautenbacher, Jr. Margaret A. Davidson 

Secretary Vice Admiral (Ret.), Administrator Assistant Administrator (Acting) 

This report should be cited as:

Pait, A.S. and J. O. Nelson. 2002. Endocrine Disruption in Fish: An Assessment of Recent Research and

Results. NOAA Tech. Memo. NOS NCCOS CCMA 149. Silver Spring, MD: NOAA, NOS, Center for Coastal

Monitoring and Assessment 55pp.
Contents 

Abstract .........................................................................................................................................................1

Introduction.................................................................................................................................................1

Historical Perspective............................................................................................................................2

Current Effects........................................................................................................................................2

EDC Debate.............................................................................................................................................3

Endocrine System: Overview............................................................................................................. 3

Pituitary-Hypothalamus-Gonadal Axis..............................................................................................3

Mode of Action..........................................................................................................................................4

Types and Sources of EDCs................................................................................................................ 5

Laboratory Studies...................................................................................................................................9

Alkylphenol Polyethoxylates............................................................................................................... 9

Bisphenol-A...........................................................................................................................................14

4-tert-Pentylphenol.............................................................................................................................. 15

Polycyclic Aromatic Hydrocarbons...................................................................................................15

Polychlorinated Biphenyls..................................................................................................................17

Furans and Dioxins..............................................................................................................................18

Polybrominated Diphenyl Ethers......................................................................................................18

Pesticides............................................................................................................................................... 19

Phthalate Esters.................................................................................................................................... 20

Metals.....................................................................................................................................................20

β-Sitosterol.............................................................................................................................................21

Synthetic Estrogens..............................................................................................................................21

Freshwater Field Studies................................................................................................................... 22

Sewage Treatment Plants...................................................................................................................22

United Kingdom..........................................................................................................................22

Western Sweden.......................................................................................................................... 29

Germany....................................................................................................................................... 29

New South Wales, Australia...................................................................................................... 29

Minnesota/Mississippi rivers................................................................................................... 29

Lake Mead, Nevada....................................................................................................................30

Pulp and Paper Mills..........................................................................................................................30

Lake Saimaa, Finland..................................................................................................................30

St. Maurice River, Quebec..........................................................................................................30

Lake Superior, Canada............................................................................................................... 31

St. John’s River, Florida.............................................................................................................. 31

Perdido Bay, Florida....................................................................................................................31

Fenholloway River, Florida....................................................................................................... 32

Red River, North Dakota............................................................................................................32

General Surveys.................................................................................................................................. 32

Atlantic Canada...........................................................................................................................32

Escambia and Blackwater rivers, Florida................................................................................ 33

National Survey (USGS).............................................................................................................33

Mississippi River......................................................................................................................... 35

i 
Contents (continued) 

Estuarine/Marine Studies..................................................................................................................35

Sewage Treatment Plants...................................................................................................................35

United Kingdom..........................................................................................................................35

Pulp and Paper Mills..........................................................................................................................39

St. George’s Bay, Newfoundland.............................................................................................. 39

Bothnian Bay, Finland/Sweden.................................................................................................39

Miramachi Estuary, New Brunswick........................................................................................39

General Surveys.................................................................................................................................. 40

United Kingdom..........................................................................................................................40

Wadden Sea, Netherlands..........................................................................................................40

Straits of Messina, Italy...............................................................................................................41

Tokyo Bay......................................................................................................................................41

German Baltic Coast................................................................................................................... 42

Puget Sound.................................................................................................................................42

Boston Harbor and Raritan Bay.................................................................................................43

Long Island Sound and Boston Harbor................................................................................... 43

San Pedro Bay, California...........................................................................................................44

Chesapeake Bay........................................................................................................................... 44

New York Harbor........................................................................................................................ 45

Conclusions.............................................................................................................................................. 46

Literature Cited.......................................................................................................................................48

ii 
List of Figures 

1. Schematic of typical endocrine system in teleost fish....................................................................... 3

2. Pituitary-hypothalamus-gonadal axis and the action of estradiol.................................................. 4

3. Mechanism of action of the estrogen receptor.................................................................................... 5

4. Molecular structure of 17β-estradiol, along with a number of confirmed or suspected

endocrine disrupting compounds........................................................................................................ 7

5. Vitellogenin production from cultured hepatocytes of rainbow trout in response to

alkylphenolic compounds and 17β-estradiol (E2

)............................................................................ 10

6. Production of vitellogenin in male killifish Fundulus heteroclitus

exposed to 4-nonylphenol................................................................................................................... 10

7. Effect of various compounds on GSI and plasma 17β-estradiol in female

Atlantic croaker (Micropogonias undulatus)........................................................................................17

8. Estrogenic activity of selected compounds in male rainbow trout................................................18

9. Effect of several estrogenic compounds, including a mixture of all five,

on vitellogenin production in male rainbow trout...........................................................................18

10. Results of a survey to assess estrogenicity of effluents from sewage treatment

plants to rainbow trout in England and Wales................................................................................. 23

11. Gradations of ovotestes discovered in roach fish (Rutilus rutilus) in the U.K..............................26

12. Comparison of the overall incidence of intersex roach in control, upstream

and downstream sites...........................................................................................................................27

13. Plasma vitellogenin in roach from various rivers, upstream and

downstream of STP effluents...............................................................................................................28

14. Regions and sampling sites in the USGS reconnaissance study.....................................................34

15. Vitellogenin in male flounder (Platichthys flesus) from the Tyne Estuary and

 Solway Firth in the United Kingdom................................................................................................38

16. Vitellogenin in wild male flounder from five U.K. estuaries......................................................... 40

17. Relative plasma vitellogenin concentrations of mesocosm female flounder 

Platichthys flesus, in November 1992 and May 1993.........................................................................41 

iii 
List of Tables 

1. Selected endocrine glands and hormonal action in fish....................................................................... 4

2. Confirmed or suspected endocrine disrupting compounds................................................................6

3. Relative potencies of estrogenic compounds......................................................................................... 9

4. Effects of endocrine disrupting compounds on whole fish in the laboratory..................................11

5. In vitro (fish hepatocyte) effects of endocrine disrupting compounds............................................. 13

6. Field studies of endocrine disruption in freshwater species of fish..................................................24

7. Field studies of endocrine disruption in saltwater species of fish.................................................... 36

8. Vitellogenin in male Fundulus heteroclitus from Chesapeake Bay sites.............................................45

iv 
Endocrine Disruption in Fish:

An Assessment of Recent Research and Results

Abstract. This report provides an assessment of recent 

investigations into endocrine disruption in fresh and 

saltwater species of fish. Most work to date has concentrated on reproductive endocrine disruption. Laboratory 

studies have shown a variety of synthetic and natural 

chemicals including certain industrial intermediates, 

PAHs, PCBs, pesticides, dioxins, trace elements and plant 

sterols can interfere with the endocrine system in fish. The 

potency of most of these chemicals, however, is typically 

hundreds to thousands of times less than that of endogenous hormones. Evidence of environmental endocrine 

disruption ranges from the presence of female egg proteins 

in males and reduced levels of endogenous hormones in 

both males and females, to gonadal histopathologies and 

intersex (presence of ovotestes) fish. 

Overt endocrine disruption in fish does not appear to be a 

ubiquitous environmental phenomenon, but rather more 

likely to occur near sewage treatment plants, pulp and 

paper mills, and in areas of high organic chemical contamination. However, more widespread endocrine disruption can 

occur in rivers with smaller flows 

and correspondingly large or 

numerous wastewater inputs. 

Some of the most severe examples 

of endocrine disruption in fish 

have been found adjacent to 

sewage treatment plants. Effects 

are thought to be caused primarily by natural and synthetic estrogens and to a lesser 

extent by the degradation products of alkylphenol polyethoxylate surfactants. Effects found in fish near pulp and 

paper mills include reduced levels of estrogens and 

androgens as well as masculinization of females, and has 

been linked to the presence of β-sitosterol, a plant sterol. 

Effects seen in areas of heavy industrial activity typically 

include depressed levels of estrogens and androgens as well 

as reduced gonadal growth, and may be linked to the 

presence of PAHs, PCBs, and possibly dioxins. At this 

time, however, there is no clear indication that large 

populations of fish are being seriously impacted as a result 

of endocrine disruption, although additional work is 

needed to address this possibility.

INTRODUCTION 

Recently, concern has arisen that certain environmental contaminants as well as some naturally occurring 

compounds have the potential to impact the endocrine system which regulates vital life processes, 

including development, growth, metabolism and 

reproduction. Research on the identification and 

effects of endocrine-active compounds has become 

an important area of human and environmental 

health research. In this report, compounds that 

either mimic or antagonize the action of naturally 

occurring hormones are termed endocrine disrupters 

or endocrine disrupting compounds (EDCs). 

A number of definitions currently exist for EDCs. 

Kavlock et al. (1996) (see inset) proposed a definition 

that includes not only the primary ways EDCs are 

thought to interfere with normal functioning of the 

endocrine system, but 

also some of the effects 

EDCs can have on an 

organism. In 1996, the 

European Commission 

defined an endocrine 

disrupter as “an exogenous substance that 

causes adverse health 

effects in an intact 

organism, or its progeny, 

consequent to changes in endocrine function” 

(European Commission, 1996). As with Kavlock et al. 

(1996), this definition stresses that adverse health 

effects occur as a result of one or more changes in 

endocrine function. The National Research Council 

adopted hormonally active agents as a more neutral 

mechanistic descriptor, and defined them broadly as 

substances that possess hormonelike activity, regardless of structure (NRC, 1999). 

Chemicals that can mimic or antagonize the effects of 

endogenous hormones could potentially have serious 

effects not only on the development and well being 

of an individual organism, but perhaps more importantly on the ability of that organism to reproduce, 

Endocrine disrupters have been defined as 

exogenous agents that interfere with the production, release, transport, metabolism, binding, 

action or elimination of natural hormones in the 

body responsible for the maintenance of homeostasis and the regulation of developmental 

processes (Kavlock et al., 1996). 

1
Introduction 

and its offspring to survive and eventually reproduce. In the aquatic environment, the ability of an 

EDC to affect an individual or population would 

depend on a number of factors, including the potency or efficacy of the EDC, its concentration, 

duration of exposure, bioconcentration potential, 

presence of other EDCs, life stage exposed, season, 

other environmental stressors present (e.g., temperature, salinity, and other contaminants), and mobility 

of the individual. 

Many of the investigations into EDCs in the aquatic 

environment have involved fish because of similarities in the endocrine system to higher vertebrates 

(Bond, 1979), a demonstrated sensitivity to EDCs in 

both laboratory and field investigations, ease of 

working with fish, and the fact that wild populations 

in a number of 

locations have exhibited effects thought to 

be associated with 

exposure to EDCs. 

Kime (1999) has also 

pointed out that both 

sexes produce large 

numbers of gametes 

which can be readily 

counted or examined 

for malformations, 

hatch rates are readily 

determined, and the 

resultant offspring can 

be easily monitored for 

developmental abnormalities or subsequent 

reproductive problems. 

Historical Perspective. The presence and effects of 

EDCs is not a new phenomenon. As early as 1949, it 

was known that crop dusters handling 

dichlorodiphenyltrichloroethane (DDT) frequently 

had reduced sperm counts (Patlak, 1996). DDT was 

shown to produce characteristically estrogenic 

responses in the reproductive tracts of rats and birds 

(Bitman and Cecil, 1970). The tragic use of the 

synthetic estrogen diethylstilbestrol, or DES, by 

pregnant women from the late 1940s until 1971 to 

help prevent miscarriages resulted in infertility and 

increased rates of vaginal clear cell adenocarcinomas 

in daughters (Colborn et al., 1993). In herring gulls, 

studies beginning in the mid-1960s in Lake Michigan 

suggested that environmental contaminants were 

adversely and, in some cases, severely affecting 

hatching success (Keith, 1966). 

More recently, however, the number of compounds 

identified as suspected or confirmed endocrine 

disrupters has increased substantially and includes 

industrial intermediates, such as 4-nonylphenol, 

bisphenol-A, and the phthalate ester plasticizers, as 

well as classic contaminants such as the polycyclic 

aromatic hydrocarbons (PAHs), polychlorinated 

biphenyls (PCBs), dioxins, certain pesticides, and 

even a number of trace elements (Knudsen and 

Pottinger, 1999). In plants, naturally occurring 

compounds termed phytoestrogens are also known 

to have hormonelike properties (Mitksicek, 1995). 

The identification of EDCs has come through many 

avenues of investigation, including chance discoveries while working with various chemicals and cell 

lines to detailed structure/activity investigations that 

test specific compounds 

for their ability to 

produce a response in 

cell lines or whole 

animals. Responses used 

to detect endocrine active 

compounds typically 

include cell proliferation 

(i.e., in breast cancer cell 

cultures), binding to the 

estrogen receptor, and 

production of specific 

proteins in cell cultures 

or whole animals. 

Current Effects. Examples 

of apparent reproductive 

endocrine disruption in 

aquatic organisms are, in 

some areas, disturbingly 

numerous and include intersex fish (male and female 

gonadal characteristics in gonochoristic or normally 

separate sex fish) (Jobling et al. 1998), elevated levels 

of a female egg protein in male fish (Lye et al., 1998; 

Janssen et al., 1997; Folmar et al. 1996), and degeneration of gonadal tissue (Lye et al., 1998; Janseen et al., 

1997). In Lake Apopka in Florida, a dramatic decline 

in the alligator population has been attributed to a 

spill of DDT, dicofol, and possibly 1, 2-dibromo-3-

chloropropane (Guillette et al., 1996; Semenza, 1997; 

Risebrough, 1999). 

Feminization and masculinization as well as embryonic deformities in gulls have also been linked to 

PCBs and halogenated aromatic hydrocarbons (Fox, 

1993). Fortunately, it appears that some of the effects 

seen in fish-eating birds in the Great Lakes are on the 

decrease (Fox, 1993). 

Winter flounder camouflaged to match surroundings. 

Image courtesy of NOAA/NOS. 

2
Endocrine System Overview 

EDC Debate. There has been a vigorous debate 

within the scientific and regulatory communities 

regarding the extent of endocrine disruption in the 

environment, the importance of synthetic chemicals 

relative to naturally occurring EDCs, and the relevance of endocrine disrupters compared to an 

organism’s internal suite of hormones. Some argue 

that wildlife populations have already been significantly impacted by EDCs, resulting in decreased 

fertility in fish and shellfish, demasculinization and 

feminization of male fish, birds, and mammals, and 

masculinization of female fish and birds (Colborn et

al., 1997). Others, however, are doubtful that EDCs in 

most situations are likely to have a major impact. 

They point out that EDCs typically bind to receptors 

(e.g., the estrogen receptor) hundreds or thousands of

times more 

weakly than 

endogenous 

hormones. An 

argument has also 

been made, at 

least in the case of 

environmental 

estrogens, that the 

occurrence of 

natural and 

synthetic 

antiestrogens in 

the environment 

would act to 

negate the effect 

of environmental 

estrogens (Stone, 

1994). While 

there is considerable and ongoing 

debate on the 

occurrence and significance of endocrine disruption,

there is little disagreement over the need to better 

understand the impacts EDCs may be having in the 

environment.

in fish secretes a number of hormones which affect 

growth, osmoregulation, lipid metabolism and 

reproductive development and behavior, as well as 

controlling other endocrine glands (Bone et al., 1995).

The corpuscles of Stannius secrete hypocalcin which 

is thought to be involved in calcium homeostasis an

may also be involved in controlling the ratio of 

calcium to sodium and potassium in the plasma 

(Bone et al., 1995). As in humans, the fish pancreas 

secretes insulin which aids in glucose permeability. 

Secretion of glucagon, also by the pancreas, enables 

increased glycogen and lipid metabolism. 

Although most of what is currently known about the 

effects of EDCs involves reproduction and reproductive behavior, other areas of the endocrine system, 

such as the 

thyroid, may 

also be targets 

for EDCs.

Investigations

into possible

effects on other 

targets are just 

beginning. At

an even more 

basic level,

there is still 

some uncertainty as to the

structure and 

function of

some of the

naturally 

occurring 

teleost hormones and the 

glands themselves (Bone et al., 1995), making an assessment of the 

impact of endocrine disrupting compounds, other 

than on reproduction, even more difficult. 

Pituitary-Hypothalamus-Gonadal Axis 

One of the most studied pathways that can be 

affected by EDCs is the pituitary-hypothalamusgonadal axis, which will be used to illustrate how 

EDCs can affect the endocrine system in fish. The 

release of gonadotropin releasing hormone (GnRH) 

from the hypothalamus in response to a series of 

environmental cues in fish results in the production 

and release of gonadotropin hormones (GTH) from 

the pituitary gland (Figure 2). The gonadotropins 

released into the systemic circulation elicit increased

androgen and estrogen production by the gonads 

d 

Figure 1. Schematic of typical endocrine system in teleost fish. From

Bond (1979). Reprinted by permission. 

 ENDOCRINE SYSTEM: OVERVIEW 

The endocrine system in fish consists of various 

glands located throughout the body which synthesize and secrete hormones to regulate an array of 

biological processes (Figure 1 and Table 1). For 

example, the thyroid gland secretes the hormones 

thyroxine (T4) and triiodothyronine (T3), which are 

believed to aid fish in adapting to changes in temperature and to osmotic stress. The pituitary gland

3
HYPOTHALAMUS 

Blood Hepatocyte 

circulation 

PITUITARY 

Figure 2. Pituitary-hypothalamus-gonadal axis and the action of estradiol. From Drean (1994). Reprinted by permission. 

Indirect or/and 

Synthesis of direct action? 

 GnRH 

Synthesis of 

GTH

+ 

+ 

R 

Steroid 

Binding 

Protein 

E2 

ER 

E2 

+ 

Vg gene 

Vg synthesis 

and maturation 

Vg 

R 

Phosvitin R 

Lipovitellin 

Vg 

Follicular Ovocyte cell 

+/-

+ 

E2 Synthesis 

Part A = Hypothalamus-Pituitary-gonadal axis Part B = Vitellogenesis 

Table 1. Selected endocrine glands and hormonal action in fish. 

Gland/Hormone Target Organ Effect(s) 

Pituitary 

osmoregulation, reproduction, growth, Prolactin various lipid metabolism, metabolism 

Growth hormone various stimulation of growth 

Thyroid stimulating hormone thyroid stimulation of thryroxine 

Gonadotrophic hormone gonads stimulation of gonads 

constricts gill blood vessels, systemic Isotocin, mesotocin blood vessels vasodilation 

Thyroid 

adaptation to environmental changes Thyroxin many such as temperature or osmotic stresses 

Calcitonin gills and kidney regulation of calcium metabolism 

Corpuscles of Stannius 

Hypocalcin gills calcium homeostasis 

Pancreas 

Insulin all cells increases glucose permeability 

Glucagon all cells glycogen and lipic metabolism 

Chromaffin tissue 

Adrenaline circulation gill vasodilation, system vasoconstriction 

Noradrenaline circulation increase heart and glucose metabolism 

Intrarrenal 

Corticosteroids gills, kidney stress response, osmoregulation 

Gonads 

reproductive status and behavior; also of Androgens and estrogens many, including brain other fish (as pheromones) 

Adapted from Bone et al., 1995. Reprinted by permission. 

Mode of Action 

(Bone et al., 1995). The major 

estrogen in female fish, 17βestradiol (E2

), is produced 

primarily in the ovary by the 

follicular cells. In addition to 

their importance in eliciting 

reproductive behavior and the 

development and maintenance of 

secondary sex characteristics, the 

estrogens and androgens are 

involved in the production of 

gametes (Bone et al., 1995). 

In oviparous or egg-laying fish,

as in other egg-laying animals, 

the release of E2 from the ovary 

leads to the synthesis of large 

amounts of vitellogenin by the 

hepatocytes (liver cells) (Figure 

2). This high density lipoprotein,

the precursor of egg yolk, is then 

transported from the liver via the circulatory system 

and incorporated into developing oocytes (Anderson 

et al., 1996a). 

Although estrogens are typically associated with 

females and androgens with males, that demarcation 

may not be a rigid one (Sharpe, 1997). Research 

indicates that both male and female vertebrates 

produce and use estrogens and androgens. It has 

become fairly well accepted that a minor role for 

estrogens exists in males, for example, in the regulation of GTH secretion by the pituitary gland. It is 

also possible that proper levels of 

estrogens in males may have more 

widespread effects, and may even 

be essential for fertility (Sharpe, 

1997). Exposure of an organism, 

however, to levels of natural hormones or EDCs which overwhelm 

or interfere with the proper functioning of the endocrine system has 

the potential to seriously affect the 

health of an organism and its 

progeny. 

MODE OF ACTION 

Endocrine disrupting compounds 

are believed to exert their influence 

by: 1) mimicking the effects of 

endogenous hormones, such as the 

estrogens and androgens; 2) antagonizing the effects of endogenous 

hormones; 3) altering the pattern of 

synthesis and metabolism of normal 

hormones; and 4) modifying hormone receptor levels (Soto et al., 

1995). EDCs may also interfere with 

4
Types and Sources of EDCs 

the binding proteins that act to transport endogenous 

hormones to their destination. 

The most frequently studied and best understood 

type of EDCs are those that mimic estrogens 

(Gillesby and Zacharewski, 1998; Kime, 1999). A 

current model for the binding and action of E2 is 

shown in Figure 3. Estradiol produced by the 

ovaries and transported via the circulatory system is 

passively taken up by the cell (e.g., hepatocyte) and 

then crosses the nuclear membrane. The unliganded 

estrogen receptor (ER) is maintained in an inactive 

conformation through interactions with a number of 

proteins, 

primarily 

heat shock 

proteins Hsp 

59, 70, and 90

(Gillesby and

Zacharewski,

1998). 

Following 

the binding 

of estrogen 

to the 

receptor, the 

heat shock 

proteins 

dissociate 

allowing the 

ER to change 

its conformation to the 

active form. 

Once activated, the receptor forms a homodimer complex 

which seeks out specific DNA segments, in this case 

the estrogen response elements (EREs). Binding of 

the complex to the ERE results in a rearrangement of 

the chromatin and transcription of the gene, followed

by production of the target protein (Figure 3). A 

compound able to bind to the estrogen receptor in 

the cell might very well result in transcription and 

pleiotropic responses potentially affecting numerous 

functions within the organism. 

In fish hepatocytes, the protein vitellogenin is 

produced via the pathway illustrated in Figure 3. 

While vitellogenin is normally associated with 

female fish, male fish also possess the hepatocyte ER 

and can synthesize vitellogenin when exposed to E2 

or to estrogen mimics. The production of this egg 

protein in oviparous fish, particularly in males, has 

become an important biomarker in the investigation 

of EDCs, both in the laboratory and in the field. 

Normally, vitellogenin in males is either absent or at 

very low concentrations (Sumpter and Jobling, 1995; 

Panter et al., 1998; Harries et al., 1997). In females 

vitellogenin is taken up by the ovaries, in male fish 

vitellogenin produced as a result of exposure to 

estrogens or to estrogen mimics is only slowly 

metabolized, making it a valuable biomarker. Currently, however, the ecological significance of elevated levels of vitellogenin in fish, particularly in 

males, is unclear (Jobling et al., 1998). 

Although 

the mode of 

action of 

estrogenic 

compounds 

on the 

production 

of vitellogenin in male 

fish is fairly 

well understood, there 

still remains 

much 

uncertainty 

as to how 

EDCs 

impact the 

overall 

development of 

aquatic 

organisms. 

In most cases, timing of exposure during development seems to be critical. It appears there is a labile 

period when fish are most susceptible to endocrine 

pertubation occurring just after hatching or at a 

juvenile stage, the time prior to morphological sex 

differentiation (Jobling et al., 1998). Sex reversal in 

fish by treating eggs or larvae with E2

 or testosterone 

has found widespread use in aquaculture (Yamaziki, 

1983). There is also some evidence that adult gonadal organization can be affected, but these changes 

appear to be reversible once the EDC is removed.

OH 

OH 

17β-Estradiol 

OH 

OH 

ER 

Hsp90 Hsp90

Hsp70 OH 

OH 

P 

ERE 

Hsp59 

OH 

OH 

OH 

OH 

P 

OH 

OH 

OH 

POH 

AAAAA 

AAAAA 

AAAAA 

Pleiotropic 

Response 

Nucleus 

Cell 

Figure 3. Mechanism of action of the estrogen receptor. From Gillesby and 

Zacharewski (1998). Reprinted by permission. 

 TYPES AND SOURCES OF EDCS 

EDCs include a diverse group of synthetic industrial 

and agricultural chemicals and even some naturally 

occurring compounds (Santodonato, 1997). Listed in 

5
Types and Sources of EDCs 

Table 2 and diagramed in Figure 4, 

they include industrial intermediates such as bisphenol-A (Celius et 

al., 1999; Sumpter and Jobling, 

1995) and 4-nonylphenol (Jobling 

and Sumpter, 1993; White et al., 

1994), PAHs (Santodonato, 1997; 

Thomas, 1988), pesticides such as 

the insecticides endosulfan 

(Chakravorty et al., 1992), 

carbofuran (Sukumar and 

Karpagaganapathy, 1992), lindane 

(α-HCH) (Celius et al., 1999), DDT 

(Khan and Thomas, 1998; Celius et 

al., 1999), and the herbicide 

atrazine (Wetzel et al., 1994), PCBs 

(Monosson et al., 1994; Thomas, 

1988; Thomas 1989), and a number 

of metals including lead (Thomas, 

1988), cadmium (Thomas, 1989; 

Ruby et al., 2000; Kime et al., 1996), 

and mercury (Rurangwa et al., 

1998). 

Many of the chemicals currently 

known to interact with the estrogen receptor can be found in the 

effluent from sewage treatment 

plants (STPs) (Jobling et al., 1998). 

Although the efficacy or potency 

of all these compounds is less than 

that of the endogenous hormones, 

there is concern that some have 

the ability to bioaccumulate to active or harmful 

levels over time. Additional compounds will likely 

be identified as more work is completed and better 

structure/activity relationships are developed. As 

noted by Santodonato (1997), a number of estrogenic 

EDCs or their primary oxidative metabolites, share a 

common structural relationship with the phenolic A 

ring in E2 (Figure 4). 

Alkylphenols and alkylphenol polyethoxylates, or 

APEs, have received much of the recent attention 

because of their estrogenic effects in laboratory 

studies, and their presence in the aquatic environment. APEs are used as surfactants in many applica

tions from soaps and detergents to pesticide formula

tions. Once in the environment, microbial degradation results in loss of the ethoxylates, eventually 

leaving the more persistent alkylphenol (e.g., 4-

nonylphenol). Jobling et al. (1998) have shown that 

exposure to alkylphenolic compounds result in the 

synthesis of vitellogenin in male fish. 

Table 2. Confirmed or suspected endocrine disrupting compounds. 

Chemical/Class Use/Source 

Industrial Chemicals/Byproducts 

4-Nonylphenol Surfactant intermediate/breakdown product 

Octylphenol Surfactant intermediate/breakdown product 

Bisphenol-A Monomer of polycarbonate 

4-tert-pentylphenol Industrial intermediate 

Benzo[a] pyrene Fossil fuel combustion product 

Phenanthrene Fossil fuel combustion product

Polychlorinated biphenyls Transformer oil 

Dioxins Industrial and waste incineration byproducts 

Polybrominated diphenyl ethers Flame retardants 

Butyl benzyl phthalate Plasticizer 

Butylbenzyl phthalate Plasticizer 

Di-n-butyl phthalate Plasticizer

Pesticides 

Atrazine Herbicide 

Carbofuran Insecticide 

Toxaphene Insecticide 

Endosulfan Insecticide 

Lindane Insecticide

DDT Insecticide 

DDE Degradation product of DDT 

Metals 

Mercury Industry (e.g., chloralkali plants)

Cadmium Industry (e.g., metal plating, battery production)

Lead Industry (e.g., battery production)

Natural Products 

b-Sitosterol Pulp and paper industry by-product/plant sterol 

Genistein Plant sterol 

Daidzein Plant sterol 

Enterodiol Plant sterol 

APEs enter the aquatic environment via discharges 

from STPs, textile, and pulp and paper mills (White

et al. 1994; Field and Reed 1996). 4-Nonylphenol, 

one of the degradation products of the nonylphenol

polyethoxylates, has been the subject of numerous 

studies because of its estrogenic effects and presenc

in the environment. Approximately 575 million 

pounds are manufactured annually in the U.S. 

(C&EN, 1997). 

Bisphenol-A is the monomer of the plastic polycarbonate. Approximately 1.6 billion pounds are 

produced in the U.S. each year (C&EN, 1997). Much 

of the polycarbonate eventually winds up in landfills. 

Bisphenol-A appears to be an estrogen mimic, with a 

demonstrated affinity for rat ER (Krishnan et al., 

1993). 

Polycyclic aromatic hydrocarbons, or PAHs, are 

found in fossil fuels such as oil and coal and are 

e 

-

-

6
Types and Sources of EDCs 

released into the 

environment through 

combustion, surface 

runoff, oil spills, recreational boating and 

shipping, municipal 

waste effluents and 

atmospheric deposition

(Kime, 1998). In urban 

areas it is thought that 

the majority of PAHs 

are the result of atmospheric deposition from

the combustion of fossil

fuels (McElroy et al., 

1989). Santodonato 

(1997) noted that while 

PAHs may function as 

weak ER agonists, they 

are expected to bind 

preferentially to the Ah 

receptor, triggering the 

induction of Ahresponsive genes which 

can lead to an antiestrogenic effect. 

Dioxins are not intentionally manufactured 

but are typically forme

and released through 

industrial activities suc

as chlorine bleaching at

pulp and paper mills, 

chlorination at waste 

and drinking water 

treatment plants, and from municipal solid waste 

and industrial incinerator emissions. Anderson et al. 

(1996a) have shown in vitro that both the dioxin 

TCDD (2,3,7,8 tetrachlorodibenzo-p-dioxin) and the 

furan 2,3,4,7,8-pentachlorodibenzofuran are estrogen 

antagonists. As with some of the PAHs, there 

appears to be a relationship between E2

 antagonism 

and induction of the CYP1A1 protein. 

Polychlorinated biphenyls (PCBs) were manufactured and used widely as coolants and lubricants in 

transformers, capacitors, and other electrical equipment. There are no known natural sources of PCBs. 

The manufacture of PCBs in the United States ceased 

in 1977 due to their persistence, ability to 

bioaccumulate and related health concerns. PCBs 

generally appear to produce antiestrogenic and 

possibly antiandrogenic responses. 

d 

h 

C 

C 

O 

O 

CH2 

CH2 

CH2 CH2 CH3 

O 

O 

Butylbenzyl phthalate 

2,3,7,8-Tetrachlorodibenzo-p-dioxin 

Cl 

Cl 

O 

O 

Cl 

Cl 

Cl Cl 

Cl Cl 

3,3’4,4’-Tetrachlorobiphenyl 

HO OH 

Bisphenol-A 

CH3 

CH3 

C9H19 

4-Nonylphenol 

OH 

Benzo[a]pyrene 

OH 

CH3 

H3C 

CH2 

CH3 

CH3 

CH3 

CH3 

Cl 

N N 

N 

β-Sitosterol 

(CH3)2HCHN NHC2H5

Atrazine 

Cl C Cl

Cl 

Cl 

Cl 

Cl 

O 

S 

O 

O 

Endosulfan 

Cl 

CH 

o,p’-DDT 

Cl 

CCl3 

HO 

HO 

17β-Estradiol 

CH3 

HO 

HO 

17α-Ethinylestradiol 

C 

CH3 CH 

Br 

Br 

2,2’,4,4’,6-pentabromodiphenyl ether 

(PBDE-100) 

O 

Br 

Br 

Br 

Figure 4. Molecular structure of 17β-estradiol, along with a number of confirmed 

or suspected endocrine disrupting compounds. 

Phthalate esters are used in the manufacture of 

polyvinyl chloride as a softening agent and have 

been shown to be somewhat estrogenic. Compounds 

identified to date in descending order of potency 

include butylbenzyl phthalate (BBP), dibutyl phthalate, diisobutyl phthalate, and diethyl phthalate 

(Harris et al., 1997). 

Pesticides are used to control a wide variety of insect 

and plant pests. Pesticides are usually applied as a 

formulation containing the active ingredient, along 

with other materials, such as solvents, wetting agents 

or carriers (Pait et al., 1992). Some active ingredients 

have the potential to impact the endocrine system, as 

do some of the surfactants used in the formulations. 

The carbamate insecticide carbofuran, has been 

shown to inhibit oocyte development in at least one 

7
Types and Sources of EDCs 

species of fish (Sukumar and Karpagaganapathy, 

1992). Tennant et al. (1994), working with rats, 

concluded that while the chloro-s-triazine herbicides 

atrazine and simazine did not possess any intrinsic 

estrogenic activity, these two compounds were 

capable of weak inhibition of estrogen-stimulated 

responses in the rat uterus (i.e., effect on progesterone receptor binding, and thymidine incorporation 

into uterine DNA). More recently, Crain et al. (1997) 

showed that atrazine has the ability to stimulate 

production of the enzyme aromatase which converts 

androgens to estrogens, and presumably could 

interfere with sexual differentiation and development. 

DDT, the pesticide banned 

in early 1970s, has been 

shown to induce production of the egg protein 

vitellogenin in primary fish 

hepatocytes (Celius et al., 

1999). The metabolites of 

DDT also appear capable of 

impacting the endocrine 

system. Donohoe and 

Curtis (1996) have shown 

that o’p’-DDE is estrogenic, 

however, p,p’-DDE, the 

dominant persistent 

metabolite of DDT, is not 

estrogenic using a variety 

of assays (Safe, 1995; 

Donohoe and Curtis, 1996). 

Kelce et al. (1995) while 

confirming that it had little 

ability to bind to the 

estrogen receptor in rats, 

found that p,p’-DDE was a 

potent androgen antagonist 

in male rats. 

Trace elements occur naturally in the earth’s crust but 

are concentrated and introduced into the environment through mining and manufacturing processes. 

Kime (1999) has pointed out that trace elements 

which induce the production of metallothioneins in 

the liver or gonads might disrupt gamete production 

by disturbing normal zinc homeostasis, essential for 

the development of both eggs and sperm. Cadmium, 

at aqueous concentrations of 50 ppm, has been 

shown to significantly decrease sperm motility (Kime 

et al., 1996). Mercury has been shown to have a major 

impact on sperm motility at a concentration of only 1 

ppb (Rurangwa et al., 1998), and to have a direct 

effect on the egg micropyle, preventing entry of 

sperm (Khan and Weis, 1993). In rainbow trout, 

exposure to lead resulted in smaller oocytes (eggs) 

(Ruby et al., 2000). 

A number of phytoestrogens, including genistein, 

daidzein and enterodiol are known to affect the 

endocrine system. An infertility syndrome in sheep, 

known as clover disease, can be found in animals 

grazing on subterranean clover (Cheek et al., 1998). 

Hughes (1988) has gone as far to suggest that phytoestrogens might actually be a defense strategy by 

plants to limit the fertility of grazing herbivores. 

Fungi are also known to produce several toxins 

which can affect the 

endocrine system (Celius 

et al. 1999). One of these 

is the mycotoxin 

zearalenone produced by 

Fusarium, a common 

contaminant in cereals 

and other plant products. 

Effluents from pulp and 

paper mills can affect 

reproductive function in 

fish (Tremblay and Van 

Der Kraak, 1998). One of 

the compounds that 

could be responsible is βsitosterol, a major byproduct of wood pulp 

delignification, which 

appears to be activated 

or produced in the 

presence of Mycobacterium smegmatis (Bortone 

and Cody, 1999; 

MacLatchy et al., 1997). 

In addition to the compound types discussed above, 

there is growing acknowledgment that natural and 

synthetic estrogens, when present, are likely responsible for at least some of the endocrine-related effects 

seen in fish, particularly near STPs (Harries et al., 

1999; Purdom et al., 1994; Larsson et al., 1999). The 

synthetic estrogen 17α-ethinylestradiol (Figure 4), 

used in oral contraceptives, has been found in STP 

effluents along with E2

 and the E2 metabolites estrone 

and estriol. The concentration of the natural and 

synthetic hormones needed to elicit a response, such 

as vitellogenin production in male fish, is typically 

hundreds or thousands of times less than that of an 

EDC such as 4-nonylphenol. 

Industrial activities can be the source of a variety of 

endocrine disrupting compounds including surfactants 

and polycyclic aromatic hydrocarbons. Image courtesy 

of NOAA/NOS. 

8
Table 3. Relative potencies of estrogenic compounds. 

Compound Relative Potencya Mean ED 50 

17β-estradiol 1.0 1.81 nM 

4-nonylphenol (4NP) 9.00E-06 16.15 μM 

4-tert-butylphenol (4-tBP) 1.60E-04 2.06 μM 

4-tert-octylphenol (4-tOP) 3.70E-05 2.11 μM 

nonylphenol diethoxylate (NP2EO) 6.00E-06 17.27 μM 

nonylphenoxycarboxylic acid (NP1EC) 6.30E-06 15.25 μM 

a

Refers to the mean potency of each compound relative to 17β-estradiol 

Source: Jobling and Sumpter, 1993. Reprinted by permission. 

Laboratory Studies 

 LABORATORY STUDIES 

This section contains a review of some of the laboratory studies that have addressed effects of EDCs in 

fish. A review of all laboratory investigations was 

beyond the scope of the report. Kime (1998) has 

completed an excellent and exhaustive review of the 

effects of EDCs in fish, primarily from laboratory 

studies. 

Alkylphenol Polyethoxylates. The APEs are estrogenic 

and are perhaps the most studied of the EDCs, 

particularly in male fish, in part because of their 

potency and presence in the environment. Effects 

associated with APEs typically include elevated 

levels of the egg protein vitellogenin in males, 

reduction in gonadal development, and even intersex 

fish. 

Jobling and Sumpter (1993) conducted a series of in 

vitro laboratory experiments using nonylphenol 

ethoxylate metabolites in rainbow trout 

(Oncorhynchus 

mykiss). In 

that paper, 

they introduced the 

production 

of vitellogenin in male 

fish as an 

indication of 

an estrogenic 

effect. 

Vitellogenin 

is now a 

widely used 

biomarker of 

exposure to estrogenic compounds. Jobling and 

Sumpter (1993) found that the metabolites of nonylphenol polyethoxylate, namely 4-nonylphenol, 

nonylphenol diethoxylate (NP2EO) and 

nonylphenoxycarboxylic acid (NP1EC) were all 

estrogenic in male rainbow trout, with potencies, 

however, between 1 x 10-4 to 1 x 10-6 the activity of the 

natural estrogen E2

 (Table 3). 

White et al. (1994) also investigated the ability of 

selected APEs to stimulate vitellogenin synthesis in 

rainbow trout. They found that octylphenol, 4­

nonylphenol, NP1EC and NP2EO were each capable 

of inducing vitellogenin synthesis. A summary of 

vitellogenin synthesis from cultured hepatocytes in 

comparison to E2

 is shown in Figure 5 for each of the 

compounds tested on a molar basis. The concentration of the alkylphenols or APEs needed to elicit a 

response similar to E2

, however, was much higher. 

or example, at 10-5 M octylphenol, the amount of 

itellogenin produced was roughly equal to that in 

ells exposed to only 10-8 M E2

, roughly a 1,000-fold 

oncentration difference. 

Jobling et al. (1996) assessed the effects of APEs and 

their metabolites in whole rainbow trout. They 

looked at vitellogenin synthesis and inhibition of 

testicular growth using 4-nonylphenol, NP1EC and 

NP2EO. After a 3-week exposure, all of the alkylphenolic compounds tested caused significant elevations 

in plasma vitellogenin. The order of potency was 

octylphenol > 4-nonylphenol > NP2EO > NP1EC, 

similar to what White et al. (1994) found working 

with rainbow trout hepatocytes. 

At a concentration of 30 μg/L, octylphenol produced 

a million-fold increase in the concentration of vitellogenin, while the concentration of vitellogenin in fish

exposed to 4­

nonylphenol, 

NP2EO, and 

NP1EC were 

elevated less, 

yet still by 100­

to 1,000-fold 

over background levels. 

The potent 

synthetic 

estrogen 17α

ethinylestradiol (EE2

)

(Figure 4) at a 

concentration of only 2 ng/L produced a similar 

effect. They also conducted a dose-response experiment using 4-nonylphenol and found that the lowest 

concentration needed to induce a significant elevation of plasma vitellogenin was approximately 20 

μg/L. 

Jobling et al. (1996) noted that the concentrations 

required to induce a response in vivo were two orders 

of magnitude lower than the concentrations needed 

in earlier in vitro experiments (Jobling and Sumpter, 

1993). They suggested the disparity could be due to 

the bioaccumulative properties of the compounds 

and/or the metabolism of chemicals by the fish to 

more active metabolites. This suggests that over time 

accumulation of lower concentrations of APEs and/ 

F

v

c

c

­

9
Laboratory Studies 

or their metabolites could elicit a response from fish 

in the environment. Table 4 summarizes some of the 

effects of EDCs on whole fish in the laboratory, while 

Table 5 summarizes 

some of the in vitro 

effects with primary 

hepatocyte cultures. 

Hemmer et al. (2001) 

investigated the effects 

of 4-nonylphenol on 

vitellogenin production in male sheepshead minnows 

(Cyprinodon 

variegatus). Fish were 

exposed to 0.6, 5.4, 

11.8, 23.3, and 42.7 μg/ 

L 4-nonylphenol in a 

flow through system. 

Sampling occurred on 

days 0, 2, 5, 13, 21, 35, 

and 42. Vitellogenin 

was measured using a 

direct enzyme-linked 

immunosorbent assay 

(ELISA). The sheepshead minnows 

demonstrated a rapid, dose-dependent induction of 

plasma vitellogenin (Hemmer et al., 2001). By the 

fifth day, fish exposed to 5.38 μg/L 4-nonylphenol or 

higher had significantly higher vitellogenin concentrations than fish in the control tanks. Fish produced 

as much as 100 mg vitellogenin/ml of plasma by day 

13 in the highest exposure 

(42.7 μg/L), and by day 42, 

males had produced an even 

higher vitellogenin concentration, approximately 130 mg/ 

ml. 

Pait and Nelson (in prep.) 

assessed the ability of male 

killifish (Fundulus heteroclitus) 

to produce vitellogenin in 

response to intraperitoneal 

injections of 4-nonylphenol 

over a period of 8 days 

(Figure 6). F. heteroclitus 

responded significantly to 

injections of 50 mg/kg and 

greater 4-nonylphenol, with 

the 150 mg/kg exposure 

resulting in a plasma vitellogenin concentration of 

approximately 23 mg/ml. 

Lewis and Lech (1996) 

investigated the uptake, 

disposition, and persistence of 14C-nonylphenol 

in rainbow trout. The 

compound was rapidly 

taken up into most 

tissues and had an 

apparent half-life of 19­

20 hours in muscle and 

fat. Using thin layer 

chromatography, three 

glucuronide metabolites 

were separated, however, the structure of 

these metabolites was 

not positively identified. 

They concluded that 4­

nonylphenol appears to 

be metabolized through 

oxidation and glucuronic 

acid conjugation with 

subsequent excretion in 

the bile. 

Arukwe et al. (2000) assessed the metabolism and 

organ distribution of 4-nonylphenol in Atlantic 

salmon (Salmo salar). Fish were given a single dose of 3

H-nonylphenol (25 μg) and then sampled at 24, 48, 

and 72 hours. Similar to Lewis and Lech (1996), 

Arukwe et al. (2000) found that 4-nonylphenol was 

conjugated to glucuronic acid. 

Nonylphenol-glucuronide 

was the main metabolite 

found in both biliary and 

urinary samples. In biliary 

samples, between 80 and 95 

percent of the total radioactivity was present as 

nonylphenol-glucuronide. 

Hydroxylated and oxidated 

compounds made up the 

remainder of the metabolites 

identified. 

Gray and Metcalfe (1997) 

investigated the effects of 4­

nonylphenol on Japanese 

medaka (Oryzias latipes). 

Medaka were exposed from 

hatch to 3 months of age to 

Vitellogenin (ug/ml) 

0.8 

0.6 

0.4 

0.2 

0 

-8 

[Ligand] 

-7 -6 -5 -7 -6 -5 -7 -6 -5 -7 -6 -5 

E2 OP NP NP2EO NP1EC 

Figure 5. Vitellogenin production from cultured 

hepatocytes of rainbow trout in response to alkylphenolic compounds and 17b-estradiol (E2). Compounds 

were tested at 10-8 M (-8), 10-7M (-7), 10-6M (-6), and 10 5

M (-5). From White et al., 1994. Reprinted by permission. 

­

0 10 50 100 150 

Vitellogenin (mg/ml)

30 

20 

10 

0 

4-Nonylphenol (mg/kg) 

Figure 6. Production of vitellogenin (mean 

±SEM) in male killifish (Fundulus 

heteroclitus) exposed to 4-nonylphenol. 

(Pait and Nelson, in prep.) 

10
Table 4. Effects of endocrine disrupting compounds on whole fish in the laboratory.

Fish/Species Sex Compound Method of Exposure Concentration Duration Vtg T E2 GSI GD I Reference

Rainbow trout 

(Oncorhynchus mykiss) male 4-Nonylphenol whole fish (water) 30 μg/L 3 weeks + + Jobling et al. , 1996

Killifish 

(Fundulus heteroclitus) male 4-Nonylphenol whole fish (intraperitoneal inj.) 50-150 mg/kg 8 days + + Pait and Nelson, in prep.

Sheepshead minnow 

(Cyprinodon variegatus) male 4-Nonylphenol whole fish (water) 0.6-42.7 μg/L 42 days + Hemmer et al. , 2001

Japanese medaka 

(Oryzias latipes) male 4-Nonylphenol whole fish (water) 50 μg/L 3 months + – + Gray and Metcalfe, 1997

Japanese medaka 

(Oryzias latipes) juvenile 4-Nonylphenol whole fish (water) 17.7 - 51.5 μg/L 104 days + Yokota et al. , 2001

Summer flounder 

(Paralichthys dentatus) male Octylphenol whole fish (intraperitoneal inj.) 2 - 200 mg/kg 8 weeks – Mills et al. , 2001

Fathead minnows 

(Pimphales promelas) male Bisphenol-A whole fish (water) 1 - 1,280 μg/L 164 days + – – Sohoni et al. , 2001

Rainbow trout 

(Oncorhynchus mykiss) juvenile Bisphenol-A whole fish (water) 10 - 500 μg/L 12 days + Lindholst et al. , 2000

Rainbow trout 

(Oncorhynchus mykiss) juvenile Bisphenol-A whole fish (intraperitoneal inj.) 50 mg/kg 18 days + Christiansen et al. , 2000

Common carp 

(Cyprinus carpio) male 4-tert -Pentylphenol whole fish (water) 3-1,000 μg/L 3 months + – Gimeno et al. , 1998a

Common carp 

(Cyprinus carpio) male 4-tert -Pentylphenol whole fish (water) 1,000 μg/L 140 days – + Gimeno et al. , 1998b

Japanese medaka 

(Oryzias latipes) male 17α−Ethinylestradiol whole fish (water) 0.1 μg/L 2 months + Scholz and Gutzeit, 2000

Fathead minnows 

(Pimphales promelas) male 17α−Ethinylestradiol whole fish (water) 0.064 μg/L 305 days – + Lange et al ., 2001

Japanese medaka 

(Oryzias latipes) juvenile 17α−Ethinylestradiol whole fish (water) 0.1 μg/L 110 days + Metcalfe et al. , 2001

Rainbow trout 

(Oncorhynchus mykiss) female β-Napthoflavone and 17β-estradiol whole fish (intraperitoneal inj.) 50 mg/kg β-NF and 0.5 mg/kg E 2 24 hours – Anderson et al. , 1996b

Atlantic croaker 

(Micropogonias undulatus) female Benzo[a ]pyrene whole fish (diet) 0.4 mg/70g 30 days – – – Thomas 1988

NE, no effect; Vtg, vitellogenin; T, testosterone; E2, 17β-estradiol; GSI, gonadosomatic index; GD, gonadal development; I, intersex condition

Laboratory Studies

11
Laboratory Studies Fish/Species Sex Compound Method of Exposure Concentration Duration Vtg T E2 GSI GD I Reference (Platichthys flesus) chrysene wh 12 weeks – et al.

(Lepidopsetta bilineata) 7 days – et al

(Micropogonias undulatus)

– Thomas 1988

(Micropogonias undulatus) 1254 – Thomas 1989

(Morone americana) 9 weeks NE NE – Monosson et al.

(Platichthys stellatus) Na unknown –

(Oryzias latipes) o'p + + Cheek et al.

(Micropogonias undulatus) o'p who e sh d e DDT/g +

(Cyprinodon variegatus) /L NE Hemmer et al.

(Clarias batrachus)

– et al.

(Colisa lalia)

–

Sukumar and 

(Cyprinodon variegatus) + Hemmer et al.

(Oncorhynchus mykiss) 9 days + et al

(Oncorhynchus mykiss) Lead 10 μg/L – Ruby et al.

(Micropogonias undulatus) Lead –

(Micropogonias undulatus) 1 mg/L + +

(Oncorhynchus mykiss) β 75 3 weeks + –

β – – et al.

β

Table 4. Effects of endocrine disrupting compounds on whole fish in the laboratory (continued).

European flounder Phenanthrene and female ole fish (diet) 0.4 - 12.5 nM/g food Monteiro , 2000b

Rock sole whole fish female Prudhoe Bay crude oil 0.1 - 1.0 mg/kg Johnson ., 1993 (intraperitoneal inj.)

Atlantic croaker female Aroclor 1254 whole fish (diet) 0.24 mg/70g 30 days

Atlantic croaker female Aroclor whole fish (diet) 0.24 mg/70g 30 days

White perch 3,3',4,4'- whole fish female 0.2-5.0 mg PCB/kg , 1994 Tetrachlorobiphenyl (intraperitoneal inj.)

Starry flounder female tural PCB exposure whole fish (water/diet) life Spies and Rice, 1988

Japanese medaka juvenile -DDT whole fish (water) 7.5 μg/L 2-8 weeks , 2001

Atlantic croaker 0.02 and 0.01 μg female -DDT l fi ( i t) 3-7 weeks Khan and Thomas, 1998

Sheepshead minnow male Endosulfan whole fish (water) 0.28 - 0.79 μg 40 days , 2001

Walking catfish female Endosulfan whole fish (water) 1.5 μg/L 16 days Chakravorty , 1992

Dwarf gourami female Carbofuran whole fish (water) 0.7 μg/L 20 days Karpagaganapathy, 1992

Sheepshead minnow male Methoxychlor whole fish (water) 12.1 μg /L 35 days , 2001

Rainbow trout juvenile Butylbenzyl phthalate whole fish (water) 500 mg/kg Christiansen ., 1998

Rainbow trout female whole fish (water) 12 days , 2000

Atlantic croaker female whole fish (diet) 1.34 mg/70g 30 days Thomas, 1988

Atlantic croaker female Cadmium whole fish (water) 30 days Thomas, 1989

Rainbow trout Tremblay and van Der juvenile -Sitosterol whole fish (water) - 100 μg/L Kraak, 1998

Goldfish males/females -Sitosterol whole fish (water) 75- 1,200 μg/L 12 days MacLatchy , 1997

NE, no effect; Vtg, vitellogenin; T, testosterone; E2, 17-estradiol; GSI, gonadosomatic index; GD, gonadal development; I, intersex condition

12
Table 5. In vitro (fish hepatocyte) effects of endocrine disrupting compounds.

Fish/Species Compound Concentration Duration Vtg Vtg Production Reference

Rainbow trout 

(Oncorhynchus mykiss) 4-Nonylphenol 10 μM 48 hours + 0.25 μg/ml Jobling and Sumpter, 1993

Rainbow trout 

(Oncorhynchus mykiss) 4-Nonylphenol 1-10 μM 48 hours + 0-0.45 μg/ml (approx.) Sumpter and Jobling, 1995

Rainbow trout 

(Oncorhynchus mykiss) 4-Octylphenol 10 μM 48 hours + 0.95 μg/ml (approx.) Jobling and Sumpter, 1993

Rainbow trout 

(Oncorhynchus mykiss) 4-Octylphenol 10 μM 48 hours + 0.2-0.4 μg/ml (approx.) White et al. , 1994

Rainbow trout 

(Oncorhynchus mykiss) Bisphenol-A 1-10 μM 48 hours + 0.2 - 1.7 μg/ml (approx.) Sumpter and Jobling, 1995

Atlantic salmon 

(Salmo salar) Bisphenol-A 1-10 μM 48 hours + Absorbence units; >controls Celius et al. , 1999

Rainbow trout 

(Oncorhynchus mykiss)

4-nonylphenoxycarboxylic acid (4-

NPEO) 10 μM 48 hours + 0-0.17 μg/ml (approx.) White et al. , 1994

Rainbow trout 

(Oncorhynchus mykiss)

β-Napthoflavone and 3-

methylcholanthrene 0.021 - 6.2 μM 48 hours – N A Navas and Segner, 2000

Rainbow trout 

(Oncorhynchus mykiss) Soot and 17β-estradiol 0.05-25μg/L 48 hours – N A Villalobos et al. , 1996

Rainbow trout 

(Oncorhynchus mykiss) Aroclor 1221 1-10μM 48 hours + 0.2 - 0.4 μg/ml (approx.) Sumpter and Jobling, 1995

Rainbow trout 

(Oncorhynchus mykiss)

2,3,4,7,8-Pentachlorodibenzofuran and

17β-estradiol 10-12-10-8M 48 hours – Production < with E2 alone Anderson et al ., 1996a

Rainbow trout 

(Oncorhynchus mykiss)

2,3,7,8-tetrachlorodibenzo-p -dioxin (TCDD) and 17β-estradiol 10-11-10-8M 48 hours – Production < with E2 alone Anderson et al ., 1996a

Atlantic salmon 

(Salmo salar) o'p -DDT and lindane 1-10 μM 96 hours + Absorbence units; >controls Celius et al. , 1999

Vtg, vitellogenin; NA, not available from article

Laboratory Studies

13
Laboratory Studies 

nominal concentrations of 10, 50, and 100 μg/L. At 

50 and 100 μg/L, 50 and 86 percent of the males, 

respectively, developed ovotestes, an intersex condition characterized by both testicular and ovarian 

tissue within the gonad. In addition, while the ratio 

of males to females in the control was 2:1, the ratio 

declined to 1:2 in the higher 4-nonylphenol treatment. 

In a more recent study, Yokota et al. (2001) addressed 

life cycle toxicity of 4-nonylphenol to Japanese 

medaka. Exposure began at 24 hours 

postfertilization and continued for up to 104 days 

posthatch for the F , and 60 days for the F genera­ 0 1 

tion. Both generations of medaka were exposed 

under flow through conditions to 4-nonylphenol 

concentrations of 4.2, 8.2, 17.7, 51.5 and 183 μg/L. 

The highest concentration significantly reduced 

survival (hatching and swim-up) of the F fish. At 60 0 

days posthatch, external secondary sex characteristics appeared totally skewed towards females in the 

51.5 μg/L treatment. Gonadal histology revealed 

that 20 percent of the fish in the 17.7 μg/L treatment, 

and 40 percent in the 51.5 μg/L treatment had 

ovotestes. Because of the skewed sex ratio, no fish 

from the 51.5 μg/L treatment could be used to 

examine fertility and fecundity in the F generation, 0 

and only three pairs could be selected from the 17.7 

μg/L treatment. Fish were exposed from day 71 to 

103 posthatch with concentrations ranging from 4.2 

to 17.7 μg/L. While fecundity was not affected by 

the treatments, mean fertility was reduced by 76 

percent in the 17.6 μg/L treatment. In the F genera­ 1 

tion, significant ovotestes were observed in the 8.2 (4 

percent) and the 17.7 μg/L (18 percent) concentrations. 

Mills et al. (2001) investigated the effects of 

octylphenol on juvenile male summer flounder 

(Paralichthys dentatus). Flounder were injected twice 

with octylphenol in a coconut oil carrier during the 

experiment for total concentrations of 2, 20, and 200 

mg/kg plus a control. Experiments were run for 8 

weeks in flow-through aquariums, and fish were 

sampled at 4, 6 and 8 weeks. None of the doses of 

octylphenol, however, resulted in the production of 

detectable amounts of vitellogenin in the juvenile 

male flounder as determined by direct ELISA, 

although the gonadosomatic index, or GSI (gonad 

weight/body weight X 100) showed some evidence 

of a decline in fish exposed to 200 mg/kg 4­

nonylphenol. 

Bisphenol-A. This compound is also estrogenic, 

capable of inducing vitellogenin in male fish. The 

myriad of uses and associated high production 

volume for bisphenol-A has heightened concern for 

its possible effects in the aquatic environment. 

Lindholst et al. (2000) studied the estrogenic response 

of rainbow trout exposed to bisphenol-A in a flowthrough system. Juvenile fish were exposed to 

nominal concentrations of 10, 40, 70, 100, and 500 μg/ 

L of bisphenol-A for 6 and 12 days. Plasma samples 

were analyzed for vitellogenin using a direct sandwich ELISA. While exposure to increasing concentrations of bisphenol-A resulted in increasing levels 

of vitellogenin after 6 days, only the 500 μg/L 

concentration produced significantly higher 

vitellogenin levels (approximately 1 mg/ml) than 

controls. Interestingly, after 12 days the concentration of plasma vitellogenin in those fish exposed to 

70 or 100 μg/L bisphenol-A remained the same or 

decreased slightly. Lindholst et al. (2000) attributed 

this to a possible increase in the degradation rate of 

bisphenol-A as a result of cytochrome P450 induction. After 12 days, vitellogenin levels in fish exposed to 500 μg/L continued to increase, which the 

authors suggested was the result of detoxification 

processes not being able to keep pace with the higher 

concentration of bisphenol-A. 

Sohoni et al. (2001) investigated the long-term 

reproductive effects of bisphenol-A in fathead 

minnows (Pimphales promelas). Mature male and 

female fish were exposed to nominal bisphenol-A 

concentrations of 1, 16, 160, 640 or 1,280 μg/L in 

flow-through chambers. At days 43, 71, and 164, 

selected fish were sacrificed for measurements of 

somatic growth, GSI, and plasma vitellogenin 

(ELISA). The testes were processed for histologic 

analysis. At concentrations of 640 and 1,280 μg/L, 

bisphenol-A had a significant inhibitory effect on 

weight and length in males by Day 71. 

In males, exposure to bisphenol at concentrations of 

160 μg/L and higher for at least 71 days resulted in 

vitellogenin concentrations between 120 and 144 μg/ 

L. The GSIs were also significantly decreased in both 

males and females exposed to bisphenol-A at concentrations of 640 μg/L and higher for 164 days. Histological examination revealed a decrease in spermatozoa in fish exposed to nominal bisphenol-A concentrations of 16 μg/L and greater for 164 days. Results 

from breeding pair experiments begun on Day 42 

indicated that egg production was inhibited at 1,280 

μg/L, and that hatchability was reduced at a nominal 

bisphenol-A concentration of 640 μg/L. 

14
Laboratory Studies 

Christiansen et al. (2000) investigated the potential 

for a number of xenoestrogens, including bisphenolA, to induce vitellogenesis in juvenile rainbow trout. 

Fish were injected with 50 mg/kg in a peanut oil 

carrier at days 0, 6, and 12 of an 18 day exposure. 

The average plasma vitellogenin concentration after 

18 days was 0.066 mg/ml. Christiansen et al. (2000) 

noted there was substantial variability between fish, 

with vitellogenin concentrations ranging from 0.07 to 

482 μg/ml. 

Celius et al. (1999) looked at the production of 

vitellogenin and the induction of zona radiata 

proteins (eggshell proteins) in primary hepatocytes 

of Atlantic salmon exposed to bisphenol-A. Cells 

treated with 1, 5, or 

10 uM of bisphenolA produced both 

vitellogenin and 

zona radiata proteins 

in an approximate 

dose-dependant 

manner after 96 and 

48 hours, respectively. 

Sumpter and Jobling 

(1995) assessed the 

ability of bisphenolA to induce vitellogenin production at a concentration of 1 μM in male rainbow trout hepatocytes. 

Bisphenol-A induced significant vitellogenin production after 2 days. 

4-tert-pentylphenol. This compound has also shown 

some estrogenic potential. Gimeno et al. (1998a) 

investigated the effects of 4-tert-pentylphenol (TPP), 

an industrial intermediate, on the common carp 

Cyprinus carpio during spermatogenesis. Sexually 

mature male carp were exposed for up to 3 months to 

nominal concentrations ranging from 32 to 1,000 μg/ 

L using an intermittent flow-through dosing system. 

Exposure of the male carp to TPP resulted in elevated 

vitellogenin levels and the inhibition of spermatogenesis. Exposure of fish to each concentration also 

resulted in reduced testicular growth, with a correspondingly lower testicular weight. In three out of 

five fish exposed to the higher TPP concentrations, 

disorganization of the seminiferous lobules, atrophy 

of the germinal epithelium, and some tissue necrosis 

was observed. While exposure of adult male carp to 

TPP for 3 months did not result in intersex fish, 

exposure to a high concentration of E2

 resulted in the 

development of ovotestes in one mature fish. 

Gimeno et al. (1998b) investigated the feminizing 

effects of TPP on young male carp. Genetic males 

(XY) were exposed at 50 days post hatch to nominal 

TPP concentrations of 100, 320, and 1,000 μg/L using 

an intermittent flow-through dosing system. Histological examination of the gonads was carried out 

beginning at day 20 and then continued every 10 

days until 140 days post hatch. Gimeno et al. (1998b) 

found that spermatogenesis was severely inhibited in 

the testes of the young carp. At the two highest TPP 

concentrations, a few oocytes were detected in the 

testes, along with an oviduct. 

Polycyclic Aromatic 

Hydrocarbons (PAHs). The 

PAHs appear to depress 

the production or levels 

of circulating estrogens 

as well as vitellogenin in 

female fish. There does 

not appear to be much 

information on the effects 

of PAHs in male fish. 

PAHs have also been 

shown to result in lower 

GSIs. 

A number of studies have examined the relationship 

between compounds that induce the cytochrome 

P4501A1 isozyme (CYP1A1), namely the polycyclic 

and some of the halogenated aromatic hydrocarbons, 

and endocrine disruption. This effect has been linked 

to compounds that bind to the aryl hydrocarbon (Ah) 

receptor, which functions in an analogous way to the 

ER, by forming a complex with the xenobiotic. In a 

paper by Anderson et al. (1996b), a set of experiments 

were run in vivo with juvenile rainbow trout. Fish 

were injected with various amounts of E2

 along with 

βNF (β-napthoflavone), a PAH. Trout injected with 

0.5 mg/kg E2

 plus 25 or 50 mg/kg βNF had depressed vitellogenin synthesis by the liver relative to 

E2

 injected alone. They attributed inhibition of

vitellogenin synthesis to an estrogen receptor “down 

regulation”, that is, a lowered amount of estrogen 

receptor present or available as evidenced by a 

depressed estradiol-estrogen receptor binding 

capacity of liver in trout injected with βNF. 

At higher levels of E2

 the effect was reversed, which 

he authors attributed to a possible phosphorylation 

Rainbow trout, Oncorhynchus mykiss, a fish frequently used 

in endocrine disrupter research. Image courtesy of USDOI/ 

FWS - Duane Raver. 

t

15
Laboratory Studies 

of certain cellular proteins which might synergistically activate ER-mediated gene transcription with 

E2

, independent of any changes in cellular ER content, or changes in the binding affinity of the ER for 

the ligand (Anderson et al., 1996b). They also noted 

that the antiestrogenic effect of βNF at lower concentrations could be caused by increased metabolism of 

E2

, as a result of Phase II conjugation. Santodonato 

(1997) noted that although PAHs may function as 

weak ER agonists, they are expected to bind preferentially to the Ah receptor, triggering the induction of 

the Ah-responsive genes (e.g., P450) more likely 

leading to an antiestrogenic effect. 

Navas and Segner (2000) investigated the 

antiestrogenicity of PAHs to cultured rainbow trout 

hepatocytes along with the role of the Ah receptor. 

Cells were co-exposed to PAHs and E2

, and the 

effects on 7-ethoxyresorufin-O-deethylase (EROD) 

activity (Ah receptor-regulated), and vitellogenin 

production (E2

-regulated) were assessed. Cells were 

first pre-exposed for 24 hours to serial dilutions 

(0.021-6.2 uM) of the test compounds (3-methylcholanthrene, anthracene, or β-napthoflavone), 

followed by addition of 1 uM E2

 along with the test 

compounds. Anthracene had no effect on either 

vitellogenin synthesis or EROD activity. However, 

vitellogenin synthesis decreased with increasing 

concentrations of 3-methylcholanthrene or βnapthoflavone, even in the presence of E2

. EROD 

activity, however, was found to increase with increasing concentrations of all these compounds. The 

compound with the highest EROD induction activity, 

β-napthaflavone, was also the strongest inhibitor of 

vitellogenin production. The addition of either of 

two Ah receptor inhibitors, α-napthoflavone, or 8­

methoxypsoralen (6.25 or 12.54 uM) resulted in a 

reversal of the inhibition. Because there were no 

significant differences in the E2

 concentrations in the 

control or induced cultures, Navas and Segner (2000) 

concluded that elevated levels of CYP1A1 were not 

associated with significant biotransformation of E2

 in 

the treatment cultures, which could have resulted in 

lower vitellogenin levels. Instead, the authors 

hypothesized that lower vitellogenin levels observed 

in those cultures treated with PAHs may have been 

the result of the activated Ah receptor and xenobiotic 

response element interacting or possibly inhibiting 

the regulatory regions of the vitellogenin gene. 

Monteiro et al. (2000a) assessed the ability of PAHs to 

inhibit in vitro ovarian steroidogenesis in the European flounder Platichthys flesus. Fully vitellogenic 

ovarian tissue was incubated for 24 hours with either 

phenanthrene, benzo[a]pyrene or chrysene (15μM), 

along with the steroid precursor 17ahydroxyprogesterone, at a concentration of 0.15 μM. 

All three PAHs significantly reduced the production 

of androstenedione and E .

2

In a followup study, Monteiro et al. (2000b) assessed 

the effect in vivo of phenanthrene or chrysene on 

plasma steroid levels in female European flounder. 

Fish were given the PAHs as part of the diet for a 

period of 12 weeks during the previtellogenic phase 

of the reproductive cycle. While neither phenanthrene or chrysene appeared to affect ovarian development, there was a significant effect on plasma E2 

levels. In fish exposed to 0.4 nmol chrysene/gram of 

food, or 0.5-12.5 nmol phenanthrene/gram of food, 

plasma E2

 levels decreased significantly. In female 

flounder exposed to 12.5 nmol phenanthrene, plasma 

E2

 levels were 21 pg/ml, compared to 107 pg/ml in 

control fish. 

Thomas (1988) investigated dietary effects of the 

PAH benzo[a]pyrene on endocrine function in female 

Atlantic croaker, Micropogonias undulatus. The 

experiment was carried out over 30 days during 

gonadal recrudescence (seasonal maturation). 

Benzo[a]pyrene at a rate of 0.4 mg/70g fish/day 

significantly impaired ovarian development in the 

fish. In females, the GSI was only 66 percent of 

controls (Figure 7). In addition, reduced ovarian 

growth was accompanied by a significant decline in 

circulating E2

 levels (Thomas, 1988). 

Johnson et al. (1993) injected gravid female rock sole 

(Lepidopsetta bilineata) and flathead sole 

(Hippoglossoides elassododon) from a nonurban area in 

Puget Sound with Prudhoe Bay Crude Oil (PBCO) at 

doses of 0.1, 0.5, and 1.0 mg oil/kg body weight. 

Injection of the fish with either fresh or weathered 

PBCO resulted in decreased plasma E2

 levels after 7 

days. Experiments with ovarian tissue fragments 

exposed to weathered PBCO also appeared to inhibit 

ovarian E2

 production in a dose-dependent manner. 

In work containing aspects of both a laboratory and 

field study, Knudsen et al. (1997) exposed male 

rainbow trout in a series of 500 L tanks to treated 

effluent collected from an oil refinery. Treatment 

involved skimming, sedimentation and biological 

degradation. The tanks containing the fish received 

drinking quality water mixed with either 0, 2.5, 10, or 

20 percent treated refinery effluent. At an effluent 

concentration of 20 percent, fish that were exposed 

for either 2 or 3 weeks produced a significantly 

16
15 3 

GSI 

17β-estradiol 

Control 

Lead 

Benzo[a] 

pyrene 

Aroclor 1254 

Figure 7. Effect of various compounds on GSI and plasma 17βestradiol in female Atlantic croaker (Micropogonias undulatus). 

From Thomas, 1988. Reprinted by permission. 

10 

-estradiol (ng/ml) β Plasma 17

2

GSI 

5 1 

0 0 

Compound 

Laboratory Studies 

higher concentration of vitellogenin than control fish. 

Knudsen et al. (1997) noted that while a rather high 

concentration of the effluent was needed to produce 

an effect, the treated effluent clearly contained 

bioactive substances. Although the chemical(s) 

responsible for the elevated vitellogenin levels were 

not identified, the authors felt that removal of certain 

refinery process chemicals as opposed to enhanced 

treatment might be a cost effective means of addressing the observed effects. 

Villalobos et al. 

(1996) exposed a 

primary liver cell 

culture from rainbow trout to an 

extract of soot. 

Exposure of the cells

to the extract at a 

concentration of 

0.05-25 μg/L along 

with an E2

 concentration of 1 uM 

caused an induction 

of CYP1A1 protein 

after binding to the 

Ah receptor, followed by a reduction in E2

-dependent

vitellogenin synthesis, an antiestrogenic effect. 

Polychlorinated 

Biphenyls. In 

general, PCBs 

appear to depress 

levels of estrogens as well as gonadal development 

and the GSI in female fish. Little information was 

found on effects in male fish. Although no longer 

manufactured, the widespread occurrence of PCBs in 

the environment coupled with their effects on the 

endocrine system is of concern. 

Thomas (1988) investigated the effect of Aroclor 1254 

on Atlantic croaker. Females fed Aroclor 1254 in a 

prepared diet (0.24 mg/70g fish/day) for 30 days 

had substantially lower GSIs (25 percent of controls) 

(Figure 7). Thomas (1989) conducted additional 

experiments with Aroclor 1254 and found that in 

addition to reduced ovarian growth in croaker, 

plasma E2

 levels were significantly reduced. Thomas 

(1989) hypothesized that the lower E2

 levels decreased vitellogenesis and hence ovarian growth. 

Monosson et al. (1994) injected groups of female adult 

white perch (Morone americana) with the coplanar 

PCB 3,3’,4,4’-tetrachlorobiphenyl (TCB) at three 

concentrations ranging from 0.2-5.0 mg PCB/kg 

body weight prior to the spawning season. In those 

fish receiving the highest dose of the PCB, fewer 

matured, and in those that did, the GSI was approximately half that of the control fish. Interestingly, 

however, levels of E2

, testosterone and plasma 

vitellogenin were not significantly affected by any of 

the TCB exposures (Monosson et al., 1994). Survival 

of the larvae 

was also 

affected by 

parental exposure. Larvae 

produced from 

females dosed 

with 1.0 and 5.0 

mg/kg had 

significantly 

lower survival 

rates. 

Sumpter and 

Jobling (1995) 

found that a 1 

μM concentration of Aroclor 

1221 actually 

induced vitellogenin production in male 

rainbow trout 

hepatocytes. In 

Figure 8, an 

apparent 

increase in vitellogenin production resulted from 

increasing concentrations of Aroclor 1221. This is a 

seemingly opposite effect of that found by Thomas 

(1988) for Aroclor 1254. Interestingly, when the 

hepatocytes were exposed to a mixture of chemicals, 

all at a concentration of 1 μM, Sumpter and Jobling 

(1995) found an enhanced effect (Figure 9). 

In the laboratory, Spies and Rice (1988) investigated 

reproductive success in starry flounder (Platichthys 

stellatus) from two localities in San Francisco Bay 

with differing degrees of contamination. Mature 

females from both sites were induced to spawn in the 

laboratory using carp pituitary extract, and relationships between measures of survival through successive early life stages, and the degree of contamination 

in both parent and spawned eggs were assessed. 

17
Figure 8. Estrogenic activity of selected compounds in male rainbow trout. Results are presented as the mean ±SEM. From Vitellogenin, ug/ml

2.00 

1.50 

1.00 

0.50 

0

Nonylphenol 

1 5 10 1 5 10 1 5 10 1 5 10 1 5 10 

Octylphenol o,p’-DDT Aroclor 1221 Bisphenol-A 

Concentration, uM 

Sumpter and Jobling, 1995. Reprinted by permission.

Control

Nonylphenol 

’-DDT o,p

Octylphenol

Aroclor 1221

Bisphenol-A

All 5 chemicals 

Compounds, 1 μM 

0.20 

0 

1.40 

1.20 

1.00 

0.80 

0.60 

Vitellogenin, ug/ml 

0.40 

Figure 9. Effect of several estrogenic compounds, 

including a mixture of all five, on vitellogenin production in male rainbow trout. From Sumpter and Jobling, 

1995. Reprinted by permission. 

Laboratory Studies 

Significant negative correlations existed between 

PCB content of the eggs and percent fertilized eggs

that hatched, termed embryological success. There

were also negative correlations between maternal 

hepatic aryl 

hydrocarbon 

hydroxylase 

(AHH) activity 

and percent 

viable eggs, 

fertilization 

success and 

embryological 

success. In 

addition, Spies 

and Rice (1988) 

found that 

females captured 

from the more 

urbanized 

location (Berkeley) had significantly fewer 

viable eggs, lower 

fertilization rates 

and higher AHH activity than fish from the less 

contaminated site in San Pablo Bay. 

Furans and Dioxins. Although little work has been 

done on the effects of furans and dioxins on the 

endocrine system in fish, 

they appear to be 

antiestrogenic, at least in 

vitro, perhaps sharing a 

similar mechanism with 

the PAHs. Using rainbow trout hepatocytes, 

Anderson et al. (1996a) 

investigated the in vitro 

modulation of E2 -

induced vitellogenin 

synthesis by a group of 

compounds, including 

2,3,4,7,8­

pentachlorodibenzofuran,

2,3,7,8­

tetrachlorodibenzo-pdioxin (TCDD), and 

2,3,7,8­

tetrachlorodibenzofuran. 

All were antiestrogenic, 

that is, reducing the 

amount of vitellogenin 

synthesis in trout cells 

coexposed to E2

. Anderson et al. (1996a) noted that 

the potency of inhibition was directly related to the 

strength of the compound as an inducer of CYP1A1 

protein, similar to that found by Navas and Segner 

(2000) for PAHs. 

Polybrominated 

Diphenyl Ethers 

(PBDEs). Since 

the 1960s, polybrominated 

diphenyl ethers 

or PBDEs (Figure 

4) have found 

widespread use 

as flame retardants in products 

from television 

sets to rug and 

drapery textiles. 

Commercial 

PBDE mixtures 

range from pentato deca-BDEs, 

with worldwide 

annual production estimated at 148 million pounds 

(Hooper and McDonald, 2000). PBDEs have been 

found in humans, wildlife, and sediments, and 

residues appear to be on the increase (McDonald, 

2002). Meerts et al. (2001) found that a number of 

PBDEs were estrogenic 

using an estrogen 

responsive human cell 

line. In addition to 

being estrogenic, there 

is also evidence that 

PBDEs can interfere 

with thyroid hormone 

balance and function. 

Mice dosed with

penta-BDE showed 

decreased levels of T4 

after 8 days (Fowles et 

al., 1994). In addition, 

a number of PBDEs or 

their metabolites bind 

with high affinity to 

the thyroid hormone 

transport protein 

transthyretin (Meerts et 

al., 2000), and to the

thyroid hormone 

receptors (Marsh et al., 

1998). Although no 

18 
Laboratory Studies 

articles were located on endocrine related effects of 

PBDEs in fish, this emerging class of EDCs would 

appear to be good candidates for future laboratory 

and field investigations. 

Pesticides. A number of pesticides have been identified as EDCs. Some appear to be antiestrogenic, 

while others, such as o,p’-DDT and methoxychlor, 

appear capable of inducing vitellogenesis in male 

fish. 

Chakravorty et al. (1992) investigated the effect of the 

insecticide endosulfan on vitellogenesis in the catfish 

Clarias batrachus. Female fish were exposed to a 

nominal concentration of 1.5 μg/L endosulfan in a 

static renewal system for 

a period of 16 days. 

There was a significant 

decline in plasma 

vitellogenin compared 

to controls (approximately 0.8 mg/ml) after 

16 hours, with a maximum low (approximately 0.1 μg/ml) 

reached after 48 hours 

(Chakravorty et al. 1992). 

Hemmer et al. (2001) 

assessed the effects of 

endosulfan on 

vitellogenin production 

in male sheepshead 

minnows. Fish were 

exposed to measured 

concentrations ranging 

from 0.28 - 0.79 μg/L for 

up to 40 days in a flow

through system. None of these exposures, however,

resulted in the accumulation of detectable levels of

vitellogenin in the plasma.

Hemmer et al. (2001) also assessed the production of

vitellogenin in sheepshead minnows exposed to

methoxychlor. They noted that while the parent

compound is not estrogenic, the demethylated

metabolites are active and bind to the estrogen

receptor. Exposure of male sheepshead minnows to a

methoxychlor concentration of 5.6 and 12.1 μg/L

resulted in fairly rapid increases in plasma

vitellogenin. By day 35, fish exposed to 12.1 μg/L

produced a substantial amount of vitellogenin,

approximately 120 mg vitellogenin/ml of plasma.

Sukumar and Karpagaganapathy (1992) investigated 

the effects of sublethal concentrations of the carbamate insecticide carbofuran on the dwarf gourami, 

Colisa lalia. Female fish were exposed to a concentration of 0.7 μg/L technical grade carbofuran in a static 

renewal exposure for a period of 20 days. The fish 

exposed to the carbofuran had fewer mature oocytes, 

and the mature oocytes that were observed appeared 

atretic (Sukumar and Karpagaganapathy,1992). 

Ghosh and Thomas (1995) assessed the antagonistic 

effects of the insecticide kepone (chlordecone) on 

oocyte maturation in vitro in Atlantic croaker. The 

breakdown of the germinal vesicle in Atlantic croaker 

is controlled by the maturation-inducing hormone 

17α,20β,21−

trihydroxy-4­

pregnen-3-one 

(20β-S). Ghosh 

and Thomas 

(1995) exposed 

fully grown and 

primed oocytes 

to either 20β-S, or 

20β-S and 

kepone. Kepone 

concentrations 

ranged from 100 

nM to 100 μM. 

Exposure of the 

oocytes for as 

little as one 

minute to 0.1 μM 

kepone resulted 

in a significantly 

lower breakdown 

of the germinal 

vesicles when 

compared to controls. Exposure of the oocytes to 

kepone followed by washing and then incubation 

with 20β-S resulted in a normal germinal vesicle 

breakdown process, leading the authors to conclude 

that kepone competes with the hormone for the 

receptor, rather than simply being toxic to the 

oocytes. 

Celius et al. (1999) investigated the production of 

vitellogenin and zona radiata (eggshell) proteins in 

primary hepatocytes of Atlantic salmon exposed to a 

variety of suspected EDCs including o,p’-DDT and 

lindane (γ−HCH). Cells treated with 1, 5, and 10 μM 

DDT and lindane induced vitellogenin and zona 

radiata proteins in an approximate dose-dependant 

manner. 

Certain pesticides or compounds used in pesticide formulations are 

suspected of being endocrine disrupters. Image courtesy of NOAA/ 

NOS. 

19
Laboratory Studies 

Cheek et al. (2001) reported on reproductive impairment in medaka by o,p’-DDT. Fish were exposed for 

either 2 or 8 weeks post hatch to nominal concentrations ranging from 0 to 7.5 μg/L in intermittent flow 

through chambers. Fish were then sampled at 2, 4, or 

8 weeks for vitellogenin content and gonad development. The remaining fish were then transferred to 

clean water and grown to sexual maturity and placed 

in mating pairs. Measurements of fecundity, fertility 

and hatching success were made on the eggs collected. Cheek et al. (2001) found that o,p’-DDT had 

no effect on vitellogenin expression after a 2 week 

exposure, but all doses resulted in vitellogenin 

expression after 8 weeks. In both exposures, o,p’­

DDT resulted in a female-skewed sex ratio at the two 

highest doses. Two fish exposed for 8 weeks at 1.94 

μg/L exhibited ovotestes. Both the 2 and 8 week 

exposures significantly affected fertility and in 

particular, hatching success at all concentrations. 

Because vitellogenin expression was not as sensitive 

an indicator as the other parameters measured, 

Cheek et al. (2001) concluded that the absence of 

vitellogenin expression should not be used to indicate a lack of reproductive endocrine disruption. 

Khan and Thomas (1998) also evaluated the effect of 

o,p’-DDT, however, looking at the release of gonadotropin hormone (GTH) in Atlantic croaker. Fish were 

exposed to o,p’-DDT through the diet (0.02 and 0.1 μg 

DDT/g body weight per day) for 3 and 7 weeks 

during gonadal recrudescence. The o,p’-DDT resulted in a significant increase in plasma GTH after 

both the 3 and 7 week exposures. The increase in 

GTH was accompanied by a slight increase in the 

relative size of the ovary as indicated by an increase 

in the GSI (Khan and Thomas, 1998). 

Knudson and Pottinger (1999) assessed the binding 

affinity of the pesticides toxaphene and dieldrin to 

both the E2

 receptor (ER) and the testosterone receptor (TR) in rainbow trout, using a competitive 

binding assay with 1 pmol radiolabeled E2

 or testosterone. Concentrations of toxaphene and dieldrin 

up to 50,000 times that of E2

 or testosterone failed to 

dislodge the endogenous hormones, indicating a low 

affinity of these two pesticides for rainbow trout ER 

and TR. 

Phthalate Esters. Certain phthalates appear to be 

weakly estrogenic in whole fish. Jobling et al. (1995) 

investigated the estrogenicity of a variety of phthalate esters including di(2-ethyhexyl) phthalate 

(DEHP), butylbenzyl phthalate (BBP), and di-nbutylphthalate (DBP), in an assay to assess the ability 

of the compounds to inhibit binding of E2

 to the ER in 

rainbow trout liver. DEHP showed the lowest 

affinity for trout ER, with an affinity of approximately 1:100,000 that of E2

 for ER. 

BBP showed an affinity of 1:10,000, however, DBP 

showed an affinity approaching 1:10 that of E2

. 

Jobling et al. (1995) noted that phthalates are the most 

ubiquitous man made contaminants in the environment, and that thousands of tons of plastics containing phthalates are disposed of annually in landfill 

sites, enabling these compounds to migrate into 

surface and groundwaters. Jobling et al. (1995) also 

stated that the presence of these compounds in the 

aqueous environment is well known in river, waste, 

and drinking water as well as in fish and sediments. 

Christiansen et al. (1998) injected immature rainbow 

trout with 500 mg/kg of either BBP or DBP. A preexposure blood sample was taken on Day 0, followed 

by a blood sample and sacrifice on Day 9. Vitellogenin was measured using a direct sandwich ELISA 

technique. BBP increased vitellogenin by a factor of 

approximately three, while the DBP did not increase 

the vitellogenin concentration above the detection 

limit (Christiansen et al., 1998). It is interesting to 

note here that while Christiansen et al. (1998) did not 

detect an increase in vitellogenin production in 

hepatocytes exposed to DBP, this same compound 

showed a fairly strong affinity for rainbow trout liver 

ER (Jobling et al., 1995), possibly indicating an 

antagonistic potential for DBP. 

Metals. In general, a number of trace elements have 

been shown to have a negative effect on endocrine 

function in fish. Ruby et al. (2000) investigated the 

effects of lead exposure on sexually maturing female 

rainbow trout. Fish exposed to 10 μg/L of waterborne lead (Pb(NO ) ) during the period of recrudes­ 3 2

cence had significantly lower GSIs and oocyte 

diameters than control fish on day 12. In addition, 

Ruby et al. (2000) assessed the effect of lead on the 

pituitary gland, the source of GTH. They found that 

the mean number of GTH producing granular 

pituitary basophils was lower in lead exposed 

females than in controls, suggesting the site of action 

of lead may be the pituitary. 

Thomas (1988) administered lead to Atlantic croaker 

in the diet (1.34 mg/70g fish/day). After 30 days, the 

GSI in the croaker was only 32 percent of controls. 

The exposure to lead also resulted in lower circulat20
Cadmium has been associated with elevated levels of 

E2 and enhanced gonadal growth in female Atlantic 

croaker (Micropogonias undulatus). Image courtesy 

of NOAA/NMFS. 

Laboratory Studies 

ing E2

 levels. However, in vitro experiments showed 

that the steroidogenic capacity of the ovarian tissue 

was not decreased by exposure to lead. 

Thomas (1989) looked at the effects of cadmium on 

Atlantic croaker. Females at an early stage of gonadal recrudescence were exposed to cadmium for 30

days in 30‰ seawater at a concentration of 1 mg/L. 

In contrast to lead, cadmium caused an accelerated 

rate of ovarian growth when compared to controls. 

In addition, croaker exposed to cadmium had 

elevated levels of plasma E2

, leading Thomas (1989) 

to suggest that vitellogenesis was enhanced by 

exposure to cadmium. 

In vitro experiments 

revealed increased GTH 

secretion by the 

hemipituitaries in 

cadmium treated fish. 

Ricard et al. (1998) 

investigated the effects 

of CdCl2 in rainbow

trout. Adults were 

exposed to 10 and 25 μg 

Cd/L. Exposure to 

cadmium tended to 

decrease (though not 

significantly), concentrations of the thyroid 

hormones T (triiodothyronine) and T4

 (thyroxine), which are believed to be 

involved in the adaptation of fish to environmental 

changes (e.g., temperature or salinity). In addition, 

exposure significantly increased concentrations of 

cortisol, a hormone involved in stress response in 

trout and other fish. 

β-Sitosterol. Pulp and paper mill effluents can impact 

reproductive development in fish (Tremblay and Van 

Der Kraak, 1998; McMaster et al., 1991). β-Sitosterol 

is one of the major plant sterols discharged from pulp 

and paper mills. MacLatchy et al. (1997) exposed 

goldfish to 75, 300, 600, and 1,200 μg/L of β-sitosterol, typical of concentrations from bleached kraft 

mill effluent. The experiment was run for 12 days 

using a static system renewed every two days. In 

general, plasma reproductive steroid levels in both 

males and females were decreased as a result of the 

treatment. MacLatchy et al. (1997) noted that gonadal 

cholesterol levels in fish treated with the highest level 

of β-sitosterol were also decreased, while gonadal 

cholesterol levels were unchanged in E2

 treated fish. 

This led the authors to suggest that decreases in 

plasma steroids could be related to the availability of 

cholesterol to the P450 enzymes (e.g., P-450scc and P45017α) involved in the conversion of cholesterol to 

the androgens and estrogens. 

Tremblay and Van Der Kraak (1998) investigated the 

effects of β-sitosterol on rainbow trout. β-Sitosterol 

induced the production of vitellogenin in immature 

rainbow trout during a 3 week exposure, although 

the induction appeared to be somewhat lower than 

that produced by 4-nonylphenol. At higher levels of 

β-sitosterol (75 and 100 μg/L), the levels of plasma 

testosterone fell below the detection limit of the 

radioimmunoassay 

(Tremblay and Van Der 

Kraak, 1998), similar to 

what McLatchy et al. 

(1997) found in goldfish. 

Synthetic Estrogens. As 

noted earlier, the synthetic estrogen EE2

 used

in oral contraceptives can 

also have a significant 

impact on the endocrine 

system in fish at very low 

concentrations. Scholz 

and Gutzeit (2000) 

investigated long term 

effects of EE2

 exposure to 

Japanese medaka. 

Freshly hatched males (XY) were exposed in a semistatic system to 100 ng/L EE2

 for 2 months, followed

by a 6 week recovery period. At this concentration, 

all medaka were sex reversed, having developed an 

ovary. In females (XX), a reduced production of eggs 

was noted and was reflected in the GSI at 10 and 100 

ng/L. At a 10 ng/L concentration, no alterations of 

testicular structure or function was detected. 

Lange et al. (2001) assessed the effects of EE2

 on the

life cycle of fathead minnows. Fish were exposed in 

a continuous flow-through system to concentrations 

of 0.2, 1.0, 4.0, 16, and 64 ng/L (nominal concentrations) for 305 days. Male fish exposed to ≥ 4.0 ng/L 

EE2

 failed to develop normal secondary sexual

characteristics. Histological analysis at 56 days 

posthatch revealed the female/male ratio in the 0.2 

and 1.0 ng/L exposures to be approximately 50:50. 

In the 4.0 ng/L exposure, the sex ratio was 84:5, with 

ovotestes contained in 11 percent of the fish. In fish 

exposed to 4.0 ng/L for 152 days, no testicular tissue 

was observed at all. Lange et al. (2001) calculated an 

overall no observed adverse effect EE2 concentration

of 1.0 ng/L. 

3

21
Freshwater Field Studies 

Metcalfe et al. (2001) assessed the histologic effects of 

EE2 on Japanese medaka. Fish were exposed in a 

static renewal system for 85-110 days post hatch. 

Experiments were begun one day after hatch, and 

fish were exposed in a static renewal system to 0, 0.1, 

1.0, 10, 100, and 1,000 ng/L EE2 (nominal concentration). The highest concentration was lethal to most of 

the fish, but those that did survive contained eosinophilic fluid in the organs and body cavity. In the 100 

ng/L exposure, 91 percent of the fish were identified 

as female, significantly different from the controls. 

Ovotestes were observed in males from all treatments 

except the 100 ng/L.

FRESHWATER FIELD STUDIES 

The next two sections contain summaries of some of 

the endocrine disrupter-related field work completed 

on fish in fresh and saltwater systems. Much of the 

work to date on environmental endocrine disruption 

has occurred in Europe. Studies reviewed in both 

sections fell into three broad categories and included 

investigations adjacent 

to STPs, pulp and paper 

mills, and general 

surveys that covered a 

defined geographic area 

rather than specific 

point source discharge 

types. Investigations 

occurring near these 

activties have produced 

some of the most 

significant findings. 

Table 6 contains a 

summary of results 

from a number of the 

freshwater investigations reviewed in this 

report. 

Sewage Treatment Plants. Some of the most dramatic examples of endocrine disruption in fish have 

been observed near the outfalls of STPs. All of the 

effects appear to be estrogenic. Recent work indicates that natural and synthetic estrogens are probably responsible for much of the estrogenicity in STP 

effluents and effects seen in fish, although the APEs 

may also be contributing to observed effects under 

certain influent characteristics. A number of studies 

that have attempted to characterize the estrogenic 

components of STP effluents are also included in this 

section. 

United Kingdom. Part of the impetus for research into 

endocrine disruption in fish, can be traced to casual 

observations by anglers in the mid-1980s of her

maphrodite roach (Rutilus rutilus) inhabiting STP 

lagoons (Routledge et al., 1998; Purdom et al., 1994). 

In one of the earlier studies of the association between STP effluents and estrogenic effects, Purdom et 

al. (1994) carried out a number of field exposures 

with rainbow trout (Oncorhynchus mykiss) at 15 STPs 

throughout England and Wales. Controls consisted 

of maintaining fish in well or spring water. Male and 

female fish were exposed in or near the effluent 

discharges from 1 to 3 weeks. The original hypothesis was that the synthetic estrogen EE2

 found in oral

contraceptives contributed to the estrogenic effect of 

the STP effluents. 

At each site, Purdom et al. (1994) found clear evidence of increased plasma vitellogenin in all fish, 

both male and female. Vitellogenin in 

male fish at the control sites ranged from 0.05-1.80 

μg/ml, and in females from 4.5-88.3 μg/ml. In males 

exposed to effluents from STPs, plasma vitellogenin 

values ranged from 23­

147,000 μg/ml, and in 

females from 470­

112,000 μg/ml. In most 

cases, the effect of the 

effluent was so obvious 

that tests of statistical 

significance were 

deemed inappropriate 

(Purdom et al., 1994). 

From these results, the 

authors concluded there 

was a nationwide 

distribution of estrogens 

in STP effluents. 

Although one of the 

goals of the research had 

been to determine whether EE2

 contributed to the

effect, attempts to measure levels in the effluents 

using GC/MS as well as radioimmunoassay were 

unsuccessful. Purdom et al. (1994), however, felt that 

EE2, and possibly the APEs were likely candidates for 

the observed effects. 

Sumpter and Jobling (1995) also conducted a nationwide survey in England and Wales, exposing caged 

male rainbow trout to effluents from 28 STPs cover

ing all 10 Water Authority areas, in order to deter

mine the extent of estrogenic effects. They noted that 

many of the compounds identified as endocrine 

­

­

­

The discovery of hermaphrodite specimens of the roach 

Rutilus rutilus in rivers close to sewage outfalls in the U.K. 

helped spur investigations of endocrine disruption in the 

U.K. Image courtesy of FishBase.org. 

22
10 6 

10 5 

C1 1 2 3 4 C2 5 6 7 8 C3 9 10 C4 11 12 C5 13 14 15 

Treatment 

Vitellogenin, ug/ml 10 4 

10 3 

10 2 

10 1 

1 

10 -1 

10 -2 

Figure 10. Results of a survey to assess estrogenicity of effluents 

from sewage treatment plants to rainbow trout in England and 

Wales. Numbers represent the 15 STPs where fish survived; C1-

C5 were control sites. Mean ±SEM. From Sumpter and Jobling, 

1995. Reprinted by permission. 

Freshwater Field Studies 

disrupters inevitably find their way into rivers via 

STPs. At 13 of the 28 sites, trout were not able to 

survive the 3 week experiment. At the 15 sites wher

fish did survive, there was a pronounced increase in

plasma vitellogenin in all fish, compared to controls

(Figure 10). Variability in the response ranged from

500 to over 50,000 fold at the different sites, which 

they attributed 

to a number of 

factors including 

the composition 

of the effluent 

(i.e., which 

compounds were 

present and at 

what concentrations), water 

temperature, and 

the age of the 

fish. 

Sumpter and 

Jobling (1995) 

concluded that 

effluents from al

STPs in the U.K. 

were probably 

estrogenic to 

fish. They also 

noted it was 

possible that 

some synthetic 

estrogens, 

particularly EE2,

may be a contaminant in the 

aquatic environment. 

e 

l 

Harries et al. (1996) completed a series of experiments 

in the U.K. with male rainbow trout placed in cages 

at the point of discharge and varying distances 

downstream of five sewage treatment outfalls on the 

River Lea, which flows past the northeastern portion 

of London. During the summer months, effluent 

from STPs comprise as much as 82 percent of the 

total flow of the upper portion of the river (Harries et 

al., 1996). After 3 weeks at the majority of sites, fish 

held up to 15 kilometers downstream of the inputs 

showed an increase in plasma vitellogenin, with 

statistically significant elevations up to 4.5 kilometers 

downstream. 

Levels of vitellogenin in the controls (reservoir and 

laboratory) were approximately 0.01 μg vitellogenin/ 

ml plasma. Levels of vitellogenin in male fish 

exposed to River Lea water ranged from approximately 0.1 to 1,000 μg/ml vitellogenin/ml plasma 

during the July/August timeframe. Field exposures 

repeated during November resulted in lower re

sponses. Harries et al. (1996) attributed this to an 

overall increase in dilution resulting from additional 

rainfall during that time of year. They also concluded that the 

elevated levels of 

vitellogenin were the 

result of treated 

sewage effluent in 

the river, as there 

was a general

pattern of decline in 

vitellogenin levels 

with distance 

downstream from 

each outfall.

Although the 

compounds responsible for induction 

were not identified, 

the authors speculated that the 

alkylphenols or the 

parent APEs, along 

with the synthetic 

estrogen EE2

 might

have contributed to 

the observed estrogenicity. Harries et 

al. (1996) noted, 

however, that no 

gross abnormalities have been reported in the 

population structure of rainbow trout in the River 

Lea, but that the stocking of fish in the river would 

tend to reduce the importance of natural reproduction. 

In another study by Harries (Harries et al., 1997), an 

assessment was made of vitellogenesis in caged male 

rainbow trout placed at various distances downstream of STPs in five other rivers in England (Great 

Stour, Arun, Chelmur, Stour, and Aire). In four cases 

where the cages were placed directly in the effluent, 

there were rapid and very significant increases in 

plasma vitellogenin in all fish, ranging from 0.03 to 

almost 52 μg vitellogenin/ml plasma, demonstrating 

the effluent was estrogenic. Initial vitellogenin levels 

in the males were around 0.01 μg/ml. The GSIs of 

the exposed fish were also significantly reduced. 

­

23
Freshwater Field Studies

24

Table 6. Field studies of endocrine disruption in freshwater species of fish.

Fish/Species Sex Land Use Location/ Study Site Caged/wild fish Vtg I T E2 GSI GD EM Possible Cause(s) Reference

Roach 

(Rutilus rutilus) M/F Sewage treatment plants 15 sites in England caged; 1-3 weeks + 17α-ethinylestradiol Purdom et al. , 1994

Rainbow trout 

(Oncorhynchus mykiss) M Sewage treatment plants

15 sites in 

England and 

Wales

caged; 

3 weeks + 17α-ethinylestradiol Sumpter and Jobling, 1995

Rainbow trout 

(Oncorhynchus mykiss) M Sewage treatment plants 5 sites on River Lea in U.K. caged; 3 weeks + 17α-ethinylestradiol and alkylphenols Harries et al., 1996

Rainbow trout 

(Oncorhynchus mykiss) M Sewage treatment plants 5 sites on River Lea in U.K. caged; 3 weeks + – Estrogens and nonylphenol Harries et al., 1997

Rainbow trout 

(Oncorhynchus mykiss) M Sewage treatment plants Berlin STP 6 months + Estrogens and nonylphenol Hansen et al. , 1998

Roach 

(Rutilus rutilus) M Sewage treatment plants 8 rivers in the U.K. wild + + – Estrogens and alkylphenols Jobling et al., 1998

Roach 

(Rutilus rutilus) M Sewage treatment plants Chelmsford STP, U.K. caged; 1 month + + Estrogens and alkylphenols Rodgers-Gray et al. , 2000

Rainbow trout 

(Oncorhynchus mykiss) M Sewage treatment plants River Aire, U.K. caged; 3 weeks + Alkylphenols from wool scouring plant Sheahan et al ., 2002a

Gudgeon 

(Gobio gobio) M Sewage treatment plants River Aire, U.K. wild + + Estrogens and alkylphenols van Aerle et al ., 2001

Rainbow trout 

(Oncorhynchus mykiss) M Sewage treatment plant Western Sweden caged; 2 weeks + Estrogens and alkylphenols Larsson et al., 1999

Mosquitofish 

(Gambusia a. holbrooki) M Sewage treatment plant

Hawkesbury 

River, 

Australia 

wild – Estrogens and alkylphenols Batty and Lim, 1999

Common carp 

(Cyprinus carpio) M Sewage treatment plants St. Paul, Minnesota wild + NE Estrogens Folmaret al., 1996

Walleye 

(Stizostedion vitreum) M Sewage treatment plants St. Paul, Minnesota wild + – + – Estrogens and alkylphenols Folmaret al., 2001

Walleye 

(Stizostedion vitreum) F Sewage treatment plants St. Paul, Minnesota wild + + – Estrogens and alkylphenols Folmaret al., 2001

M/F, male and female fish collected in study and effects same in both sexes; NE, no effect; Vtg, vitellogenin; I, intersex condition; T, testosterone; E2, 17 β-estradiol; 

GSI, gonadosomatic index; GD, gonadal development; EM effect on external male/female morphology
Freshwater Field Studies

25

Fish/Species Sex Land Use Location/ Study Site Caged/wild fish Vtg I T E2 GSI GD EM Reference

Largemouth bass 

(Micropterus salmoides 

floridanus)

M Pulp and paper mill St. Johns River wild – β-Sitosterol Sepúlvedaet al., 2002

Largemouth bass 

(Micropterus salmoides 

floridanus)

F Pulp and paper mill St. Johns River wild – – β-Sitosterol Sepúlvedaet al., 2002

Whitefish 

(Coregonus lavaretus) M/F Pulp and paper mill Lake Saimaa, Finland caged; 1 month + β-Sitosterol Soimasuoet al., 1998

Perch 

(Perca fluviatilis L.) F Pulp and paper mill Lake Saimaa, Finland wild – β-Sitosterol Karels et al. , 2001

White sucker 

(Catostomus commerson M/F Pulp and paper mill St. Maurice River, Quebec wild – – β-Sitosterol Gagnon et al., 1994

White sucker 

(Catostomus commerson M/F Pulp and paper mill Lake Superior, Canada wild – – – β-Sitosterol McMaster et al., 1991

Mosquitofish 

(Gambusia holbrooki) F Pulp and paper mill St. John's River, Florida wild – β-Sitosterol Bortone and Cody, 1999

Mosquitofish 

(Gambusia holbrooki) F Pulp and paper mill Perdido Bay, Florida wild – β-Sitosterol Cody and Bortone, 1997

Mosquitofish 

(Gambusia holbrooki) F Pulp and paper mill Fenholloway River, Florida wild – β-Sitosterol Jenkinset al. , 2001

Channel catfish 

(Ictalurus punctatus) F Sugar beet processing plant Red River, North Dakota wild – β-Sitosterol Hegrenes, 1999

Largemouth bass 

(Micropterus salmoides) M/F Power plant, chemical manufacture Escambia/Blackwater River, Florida wild + – Contaminants Orlando et al., 1999

Common carp 

(Cyprinus carpio) M General survey Various locations in U.S. wild +/NE –/NE –/NE Contaminants Goodbred et al. , 1997

Common carp 

(Cyprinus carpio) F General survey Various locations in U.S. wild –/NE Contaminants Goodbred et al. , 1997

Fathead minnow 

(Pimphales promelas) M/F General survey Mississippi River wild NE None Parks et al., 1997

M/F, male and female fish collected in study and effects same in both sexes; NE, no effect; Vtg, vitellogenin; I, intersex condition; T, testosterone; E2, 17 β-estradiol; 

GSI, gonadosomatic index; GD, gonadal development; EM effect on external male/female morphology

Table 6. Field studies of endocrine disruption in freshwater species of fish (continued).

Possible Cause(s)
Freshwater Field Studies 

On the River Aire, the effluent was extremely estrogenic. Vitellogenin levels at the sites downstream of 

the Marley STP in Keighley on the River Aire ranged 

from approximately 1 - 10,000 μg vitellogenin/ml 

plasma. Harries et al. (1997) noted there are a number of wool scouring mills in the basin that use 

alkylphenol polyethoxylate detergents to wash the 

fleece. An 

analysis of water 

samples near the 

Marley STP 

confirmed the 

presence of 4­

nonylphenol, 

with concentrations as high as 

330 μg/L, compared with 

concentrations of 

2.9 and 0.9 μg/L 

near other STPs. 

Harries et al. 

(1997) remarked 

that the estrogenicity at other 

sites could be due

to additional 

contaminants 

such as dieldrin, 

lindane, and 

phthalates. 

Sheahan et al. 

(2002a) conducted 

a followup study 

at the Marley STP 

in Keighley in an effort to identify the estrogenic 

contaminants in the effluent. High performance 

liquid chromatography fractionation paired with 

recombinant yeast assays revealed the estrogenic 

activity of the effluent was primarily due to the 

alkylphenolic compounds 4-nonylphenol, 

nonylphenol-1-ethoxylate, and nonylphenol-2­

ethoxylate, along with the natural estrogens E2

 and

estrone. Studies with caged male rainbow trout (1 

year old) exposed for 3 weeks to the effluent resulted 

in vitellogenin levels as high as 10,000 μg/ml. 

Sheahan et al. (2002a) noted that the high use of APEs 

by the wool scouring mills which discharge to the 

STP created an unusual situation, one in which the 

alkylphenolic chemicals may potentially have been 

responsible for up to 90 percent of the vitellogenin 

induction observed in caged male trout held in the 

final effluent discharge from the STP. Sheahan et al. 

(2002a) noted, however, that given the differential 

uptake, metabolism and bioconcentration of the 

alkylphenolic chemicals versus the estrogens in fish, 

it was very difficult to assess which of these two 

groups contributed most to the effects seen in the 

fish. 

Efforts have been underway since the early 1990s at 

the Marley STP in Keighley to tighten the discharge 

limits in an effort to 

improve downstream water quality, 

including reducing 

the discharge of 

APEs. Sheahan et al. 

(2002b) monitored 

the concentration of 

the alkylphenolic 

chemicals in the 

effluent as well as 

the effect on caged 

rainbow trout from 

1994 to 1997. Reductions in the concentration of the APEs 

as a result of improved treatment 

processes, and a 

subsequent reduction in vitellogenin 

production and an 

increase in the GSI in 

caged male rainbow 

trout has been 

observed. Concentrations are likely to 

continue to decrease 

as the major wool scouring mill that discharged to 

the STP has since relocated to another basin (Sheahan 

et al., 2002b). 

Jobling et al. (1998) demonstrated the widespread 

occurrence (Figure 11) of intersex roach Rutilus 

rutilus, a cyprinid fish in U.K. rivers, and that it was 

associated with exposure to effluents from STPs. 

Wild populations of roach were sampled upstream 

and downstream of STPs in eight rivers (Wreake/ 

Eye, Ouse, Lea, Arun, Nene, Trent, Rea, and Aire), 

representing a range in water quality from very good 

to poor (Jobling et al., 1998). Upstream sites were 

separated from downstream sites by one or more 

physical barriers, and upstream sites were usually 

separated from downstream sites by several kilometers. Upstream sites were still subject to exposure 

from sewage effluents from smaller treatment plants 

even further upstream. Because no riverine sites 

Figure 11. Gradations of ovotestes discovered in roach fish 

(Rutilus rutilus) in the U.K. The degree of intersexuality ranged 

from normal testes to testes with both spermatogenic tissue (t) 

and occasional primary oocytes (po) to gonads that contained 

primary and secondary oocytes (so) and an ovarian cavity (oc). 

From Jobling et al. (1998). Reprinted by permission. 

26
% Intersex

100 

75 

50 

25 

0 

Controls Upstream Downstream 

Sites 

** 

Figure 12. Comparison of the overall 

incidence of intersex roach in control, 

upstream and downstream sites. From 

Jobling et al., 1998. Reprinted by permission. 

Freshwater Field Studies 

could be located that were not impacted by STPs, 

controls consisted of lakes and canals. 

Although the level of treatment at the various STPs 

was not provided, Jobling et al. (1998) characterized 

the fish capture sites in terms of the strength of the 

influent or load coming into each plant, and the 

diluting ability of the receiving waters. The level of 

vitellogenin in male and female fish were character

ized using a radioimmunoassay, and the gonads of 

the fish were examined histologically for gonadal 

anomalies. 

When examined externally, all fish collected appeared to be either male or 

female. Histological examination of the fish, however, 

revealed a large proportion of 

the males examined were in fact 

intersex. Specifically, intersex 

males were characterized by the 

presence of ovotestes and/or an 

ovarian cavity in the testicular 

tissue (Figure 11). 

The incidence of intersexuality 

in males (Figure 12) ranged 

from 16 to 100 percent at sites 

downstream of the STPs, and 

from 12 to 44 percent at upstream sites. In some males, the 

degree of intersexuality was 

slight, while in some fish, more 

than 50 percent of the gonadal 

tissue was ovarian. The degree 

of intersexuality was also 

significant at downstream sites. Intersexuality at the 

control sites ranged from 4 to 18 percent. Jobling et 

al. (1998) noted that while intersexuality is considered rare in roach, rates of up to 5 percent have been 

observed in the common carp (Cyprinus carpio), a 

cyprinid fish related to the roach. The authors stated,

however, that even though a low level of intersexuality in roach may be “natural”, the results of the 

survey suggested the incidence of intersexuality in 

U.K. rivers was considerably higher than what might

be expected, and that the condition was strongly 

associated with discharges from STPs. 

An evaluation of plasma vitellogenin in male roach 

also provided evidence that some populations were 

exposed to estrogenic contaminants. Jobling et al. 

(1998) noted that the concentration of vitellogenin 

found in intersex fish was intermediate between 

­

female roach and males (Figure 13). There was also a 

positive correlation between the degree of intersexuality and plasma vitellogenin concentrations, and a 

negative correlation between vitellogenin levels in 

males and the GSI. A second species of fish, the 

gudgeon (Gobio gobio) from several rivers exhibited 

similar anomalies, indicating the effects were unlikely to be species specific. 

Jobling et al. (1998) discussed possible causes for the 

observed effects. There was a significant correlation 

between the proportion of intersex fish and the 

strength of the effluent from the STPs, leading the 

authors to conclude that estrogenic constituents of 

the sewage effluents were 

responsible for the occur

rence of intersexuality in 

the fish. While the authors 

did not investigate the 

chemicals responsible, they 

noted that natural (E2

 and

estrone) and synthetic (EE2)

estrogens have been re

ported in STP effluents, 

including some of the 

effluents that impact the 

sites sampled in the study. 

In addition, they noted that 

the concentrations of 

estrogens reported in STP 

effluents were high enough 

to cause vitellogenin 

synthesis in males in 

laboratory studies. They 

cautioned, however, that 

other chemicals, namely the alkylphenolic compounds, are also present in many sewage effluents, 

and thus the role of these types of chemicals should 

not be underestimated. 

In a followup study, van Aerle et al. (2001) assessed 

the extent of reproductive endocrine disruption in 

gudgeon in the River Aire. Fish were collected at a 

series of locations near STPs, as well as control (lake) 

sites. Vitellogenin was measured using an ELISA. 

The GSI was calculated and the gonads were examined histologically for gonadal anomalies. Intersex 

gudgeon were found at all sites, with frequencies as 

high as 15 percent. Fish from Thwaite Weir on the 

River Aire had the greatest severity of intersexuality, 

with over 40 percent of the fish collected having 

greater than 50 percent ovarian tissue containing 

both primary and secondary oocytes interspersed 

within the testicular tissue. 

­

­

27
 3 Lab Control 10

Control 

Vitellogenin (ug/ml) 

Upstream ****** 

10 2 

Downstream 

*** 

10 1 

*** 

1 *** 

10-1

10-2

Male Intersex Female 

Figure 13. Plasma vitellogenin in roach 

from various rivers, upstream and 

downstream of STP effluents. Asterisks 

denote differences at 0.001 level of 

significance from field control sites. 

From Jobling et al. (1998). Reprinted by 

permission. 

Freshwater Field Studies 

Interestingly, while no intersex fish were found at a 

site downstream from the Marley STP at Keighley, 97 

percent of the gudgeon collected from that site were 

female. Although they noted that monosex populations may be natural for this species, because this 

phenomenon was not observed at any other locations

on the River Aire, van Aerle et al. (2001) stated they 

could not rule out the possibility that the almost 

exclusively female population of gudgeon was the 

result of complete sex reversal 

in males downstream of the 

Marley STP. At all locations, 

unlike with roach, there were 

no differences in the GSI 

between sites. Although 

intersex gudgeon were found 

at the sites sampled, the 

incidence was not as high as 

that found by Jobling et al. 

(1998) for roach. van Aerle et 

al. (2001) stated this was 

somewhat surprising as 

gudgeon are a more demersal 

species, frequently in contact 

with sediments, which would 

seem likely to increase their 

exposure to estrogenic compounds. As a possible explanation, they noted that most 

of the gudgeon collected were 

from areas with gravel beds as 

opposed to mud, and also that 

species likely differ in terms 

of their susceptibility to 

endocrine disruption. van Aerle et al. (2001) concluded, as did Jobling et al. (1998), that the occurence 

of intersex fish was not species specific. 

Desbrow et al. (1998) reported on the results of a 

chemical fractionation and screening technique to 

isolate estrogenic components from seven domestic 

STP effluents in the U.K., using a recombinant yeast 

assay. Three sterols were isolated from the estrogenic 

fractions of the sewage extracts, and were identified 

as the natural hormones E2

 and estrone, and the 

synthetic hormone EE2. Concentrations of E2 and

estrone ranged from 1 ng/L to almost 50 and 80 ng/ 

L, respectively. The concentration of EE2 was gener

ally below the detection limit but in three of the 

effluent samples was found at concentrations ranging 

from 0.2-7.0 ng/L. They suggested that natural and 

synthetic hormones might be responsible for the 

observed induction of vitellogenin in males placed 

downstream of discharges from STPs that received 

mainly domestic wastes (Desbrow et al., 1998). 

In a follow-up paper, Routledge et al. (1998) conducted a series of in-vivo experiments with adult 

male rainbow trout and adult roach for 21 days usi

similar, environmentally relevant concentrations of 

E2

, and estrone (1, 10, and 100 ng/L). They found 

that both male rainbow trout and roach were very 

sensitive to these concentrations of E2

 and estrone, producing high levels of plasma 

vitellogenin. The threshold E2 

concentration in the trout was 

estimated to be between 1 and 

10 ng/L. Male rainbow trout 

seemed more sensitive to E2 

than male roach, which they 

suggested could be due to the 

smaller egg size and subsequently lower vitellogenin 

requirements of the roach. 

They also noted that the 

synthetic estrogen EE2 is not

only a potent estrogen but is 

also persistent due to the 17α

ethinyl group which acts to 

reduce its rate of metabolism, 

and that when present could be 

a major contributor to any 

estrogenic response. Routledge 

et al. (1998) concluded that the 

steroidal estrogens identified in 

the domestic sewage effluent 

were present in sufficient amounts to cause the 

synthesis of vitellogenin in vivo. 

Rodgers-Gray et al. (2000) investigated the estrogenicity of effluent from the Chelmsford STP in the 

U.K., using a toxicity identification and evaluation 

(TIE) method. Estrogenic fractions of the effluent 

from a particular STP were identified using a estrogen-responsive recombinant yeast screen followed by 

identification of the estrogenic components using gas 

chromatography/mass spectrometry. Rodgers-Gray 

et al. (2000) were able to show that the natural 

estrogens E2

 and estrone were the major contributors 

to the estrogenic activity of the effluent. The synthetic hormone EE2

 was also detected occasionally

along with 4-nonylphenol and nonylphenol 

ethoxylates, although the APEs were not identified 

as contributing to the estrogenicity of the samples of 

STP effluent. 

ng 

­

­

28
Freshwater Field Studies 

Male roach were also exposed to various concentrations of the effluent. After only 1 month in the 100 

percent effluent experiment, the concentration of 

plasma vitellogenin in exposed male roach increased 

to 109 μg/ml. The concentrations of E2 (7-88 ng/ml) 

and estrone (15-220 ng/ml) measured in the effluent 

at the time of the exposures suggested to RodgersGray et al. (2000) that the natural steroids alone were 

probably capable of producing the observed induction of vitellogenin. In a little over 1 percent of the 

males exposed to the effluent, primary oocytes were 

occasionally detected in testicular tissue, although 

the intersexuality was not correlated with effluent 

concentration or duration of exposure. 

Western Sweden. Larsson et al. (1999) investigated the 

estrogenicity of an effluent from a small STP in 

western Sweden receiving primarily domestic 

wastewater. They exposed juvenile rainbow trout in 

cages downstream of the STP for a period of 2 weeks 

and found a very significant induction of vitellogenin 

(1.5 mg/ml) in all eight exposed fish. 

Gas chromatography/mass spectrometry revealed 

the presence of the synthetic estrogen EE (4.5 ng/L), 2

the natural estrogens estrone (5.8 ng/L), and E2 (1.1 

ng/L), along with 4-nonylphenol (840 ng/L) and 

bisphenol-A (490 ng/L). They noted that because the 

estrogenicity of 4-nonylphenol and bisphenol-A is 

typically 10-3 to 10-4 that of natural estrogens, the 

relative importance of these phenolic compounds in 

the effluent was probably minor. On the other hand, 

the level of EE exceeded levels shown to be estro­ 2 

genic in fish by 45 times. They concluded that levels 

of EE found in the effluent were sufficient to explain 2

a major part of the observed estrogenicity, and that 

this compound poses a potential hazard to fish and 

possibly other aquatic organisms. 

Germany. Hansen et al. (1998) investigated the 

estrogenicity of effluents from a Berlin STP to male 

rainbow trout exposed to varying concentrations of 

effluent (10, 20, 30, and 40 percent), and found there 

was a marked increase in vitellogenin in males 

exposed to 20 percent and greater STP effluent. 

In southwestern Germany, Korner et al. (2001) 

analyzed effluents from 16 municipal STPs in the 

state of Baden-Württemberg using the E-screen 

(MCF-7) assay to detect estrogenic activity. Cell 

number was assessd by the measurement of total 

protein using sulforhodamine B and extinction at 550 

nm. Effluents from all the STPs contained concentration dependent estogenic activity or EEQ in the range 

of 0.2 - 7.8 ng/L, with the natural estrogens E2

, EE2, 

and, to a lesser extent, estrone contributing 90 

percent or more of the E2 equivalent concentration. 

New South Wales, Australia. Batty and Lim (1999) 

investigated the effects of sewage effluent on the 

morphology of male mosquitofish (Gambusia a. 

holbrooki) in a river in New South Wales. In 

mosquitofish, the modified anal fin, or gonopodium, 

is formed under the influence of testosterone, and is 

critical for sperm transfer. Batty and Lim (1999) 

looked at differences in gonopodium length in fish 

inhabiting reference sites versus males inhabiting 

sites downstream of the STP. The gonopodium in 

males inhabiting the site downstream of the STP was 

significantly reduced in size compared to the reference site. They also established a site that received 

agricultural runoff (fertilizers and pesticides), but 

found no differences between the fish at this site 

versus the reference site. 

Minnesota/Mississippi Rivers. Folmar et al. (1996) 

investigated the levels of vitellogenin in common 

carp collected at five riverine locations around St. 

Paul, Minnesota. Vitellogenin was identified using 

Western blot analysis and quantified using a capture 

ELISA assay. 

Seven of 10 male fish collected from an effluent 

channel below the St. Paul, Minnesota metropolitan 

STP plant had detectable levels of vitellogenin; six 

males had vitellogenin approaching those of females 

captured at the same site (Folmar et al.,1996). Interestingly, males from the same location had plasma 

testosterone concentrations comparable to carp taken 

from the St. Croix River reference site. Folmar et al. 

(1996) also noted that male fish collected on the 

Minnesota River, a tributary that receives extensive 

agricultural runoff, did not show any induction of 

vitellogenin. As noted earlier, certain pesticides or 

components of pesticide formulations may be 

endocrine disrupters. 

In a followup study, Folmar et al. (2001) assessed 

altered sex steroid levels and vitellogenin in walleye 

(Stizostedion vitreum) at two sites near the same STP 

in St. Paul, Minnesota and at a reference site on the 

Snake River. All fish collected, both males and 

females, from the STP effluent channel contained 

measurable levels of vitellogenin. Vitellogenin in 

males ranged as high as 3.2 mg/ml near the STP; no 

vitellogenin was detected in males from the downstream or reference site. Males from the effluent 

channel as well as a downstream site also had 

29
Freshwater Field Studies 

significantly depressed levels of testosterone, and in 

the effluent channel, significantly elevated levels of 

E2. Females taken adjacent to the STP also had 

significantly elevated levels of E2

 as compared to 

those from the downstream or reference sites. 

Finally, spermatogenesis appeared somewhat reduced as a number of fish taken near the STP contained spermatogonia with no evidence of spermatocytes/spermatids. Folmar et al. (2001) noted, however, that there was no correlation between E2 

concentration and vitellogenin in males which they 

suggested might be due to the presence of other 

natural and synthetic estrogens, and possibly to 4­

nonylphenol and 4-nonyphenol ethoxylates. 

Lake Mead, Nevada. Snyder et al. (2001) used the EScreen with an estrogen receptor controlled luciferase 

report gene construct (MCF-7-luc) to identify and 

quantify estrogen receptor agonists in STP effluents 

at five sites in Lake Mead. Samples of effluent were 

fractionated (HPLC) to assess the contribution of 

PAHs, PCBs, and organochlorine pesticides, 

nonyl and octylphenols, or 

E and EE . Results of the 2 2

E-screen assays for the 

Lake Mead samples 

indicated that E and EE 2 2 

were the dominant environmental estrogens, 

accounting for over 99 

percent of the observed 

estrogencity in the 

samples. 

Pulp and Paper Mills. 

Studies adjacent to pulp and paper mill discharges in 

Europe and in North America have revealed a 

number of effects, including depressed levels of 

reproductive hormones, decreased GSIs in both male 

and female fish, and external masculinization of 

females. 

Lake Saimaa, Finland. Soimasuo et al. (1998) studied 

the effects of pulp and paper mill effluents on 

whitefish (Coregonus lavaretus) in Lake Saimaa in 

Southeast Finland. Immature hatchery-reared fish 

were exposed for 1 month at a series of sites on the 

lake. Interestingly, there were no significant changes 

in E2 or testosterone concentrations in juvenile 

whitefish at any of the sites. There was expression of 

the vitellogenin gene in the vicinity of the pulp and 

paper mill discharging the highest concentrations of 

wood derived compounds (including β-sitosterol). 

More recently, Karels et al. (2001) found lowered 

levels of E and testosterone in female and male 2

perch (Perca fluviatilis L.) in Lake Saimaa at sites 

downstream of two pulp and paper mills. Roach 

collected at the same sites, however, showed no 

significant differences in these steroid hormones. 

This may be an indication of differences in susceptibility between species. The GSI and fecundity in 

female perch was significantly lower at the sites 

adjacent to the pulp and paper mills, but similar in 

roach. 

St. Maurice River, Quebec. Gagnon et al. (1994) investigated the effects of bleached kraft mill effluents on 

white sucker (Catostomus commersoni) in the St. 

Maurice River which empties into the St. Lawrence 

River. The site was chosen as there are no other 

industries or towns within 100 kilometers of the mill, 

which Gagnon et al. (1994) felt offered a unique 

opportunity to study the effects of effluents from the 

pulp and paper mill. Fish were sampled at three 

sites, one 3 kilometers 

above the mill (reference 

site), a second 2 kilometers downstream, and a 

third 95 kilometers 

further downstream. 

Fish were captured both 

during and after the 

spawning periods. In 

females, levels of E2

 and 

testosterone were 

significantly lower at 

the two downstream 

sites compared to the 

reference site. 

In males, 11-ketotestosterone was significantly lower 

in fish at the two downstream sites. MacLatchy et al. 

(1997) found a similar reduction in plasma steroids in 

both males and females exposed in the laboratory to 

β-sitosterol, one of the major byproducts of the wood 

pulp delignification process (Cody and Bortone, 

1997). 

Although plasma steroid hormone levels in both 

male and female white sucker were depressed, the 

GSI was similar at all sites, and fecundity (number of 

eggs) in females was variable. The authors concluded that the parameters measured did not allow 

them to clearly relate pertubations in plasma steroid 

hormone levels to impaired reproduction as measured by gonad weight and fecundity. 

The common carp, Cyprinus carpio has been used in a 

number of endocrine disrupter field studies. Image 

courtesy of USDOI/FWS. 

30
Freshwater Field Studies 

Lake Superior, Canada. McMaster et al. (1991) investigated the effects of effluents from a bleached kraft 

pulp mill on white sucker. Male and female fish 

were collected from two areas on Lake Superior, 

spawning streams in Jackfish Bay, which is subject to 

paper mill effluents, and Mountain Bay, which 

receives no pulp mill effluents. Neither sampling 

area received municipal or industrial effluents. Fish 

of both sexes exposed to the pulp mill effluent 

exhibited significantly reduced GSIs, increased 

hepatosomatic indices or HSIs (liver weight/body 

weight X 100), and increased age to maturity. In 

addition, there were significant reductions in 11­

ketotestosterone in males and E in females. 2

St. John’s River, Florida. 

Bortone and Cody (1999) 

investigated the phenomenon of arrhenoidy (masculinization) in female 

mosquitofish (Gambusia 

holbrooki) from Rice Creek 

and Etonia Creek, tributaries 

of the St. Johns River. Fish 

were collected at locations 

upstream and downstream of 

a large paper mill on Rice 

Creek, as well as a reference 

site (Etonia Creek). The 

parameter investigated was 

the transformation of the 

anal fin of females into the 

gonopodium-like structures 

typical of developing males. 

A simple measurement of 

anal fin length (AL) divided 

by standard body length (SL) was used to assess the 

degree of arrhenoidy in females. Females collected at 

the site immediately downstream of the mill on Rice 

Creek had significantly larger AL/SL values compared to the Etonia Creek reference site, indicating a 

masculinization, or elongation of the anal fin in 

females. Sites further downstream showed what 

appeared to be a distance/dose related response to 

the effluent exposure. Although not abundant 

enough to allow between site comparisons, Bortone 

and Cody (1999) noted gonopodial development in 

females of two other species, the least killifish 

(Heterandria formosa) and the sailfin molly (Poecilia 

latipinna) in Rice Creek. 

Sepúlveda et al. (2002) evaluated biomarkers of 

reproductive function in Florida largemouth bass 

(Micropterus salmoides floridanus), at a number of sites 

in the St. Johns River, including one in close proximity to the paper mill on Rice Creek. Parameters 

measured included GSI, plasma vitellogenin, E2

, 11­

ketotestosterone, and a histological assessment of the 

gonads. Both male and female fish were collected. 

Female bass collected near the paper mill during 

spawning season had vitellogenin levels approximately an order of magnitude lower than those fish 

collected at the reference site. E levels were also 2

significantly lower. The GSI in both female and male 

largemouth bass during the spawning season were 

lower at the site near the pulp and paper mill compared to the reference site. Levels of 11­

ketotestosterone were also 

significantly lower in males 

collected near the paper mill. 

Sepúlveda et al. (2002) 

suggested that lower levels of 

vitellogenin in females, and 

lowered GSIs in females and 

males near the mill, may 

have been the result of 

altered levels of E and 11­ 2

ketotestosterone in these fish. 

Fish collected from a site 40 

kilometers downstream of 

the paper mill also exhibited 

evidence (significantly lower 

GSI and vitellogenin in 

females, significantly lower 

11-ketotestosterone in males) 

of endocrine disruption. 

Sepúlveda et al. (2002) 

suggested though that 

elevated levels of PAHs and 

PCBs in the tissues of the fish may have been responsible, and were a result of contamination associated 

with marine shipping near this site and not from the 

paper mill upstream. 

Perdido Bay, Florida. In the late 1970s, Howell et al. 

(1980) had reported on the presence of arrhenoidy in 

mosquitofish in Elevenmile Creek which receives 

pulp and paper mill effluent. More recently, Cody 

and Bortone (1997) completed a followup investigation in Elevenmile Creek after process changes had 

been made at the mill. These changes included the 

reclamation of 97 percent of tall oils and terpenes, byproducts of the pulping process. The phytosterol βsitosterol is a major component of the tall oils discharged. The reference or control site in this study 

was Eightmile Creek, which receives no paper mill 

effluent. 

Pulp and paper mills can be a source of endocrine 

disrupting compounds. Image courtesy of 

USDOI/NPS. 

31
Freshwater Field Studies 

Female fish collected at the two sites on Elevenmile 

Creek, both below the effluent discharge point for the 

mill, exhibited significant elongation of the anal fins, 

compared with the Eightmile Creek reference site. 

Cody and Bortone (1997) concluded that masculinization seen among female mosquitofish from 

Elevenmile Creek exposed to effluents from the pulp 

mill prior to the initiation of effluent treatment 

improvements had persisted. However, the lower 

frequency of highly masculinized females originally 

seen by Howell et al. (1980) suggested that an overall 

reduction in the degree of 

masculinization was associated with tall oil reclamation 

efforts. 

Fenholloway River, Florida. 

Jenkins et al. (2001) assessed 

arrhenoidy in female 

mosquitofish from the 

Fenholloway River, and also 

analyzed water samples in an 

effort to identify possible 

androgenic components. The 

sampling site on the 

Fenhollway River was 

approximately 4 kilometers 

downstream from the settling 

ponds of a large paper mill. 

All female mosquitofish taken from the Fenholloway 

River had significantly elongated anal fins as compared to the Spring Creek reference site (Jenkins et 

al., 2001). 

In the laboratory, water samples from the 

Fenholloway River were eluted through solid-phase 

extraction cartridges. Androgen agonist activity of 

the fractions was then tested using androgenreceptor transcription assays. Those fractions with 

androgen receptor agonist activity were analyzed 

using high performance liquid chromatographymass spectrometry. Jenkins et al. (2001) identified 

one of the androgenic components in the effluent as 

androstendione, a precursor in androgen formation. 

The authors noted that its presence could have been 

brought about as a result of the microbial modification of plant sterols, including β-sitosterol, or that the 

androstendione might be a natural component of the 

original wood. Jenkins et al. (2001) cautioned, 

however, that other unidentified compounds in the 

effluent could be acting as androgens or androgen 

precursors, or that compounds within the effluent, 

such as β-sitosterol, might also act to interfere with 

steroid biosynthesis or degradation. 

Red River, North Dakota. Although not adjacent to a 

pulp and paper mill, Hegrenes (1999) examined 

channel catfish (Ictalurus punctatus) for indications of 

arrhenoidy on the Red River near a sugar beet 

processing plant. Fish were identified as being either 

male or female, and then a subsample of the almost 

900 fish captured were examined internally to assess 

spawning readiness, and to confirm external sex 

identification. The internal evaluation revealed that 

2 percent of the females identified through dissection 

exhibited external male features. The masculinized 

female catfish were collected 

5 miles downstream of 

Fargo, North Dakota. 

Nearby point sources 

included the Moorhead 

sewage treatment ponds, as 

well as a sugar beet processing plant. Hegrenes (1999) 

noted that while the sewage 

effluent might be a logical 

source of estrogenic compounds, chemicals in the 

effluent associated with 

sugar beet processing such 

as β-sitosterol might be 

more likely to have induced 

the observed arrhenoidy. 

General Surveys. A number of general surveys, not 

tied to specific land use types were found for freshwater fish in North America. These types of studies 

are important as they provide an indicator of the 

general level of endocrine disruption in a larger 

geographic area. 

Atlantic Canada. Fairchild et al. (1999) used historical 

records of the aerial application of the insecticide 

Matacil 1.8D and catch data for the Atlantic salmon 

(Salmo salar) to study the potential effects of 4­

nonylphenol. Matacil 1.8D contains the carbamate 

insecticide aminocarb as the active ingredient and the 

EDC 4-nonylphenol as a primary solvent. 

Between 1975 and 1985, Matacil 1.8D was applied to 

forests in Atlantic Canada to control damage caused 

by the spruce budworm (Choristoneura fumiferana). 

The pesticide was applied around the time of salmon 

smoltification, the developmental process involving 

several hormonal systems that enable the salmon to 

adapt to life in seawater. The author used the ratio of 

aminocarb to 4-nonylphenol in the formulation to 

estimate the aqueous concentrations of 4-nonylphenol that occurred after application. 

Female eastern mosquitofish, Gambusia holbrooki 

have been affected by effluents from certain pulp 

and paper mills. Image courtesy of USDOI/ 

USGS - Chris Appleby. 

32
Freshwater Field Studies 

There is evidence that Atlantic salmon (Salmo salar) may have 

been impacted by 4-nonylphenol contained in a pesticide formulation used in the forests of New Brunswick and Newfoundland. 

Image courtesy of USDOI/FWS - William Hartley. 

In the Restigouche River basin in 1977, the authors 

found a significant negative relationship between 

salmon return and the proportion of tributaries 

sprayed with Matacil 1.8D. Fairchild et al. (1999) also 

found that for 16 rivers in the region, a significant 

proportion of the lowest salmon catches between 

1973 and 1990 coincided with application of the 

pesticide formulation. Significantly, no relationship 

was found between application and salmon catch for 

Matacil 1.8F, a flowable form of Matacil that does not 

contain 4-nonylphenol as a solvent. Following a 

large application of Matacil 1.8D in 1979, the catch 

per unit effort in 1984 of the 1979 year class was the 

lowest recorded for 

any catches between 1981 and 

1987. The authors 

also found a decline 

in recruitment and 

the spraying of 

Matacil 1.8D for a 

second species of 

fish, the blueback 

herring, (Alosa 

aestivalis) in New 

Brunswick. 

The authors hypothesized that the 4-nonylphenol contained in 

Matacil 1.8D was the causative agent in the decline of 

both the salmon and blueback herring. Fairchild et 

al. (1999) also stated that the estimated levels of 4­

nonylphenol present after forest spraying would fall 

into the same range as those currently found in 

discharges from STPs, industrial effluents and 

agricultural wastes. Although 4-nonylphenol is 

estrogenic, its possible effects on the thyroid, particularly to T and T might be useful here. 3 4

Escambia and Blackwater Rivers, Florida. Orlando et al. 

(1999) examined the reproductive physiology of 

largemouth bass (Micropterus salmoides) collected 

from the Escambia and Blackwater Rivers which 

empty into Pensacola Bay in northwest Florida. In 

Escambia Bay, fish were collected downstream from a 

coal fired electric utility and a nylon fiber and 

chemical intermediates manufacturing plant. The 

Blackwater River was used as a reference site. In this 

study, parameters measured included GSI and the 

HSI, plasma concentrations of E2 and testosterone, 

and vitellogenin. 

There were no differences in the GSIs for males or 

females between sites. While there was no difference 

in HSI for females between sites, males from the 

Escambia Bay site had larger HSIs relative to males 

from the Blackwater River. Levels of E did not differ 2

in females between sites, however, males from the 

Escambia River had significantly lower levels of 

testosterone than males from the Blackwater River. 

There was no difference in the number of females 

with detectable vitellogenin between sites. However, 

three out of 15 males from the Escambia River had 

detectable vitellogenin, while detectable vitellogenin 

levels were not found in any male fish from Blackwater River. 

National Survey (U.S.). 

Goodbred et al. (1997) 

completed a survey 

covering a number of 

regions in the U.S. 

(Figure 14), to assess 

the potential for 

contaminant-induced 

endocrine disruption 

in the common carp 

(Cyprinus carpio). The 

design of the project 

provided the opportunity to assess fish 

across a wide range of geographies and land use 

types. The goals of the study were to: 1) determine if 

endocrine disruption is widespread in the United 

States in areas inhabited by the carp, 2) evaluate any 

relationships between endocrine disruption and 

contaminants, and 3) aid in determining whether 

further studies are needed, as well as the type of 

investigation that would be most useful (Goodbred et 

al., 1997). Twenty-five sites (23 streams and two 

impoundments) representing a wide range of geographic areas, land use types and contaminant levels 

were selected. 

Between August and December 1994, during the 

period of gonadal maturation, approximately 650 

male and female adult carp were sampled. Parameters measured included E2

, 11-ketotestosterone, 

vitellogenin, and gonadal histopathology. Because 

most of the sites sampled are part of the National 

Water Quality Assessment (NAWQA) Program, 

contaminant levels were already characterized. Data 

on organochlorine pesticides and PCBs in tissues, 

phthalates, phenols and PAHs in bed sediments, and 

dissolved concentrations of pesticides, were evaluated for correlations with the reproductive health of 

carp. 

33
Freshwater Field Studies 

Goodbred et al. (1997) found no significant regional 

differences in steroid hormones for males, but steroid 

hormones in females from the Northern and Southern Midcontinent sampling region were significantly 

different from other regions of the country. Within 

all regions, there were significant differences in one 

or both hormones (E2 and 11-ketotestosterone) for 

male and female fish between sites. 

The lowest 

E2 concentration in 

females 

occurred in 

Sanger, 

Texas north 

of Dallas. 

Conversely, 

the highest 

mean 

concentration of E2

 in

both male 

and female 

fish 

occurred in 

a reservoir 

northeast 

of 

Modesto, 

California. 

The lowest 

mean 11

ketotestosterone concentration occurred in male carp 

from the Shenandoah River in West Virginia. 

Goodbred et al. (1997) also examined the ratio of E2/ 

11-ketotestosterone as a possible indicator of endocrine disruption. The lowest E2

 to 11-ketotestosterone

ratio in both male and female carp occurred on the 

Platte River in Nebraska, which also had the highest 

dissolved pesticide concentration. Goodbred et al. 

(1997) found a significant negative correlation in 

male carp between 11-ketotestosterone and bed 

sediment concentrations of phenols. There was also 

a significant negative correlation between this ratio 

for both male and female carp and the concentration 

of dissolved pesticides. 

There was substantial variability in the concentrations of E2

 and 11-ketotestosterone in individual male

and female fish, even within a site. At most sites, 

concentrations of these hormones were within a 

factor of five of each other. Goodbred et al. (1997) 

cautioned that using levels of sex steroid hormones 

in fish, even within the same period of the reproductive cycle, is difficult because of natural variability. 

Perhaps more importantly, they concluded that 

subtle effects of endocrine disrupters are unlikely to 

be detected using differences in sex steroid hor

mones, and that only the strongest influences of 

contaminants are likely to be evident. 

Vitellogenin in 

carp was assessed using a

capture ELISA 

technique and a

monoclonal

antibody specific

for carp vitellogenin. The

majority of male

plasma samples,

however, were 

negative for

vitellogenin.

Overall, approximately 16

percent of the 

male plasma

samples contained detectable

levels of vitellogenin. 

The detection frequency of vitellogenin in male carp 

was not significantly correlated with any of the 

contaminant groups. In addition, the detection 

frequency of vitellogenin in male carp was not 

significantly correlated with concentrations of either 

11-ketestosterone or E2

.

The highest frequency of vitellogenin detection at 

any site in male carp (approximately 53 percent) 

occurred in the Hudson River, New York, south of 

Lake Luzerne. Interestingly, Goodbred et al. (1997) 

reported that land use in this basin is primarily forest 

(91 percent). 

Thirty percent of male fish caught in the Red River 

site near Drayton, North Dakota had detectable 

levels of vitellogenin, and on the White River at 

Hazleton, Indiana, vitellogenin was detected in 26 

percent of male carp. The highest amount of vitellogenin detected in male carp at any site (0.83 mg/ml) 

­

Figure 14. Regions and sampling sites in the USGS reconnaissance study. 

From Goodbred et al. (1997). Reprinted by permission. 

­

34
Estuarine/Marine Studies 

also occurred at the White River site. Earlier, 

Sumpter and Jobling (1995) found approximately 20 

mg vitellogenin/ml of plasma in male rainbow trout 

downstream of STPs in the U.K. Goodbred et al. 

(1997) noted that levels of vitellogenin found in 

males in their study was far below the normal range 

found in females. They also noted that vitellogenin 

was detected in males at several minimally contaminated sites with no STP effluent sources, which might

indicate that some male fish have low background 

levels of vitellogenin during a portion of the reproductive cycle. 

Gonadal anomalies in the carp were also investigated. Out of 438 male and female carp examined, 

only one abnormality was found. This was the 

presence of a few basophilic cells found in a male 

carp from the Shenandoah River, which Goodbred et 

al. (1997) concluded were probably primary oocytes. 

This fish, however, contained no detectable vitellogenin in the plasma. This single abnormality can be 

compared with the nearly 25 percent of roach being 

classified as intersex above the STPs, and nearly 60 

percent downstream, on eight rivers in the U.K 

(Jobling et al., 1998). In addition, rates on intersexuality of up to 5 percent have been observed in carp 

(Jobling et al., 1998). 

Mississippi River. Parks et al. (1997) carried out a field 

study on carp from six sites along the Mississippi 

River in Tennessee, Mississippi and Louisiana. Five 

adult male and five adult female carp were taken at 

each site, and plasma vitellogenin was measured 

using a competitive ELISA employing a polyclonal 

antibody for fathead minnow (Pimphales promelas), 

that also recognized carp vitellogenin. Parks et al. 

(1997) found no indication of vitellogenin in male 

fish, nor was there any evidence of decreased vitellogenin in females between sites.

 ESTUARINE /MARINE STUDIES 

freshwater than in estuarine/marine systems, which 

is particularly true for studies adjacent to STPs. As in 

freshwater, estrogenic effects are detected near STPs, 

and range from the presence of vitellogenin in males, 

to gonadal anomalies and even intersex fish. Almost 

no investigations on the effects of sewage treatment 

effluents on estuarine/marine species in the U.S. 

were found. 

United Kingdom. Lye et al. (1997) investigated abnor

malities in reproductive health of the European 

flounder (Platichthys flesus) exposed to effluent from 

an STP in the U.K. Three sites were chosen, two on 

the Outer Tyne Estuary where a primary level 

sewage treatment plant is located, and one on the 

Solway Firth which drains a more sparsely populated 

basin. On the Tyne Estuary, one site was located 

adjacent to a large primary treatment STP that serves 

approximately 1.25 million people, the other site was 

located 2 kilometers upstream of the outfall. 

Mature male and female flounder were collected in 

July, 1995. Parameters measured included plasma 

vitellogenin (expressed as a percentage of protein 

content compared to a control female), HSI, and 

macroscopic examination of the gonads. 

Vitellogenin was found in the plasma of male flounders from all sites investigated. Sixty percent of the 

males found at the site adjacent to the STP on the 

Outer Tyne Estuary contained detectable levels of 

vitellogenin, compared with 20 percent of male fish 

from the Solway Firth. A similar pattern was found 

for female flounder; 73-80 percent from the Tyne 

Estuary had detectable levels of vitellogenin compared with 55 percent of females from the Solway 

Firth. Lye et al. (1997) also found that male flounder 

from the Tyne Estuary had higher levels of vitellogenin than males from the Solway Firth as did females, 

although the differences in both males and females 

were not statistically significant due to variability 

within the population. 

Although not observed in any fish from Solway Firth, 

53 percent of the males taken at the site adjacent to 

the STP contained malformed testes, consisting of 

truncated lobes and thick interstitial tissue. No 

severe abnormalities were observed in females from 

any of the sites. Lye et al. (1997) concluded that the 

presence of vitellogenin in at least 20 percent of the 

fish from all sites indicated this phenomenon may be 

widespread. 

­

This section summarizes some of the endocrine 

disrupter-related field work in estuarine and marine 

species of fish. As in the last section, studies were 

organized into three broad categories and included 

investigations adjacent to STPs, pulp and paper mills, 

and general surveys that covered a defined geographic area. Table 7 summarizes results of some of 

the studies reviewed. 

Sewage Treatment Plants. It appears that more 

endocrine disrupter work has been conducted in 

35
36

Sex Land Use IS E2 GSI GD S Reference

European flounder 

(Platichthys flesus) Sewage treatment plant Outer Tyne Estuary, U.K. wild + – Natural estrogens and organic contaminants Lye et al., 1997

European flounder 

(Platichthys flesus) F Sewage treatment plant Outer Tyne Estuary, U.K. wild + Natural estrogens and organic contaminants Lye et al., 1997

European flounder 

(Platichthys flesus) Sewage treatment plant Outer Tyne Estuary, U.K. wild + + – Natural estrogens and organic contaminants Lye et al., 1998

European flounder 

(Platichthys flesus) F Sewage treatment plant Outer Tyne Estuary, U.K. wild + + – Natural estrogens and organic contaminants Lye et al., 1998

European flounder 

(Platichthys flesus) Sewage treatment plant Outer Tyne Estuary, U.K. wild + Natural estrogens and nonylphenols Lye et al., 1999

Burbot 

(Lota lota)

Sewage treatment and 

industrial facilities Bothnian Bay, Finland wild – Natural estrogens and organic contaminants Pulliainen et al. , 1992

Burbot 

(Lota lota) F Sewage treatment and industrial facilities Bothnian Bay, Finland wild – Natural estrogens and organic contaminants Pulliainen et al. , 1992

(Pleuronectes americanus Pulp and paper mill St. Georges Bay, Newfoundland wild + – Chemicals in paper mill effluent Barker et al. , 1994 

(Pleuronectes americanus F Pulp and paper mill St. Georges Bay, Newfoundland wild + – Chemicals in paper mill effluent Barker et al. , 1994 

Killifish 

(Fundulus heteroclitus) Pulp and paper mill Brunswick wild – Chemicals in paper mill effluent Leblanc et al. , 1997

Killifish 

(Fundulus heteroclitus) F Pulp and paper mill Brunswick wild – Chemicals in paper mill effluent Leblanc et al. , 1997

European flounder 

(Platichthys flesus) F contaminated sediments Netherlands wild + + NE – PCBs and PAHs Janseen et al ., 1997

European flounder 

(Platichthys flesus) contaminated sediments Netherlands wild NE PCBs and PAHs Janseen et al ., 1997

Abbreviations: NE, no effect; Vtg, vitellogenin; IS, intersex condition; T, testosterone; E2, 17 β-estradiol; GSI, gonadosomatic index; 

Table 7. Field studies of endocrine disruption in saltwater species of fish.

Caged/wild Possible causes/ Fish/Species Location/ Study Site Vtg EM fish Contaminants present

M

M

M

M

Winter flounder M

Winter flounder 

Miramachi Bay, New M

Miramachi Bay, New 

Mesocosm study; Wadden Sea, 

Mesocosm study; Wadden Sea, M

EM effect on external male/female morphology; S, effect on spawning

Estuarine/Marine Studies
Estuarine/Marine Studies

37

Sex Land Use IS E2 GSI GD S Reference

English sole 

(Parophrys vetulus) F Industrial, urban Duwamish Waterway, Puget Sound wild – – PCBs and PAHs Johnson et al. , 1988

English sole 

(Parophrys vetulus) F Industrial, urban Hylebos Waterway, Puget Sound wild ± PCBs, PAHs and other organic contaminants Collier et al. , 1998

Swordfish 

(Xiphias gladius) General survey wild + Unknown Fossi et al. , 2001

Flounder 

(Pleuronectes yokohamae General survey Tokyo Bay wild + + – Unknown Hashimoto et al. , 2000

European flounder 

(Platichthys flesus) General survey Five estuaries in the U.K. wild + + NE Natural estrogens and organic contaminants Allen et al. , 1999

Winter flounder 

(Pleuronectes americanus F General survey Long Island Sound wild + PCBs, PAHs, metals Pereira et al. , 1992

Winter flounder 

(Pleuronectes americanus F General survey Boston Harbor wild – PCBs, PAHs, metals Pereira et al. , 1992

Winter flounder 

(Pleuronectes americanus F General survey Boston Harbor wild – PCBs and PAHs Johnson et al ., 1992

White croaker 

(Genyonemus lineatus) F General survey San Pedro Bay, California wild – – DDT and PCBs Hose et al ., 1989

Killifish 

(Fundulus heteroclitus) General survey Chesapeake Bay wild NE –/NE NE Pristine to polluted areas (PAHs, PCBs, metals) Pait and Nelson, in prep.

Killifish 

(Fundulus heteroclitus) General survey New York Harbor wild NE No effects seen McArdle et al. , 1998

Killifish 

(Fundulus heteroclitus) F General survey New York Harbor wild NE No effects seen McArdle et al. , 1998

Sheepshead minnow 

(Cyprinodon variegatus) General survey Chesapeake Bay wild NE PAHs and PCBs Pait and Nelson, in prep.

Abbreviations: NE, no effect; Vtg, vitellogenin; IS, intersex condition; T, testosterone; E2, 17 β-estradiol; GSI, gonadosomatic index; 

Table 7. Field studies of endocrine disruption in saltwater species of fish (continued).

Caged/wild Possible causes/ Fish/Species Location/ Study Site Vtg EM fish Contaminants present

M Straits of Messina, Italy

M

M

M

M

M

EM effect on external male/female morphology; S, effect on spawning
Lye et al. (1998) completed a more intensive followup 

investigation on the seasonal reproductive health of 

European flounder at the same sites in the U.K. Male 

and female flounder were collected in the Tyne 

estuary, both upstream and downstream of the 

sewage treatment plant, and from the Solway Firth in 

July, September and December 1995, and April 1996. 

Lye et al. (1998) assessed relative levels of vitellogenin, gonadal histopathologies, GSI, and HSI in male

and female flounder. 

Vitellogenin was 

found in male 

fish in both 

waterways. In 

the Tyne, 64 to 95 

percent of the 

male flounder 

caught contained 

detectable levels 

of plasma 

vitellogenin, 

while 27 to 50 

percent of the 

male flounder 

from the Solway 

Firth contained 

plasma vitellogenin. While not 

quantified, levels 

of vitellogenin 

were compared 

to control 

females (Figure 

15). Male fish 

from the Tyne 

Estuary sites had 

levels of vitellogenin between 60 and 155 percent of that found in 

control female flounders. 

A substantial proportion of male fish (53 to 67 

percent) from the Tyne estuary exhibited testicular 

anomalies, compared with a lower proportion in 

males from the Solway Firth (maximum of 12 per

cent). Male flounder having the highest levels of 

plasma vitellogenin also had significant pathological 

testicular abnormalities. These included disorganized and vacuolized lobules, connective tissue 

thickening, similar to that found by Lye et al. (1997), 

along with the reduction or even absence of germ 

cells within the testes. There were no indications, 

however, of intersex fish. Female flounders from the 

Tyne exhibited increased vitellogenin levels and a 

­

higher proportion of degenerated oocytes compared 

with females from the less polluted Solway Firth. 

Vitellogenin levels in both males and females varied 

somewhat by season for both sexes. In males, there 

were no differences in testicular abnormalities 

between seasons, while in females, gross ovarian 

abnormalities (e.g., large masses of aggregated and 

deformed eggs, atretic oocytes and ovarian tissues) 

were found in September, December and April, but 

were absent in July. 

The GSI in 

females from the 

Tyne in September and April 

was significantly 

larger than those 

females from the 

Solway Firth. In 

male flounder, 

significantly 

larger GSIs in 

the Tyne were 

recorded only in 

April compared 

to those in 

Solway Firth. In 

addition, the 

HSIs for both 

female and male 

flounder from 

the Tyne were 

significantly 

higher than in 

Solway Firth. 

Lye et al. (1998) 

concluded that in addition to most U.K. rivers, the 

data strongly suggested that at least some U.K. 

estuaries are experiencing contamination by estrogenic substances. They also noted that while the 

implications for reproduction and population are not 

yet known, there is concern that the observed effects 

may impact reproductive viability which could in 

turn have serious consequences at the population 

and community levels. 

Lye et al. (1999) investigated tissue (muscle and liver) 

concentrations of alkylphenols in mature male 

European flounder in the Tyne and Tees estuaries in 

the U.K. 4-Nonylphenol (5-55 ng/g, wet weight), 

and nonylphenol monoethoxylate (190-940 ng/g wet 

weight) were found in tissues of mature male flounder. These fish also had detectable plasma vitellogeRelative (%) vtg conc.

Figure 15. Vitellogenin (vtg) (mean ± SEM) in male flounder (Platichthys flesus) from the Tyne Estuary (T1, T2, and T3) and Solway Firth 

(S) in the United Kingdom. Values are expressed as a percentage of 

vitellogenin in control female. (From Lye et al., 1998). Reprinted by 

permission. 

200 

180 

160 

140 

120 

100 

80 

60 

40 

20 

0 

S T1 T2 S T1 T2 T3 S T1 T2 T3 S T1 T2 T3 

April 1996 

July 1995 

December 1995 

September 1995 

Site 

Estuarine/Marine Studies 

38 
Estuarine/Marine Studies 

nin levels. These same compounds were also found 

in sediment samples from both estuaries. However, 

given that natural and synthetic estrogens had been 

found in domestic sewage, Lye et al. (1999) concluded 

the estrogenic responses seen in the flounder were 

unlikely to be accounted for solely by the presence of 

the alkylphenols detected in tissue and sediment. 

Pulp and Paper Mills. There have been relatively 

few investigations of endocrine disruption in estuarine and marine waters adjacent to pulp and paper 

mills. When detected, impacts tend to include 

depression or delay of the reproductive cycle in 

affected fish. A number of investigations have 

concentrated on flatfish. 

St. George’s Bay, Newfoundland. Barker et al. (1994) 

investigated the effects of pulp and paper mill 

effluents on a population 

of winter flounder 

(Pleuronectes americanus). 

Male and female fish were 

collected from two sites in 

St. George’s Bay, Newfoundland, one at Port 

Harmon adjacent to a 

pulp and paper mill, and 

at St. George’s a reference 

site located 12 kilometers 

away from the pulp and 

paper mill. The mill at 

Port Harmon produces 

paper for newsprint, and 

uses sodium 

hydrosulphite for the 

bleaching process. The 

flounder (total of 400) 

were taken monthly by hook and line or captured by 

scuba divers from 1991 to 1992 during the spring 

(May) and summer (June-September) months of both 

years. 

A number of parameters were measured in the fish, 

including GSI and the state of gonadal maturation. 

Surprisingly, the GSIs from male and female Port 

Harmon (mill site) fish were higher than those from 

the St. George’s site (Barker et al., 1994). Fish taken 

from Port Harmon in the May-June time period 

contained mature gonads, while fish from the St. 

George’s (reference) site had already spawned. In 

addition, of the fish captured during July and 

August, 12% of both males and females from Port 

Harmon contained fully developed gonads, while all 

fish from St. George’s were spent. Barker et al. (1994) 

concluded that the higher percentage of 

postspawned females and significantly lower GSIs in 

the St. George’s flounder indicated delayed spawning in fish from Port Harmon. 

Bothnian Bay, Finland/Sweden. Pulliainen et al. (1992) 

investigated the effects of industrial loadings on 

burbot on the northern coast of the Bothnian Bay, 

between Finland and Sweden. Burbot (Lota lota) 

were caught over a 3 year period in gill nets at 

various locations on the northern coast of the Bay, 

known to be impacted by inputs from pulp and 

paper, municipal sewage, and steel industries. 

Pulliainen et al. (1992) reported that 87-98% (male 

and female) of the adult burbot taken off the coasts of 

the cities of Tornio and Kemi, Finland were sterile. 

Non-maturing burbot (87 percent of catch) were also 

found in the Tornio River 

approximately 10 kilometers upstream. Pulliainen et 

al. (1992) noted that typically only 10-15 percent of 

adult burbot fail to mature. 

The male/female ratio of 

the burbot was 0.83 in 

catches off Tornio, and 0.77 

in catches off Kemi. The 

authors speculated the 

effects observed could be 

related to the presence of 

toxic chemicals, even 

though a clear cause and 

effect relationship could not 

be established. 

Miramachi Estuary, New 

Brunswick. Leblanc et al. (1997) investigated the effect 

of bleached kraft pulp mill effluents on the reproductive period in the killifish F. heteroclitus. Mature male 

and female fish were collected using minnow traps at 

locations 4, 21, and 39 kilometers downstream of the 

pulp mill. Two reference sites were established on 

the Bouctouche Estuary. A variety of parameters 

were assessed, including GSI, percentage of mature 

fish, fecundity and egg size in F. heteroclitus collected 

during the spring/summer spawning period. Fish 

from Miramachi Estuary 4 kilometers and 21 kilometers downstream of the pulp mill exhibited delayed 

gonadal maturation and smaller egg size than fish 

from the site 39 kilometers downstream of the pulp 

mill. Gonadal parameters did not vary significantly 

between sites on the Bouctouche Estuary. Interestingly, fish from the site 4 kilometers downstream of 

Significant endocrine-related impacts have been 

found in burbot Lota lota at some locations in 

Finland/Sweden. Image courtesy of Idaho State 

University - Ernest Keeley. 

39
Estuarine/Marine Studies 

the pulp mill exhibited a higher fecundity (number of 

mature eggs) and GSI compared with the other sites. 

In addition, the condition factor was not reduced at 

the site closest to the pulp mill which suggested to 

Leblanc et al. (1997), that more nutrients were available to the fish collected near the mill than at the 

other sites. 

General Surveys. Investigations of endocrine disruption in fish not tied to specific land use types or 

adjacent to effluent discharges provide an opportunity to assess the ambient level of effects in a 

waterbody. A number of studies were located from 

Europe and 

North America. 

United Kingdom. 

Allen et al. 

(1999) surveyed 

the Tyne, 

Thames, 

Crouch and 

Mersey estuar

ies and a 

number of 

offshore sites 

for evidence of 

endocrine 

disruption in 

European 

flounder. The 

estuarine 

portion of the 

River Alde was 

chosen as the 

control site. 

Between two 

and five sites 

were chosen in 

each estuary, 

and male fish were sampled two times in 1996, in 

September, and again in December. Importantly, 

sites were not specifically chosen with STP outfalls in 

mind, but rather where fish could be found (Allen, 

pers. comm.). 

On the Tyne estuary, male fish collected at all five 

sites had significantly elevated plasma vitellogenin 

levels as compared with fish from the River Alde 

control site (Figure 16). On the Thames, males from 

one of three sites had significantly elevated plasma 

vitellogenin. Males from the two sites in the Mersey

estuary had the highest levels of vitellogenin of all 

sites surveyed. Allen et al. (1999) noted that as 

expected, HSIs were elevated in fish with high 

induction, however, there was no significant effect on 

the GSIs. Male flounder collected from the North Sea 

and from the English Channel, and from most sites in 

the Irish Sea, had plasma vitellogenin levels higher 

than those fish taken from the River Alde. 

Although no serious gonadal anomalies were seen in 

fish from the Tyne, Thames, or Crouch estuaries, 

approximately 20 percent of males from the Mersey 

estuary were intersex, (i.e., containing ovotestes). 

Higher levels of contamination or the possibility that 

the fish might breed within the estuary were offered 

as possible explanations for gonadal 

anomalies and high 

levels of vitellogenin in the Mersey 

fish. Allen et al. 

(1999) concluded 

that several U.K. 

estuaries are 

suffering contamination with 

estrogenic materials which is having 

adverse biological 

effects and may be 

damaging the 

ability of the fish to 

reproduce. 

Wadden Sea, 

Netherlands. 

Janssen et al. (1997) 

exposed European 

flounder for up to 3 

years in large (40m 

X 40m X 3m) 

mesocosms. The 

first mesocosm (A) contained relatively clean 

Wadden Sea sediment and water. The second (B) 

was indirectly contaminated with water from the 

third mesocosm. The third mesocosm (C) contained 

dredge spoil from the Rotterdam Harbor. Juvenile (1­

year old) male and female flounder captured from a 

relatively clean site on the Wadden Sea were introduced into these self-contained mesocosms in 1990. 

The contaminants in the dredge spoil included PCBs 

and PAHs. A relatively clean reference site was also 

established on the Wadden Sea. Parameters included 

vitellogenin measured by gel electrophoresis and 

image analysis, E2

 and testosterone (radioimmunoassay) analyses, and gonadal histology. 

­

100,000 

Plasma Vitellogenin (ug/ml) 

10,000 

1,000 

100 

10 

1.0 

0.1 

0.01 

0.001 

* 

* 

* 

* *

* 

* 

* 

* 

Alde Crouch Thames Tyne Mersey

River Site 

Figure 16. Vitellogenin in wild male flounder from five U.K. estuar

ies. Asterisks denote significantly elevated levels (p<0.05) compared 

with control estuary (Alde). From Allen et al. (1999). Reprinted by 

permission. 

­

40
Vitellogenin

9 

8 

7 

6 

5 

4 

3 

2 

1 

0 

Meso A Meso B Meso C

November 1992 

* 

Meso A Meso B Meso C

* 

May 1993 

Figure 17. Relative plasma vitellogenin concentrations 

(mean ± SEM) of mesocosm female flounder, Platichthys 

flesus, in November 1992 and May 1993. Asterisks denote 

significant (p< 0.05) difference from the reference mesocosm 

(A). From Janssen et al. (1997). Reprinted by permission. 

Estuarine/Marine Studies 

Spawning in the flounder occurs in winter. In 

November 1992, no severe abnormalities were 

observed in ovarian structure, all ovaries were in 

early or advanced stages of vitellogenesis, character

ized by smaller oocytes. In May 1993, all females 

from the Wadden Sea reference site as well as from 

mesocosms A and B were in the previtellogenic 

phase, which is natural for that time of year. Females 

from mesocosm C, however, contained not only 

previtellogenic oocytes, but also later or yolk granule 

stage eggs, more characteristic of fish in November, 

just before spawning. This asynchronous development of oocytes was not observed in fish from 

mesocosms A and B, or from the reference field site. 

There were no 

significant differ

ences, however, in 

the GSIs between 

sites during the 

May sampling. 

Relative measurements in November 

1992 revealed that 

plasma vitellogenin 

in females in 

mesocosm C was 

significantly higher 

than females from 

mesocosm A 

(Janssen et al., 1997).

Likewise, measurements made in May 

1993 revealed the 

same trend, with 

vitellogenin levels 

highest in females 

from mesocosm C 

(Figure 17). No 

plasma vitellogenin was detected in males during 

any phase of the reproductive cycle. 

To better understand possible reasons for the increased vitellogenin in females from the contaminated mesocosm, Janssen et al. (1997) made measurements of plasma E2

 and testosterone in May 1993. 

They found that E2

 was significantly elevated in 

females in mesocosm C (264 pg/ml) as compared 

with the reference, mesocosm A (116 pg/ml). An 

accompanying set of in vitro experiments with 

ovarian tissue, however, indicated no significant 

difference in E2

 synthesis capability between the 

mesocosms. 

­

Janssen et al. (1997) concluded that long term exposure to polluted water and sediments under semifield conditions caused out-of-season plasma elevations of both E2

 and testosterone. Because the

vitellogenin was prematurely elevated in females 

only, and the ovarian E2

 production capacity was 

unaltered, Janssen et al. (1997) concluded that the 

hormonal disruption was the result of a modification 

of E2

 clearance rates rather than the estrogenic action 

of pollutants. 

Straits of Messina, Italy. Fossi et al. (2001) assessed 

reproductive endocrine disruption in swordfish 

(Xiphias gladius) in Sicily. Three female and six male 

swordfish, ranging in 

ages from 1-5 years 

were collected. 

Relative amounts 

(absorbance at 492 

nm) of vitellogenin 

and zona radiata 

proteins were measured using an 

indirect ELISA. Fossi 

et al. (2001) found 

significant induction 

of vitellogenin in the 

three older males, 4 to 

9 years of age. 

Relative vitellogenin 

levels were not 

statistically different 

from the adult 

females tested. 

Similarly, zona 

radiata proteins were 

elevated in older 

males. They attributed the higher 

biomarker responses to the accumulation of EDCs 

over time in the fish. Fossi et al. (2001) also referred 

to unpublished work that male swordfish caught in 

the Atlantic (Azores) contained substantially lower 

levels of vitellogenin than those found in the Mediterranean. They concluded the results signal a 

warning of possible reproductive alterations in this 

top predator. 

Tokyo Bay. Hashimoto et al. (2000) assessed elevated 

vitellogenin levels and gonadal abnormalities in wild 

male flounder (Pleuronectes yokohamae) from Tokyo 

Bay. Mature male flounder, aged approximately 1-6 

years were collected from two sites near Haneda and 

Yokohama in Tokyo Bay, and a reference site near 

­

41
Female English sole, Parophrys vetulus collected at 

several locations in Puget Sound in the past had lower

levels of E2

 as well as delayed vitellogenesis. Image

courtesy of NOAA/NMFS. 

Estuarine/Marine Studies 

Hakkaido. Plasma vitellogenin was measured using 

an ELISA along with a fluorophore for the detection 

of vitellogenin. 

Male flounder collected from the Tokyo Bay sites had 

significantly higher levels of plasma vitellogenin 

(0.025 - 2.2 μg/ml) than fish taken from the reference 

site (0.031 - 0.091 μg/ml). There were no significant 

correlations, however, between vitellogenin and the 

HSI, E2

 or testosterone levels, although Hashimoto et 

al. (2000) noted that lower GSIs were more often 

associated with males having higher levels of 

vitellogenin. In addition, 3 out of 20 males collected 

near Haneda were intersex although the degree of 

intersexuality was slight (Hashimoto et al., 2000). 

German Baltic Coast. Hansen 

et al. (1985) investigated the 

effects of contamination on 

Baltic herring (Clupea 

harengus) hatching success. 

Eggs stripped from ripe 

females in April were fertilized with sperm and incubated in clean seawater. 

Water exchange in the aerate

incubation jars occurred 

every 48 hours, and then 

every day during the hatching period. Hatched larvae 

were counted and viable 

hatch (straight and healthy 

larvae) were determined as a 

function of contamination in parental tissues. 

Hansen et al. (1988) found that viable hatch was 

significantly affected by ovarian DDE concentrations 

higher than 18 ng/g, and PCB concentrations of more 

than 120 ng/g, both wet weight. 

Puget Sound. A significant amount of work on sole 

and flounder has occurred in the Puget Sound region

of Washington. Much of this work has centered on 

the effects of organic contaminants on the overall 

health of fish, including reproductive success. 

Johnson et al. (1988) conducted a study to assess 

whether exposure to contaminants was associated 

with altered ovarian development in adult female 

English sole (Parophrys vetulus) from Puget Sound. 

Fish were captured from four sites in the Puget 

Sound area, including three contaminated areas: 

Duwamish Waterway (high aromatic hydrocarbons 

and PCB concentrations), Eagle Harbor (high aromatic hydrocarbons and low PCB concentrations), 

d 

Sinclair Inlet (low aromatic hydrocarbon concentrations and moderate concentrations of PCBs), and Port 

Susan, a minimally contaminated site. Parameters 

associated with ovarian maturation included GSI, 

plasma vitellogenin as estimated from alkaline-labile 

phosphate (ALP), plasma E2

, and ovarian developmental stage. Exposure to contaminants was assessed by measuring concentrations of fluorescent 

aromatic compounds in the bile, hepatic aryl hydrocarbon hydroxylase (AHH) activity, and hepatic 

concentrations of PCBs. 

Johnson et al. (1988) found that female English sole 

from the Duwamish Waterway (2.53 μg/ml) and 

Eagle Harbor (1.89 μg/ml) had significantly lower 

levels of plasma E than 

those from Port Susan 

(3.92 μg/ml). Interestingly, however, plasma 

E2

 levels in females from 

Sinclair Inlet, the 

moderately contaminated site, were significantly higher than in 

female fish from Port 

Susan. 

The sampling of fish 

occurred during winter, 

the time of gonadal 

recrudescence, prior to 

spawning. There were a 

substantial number of 

females at all sites who had not entered vitellogenesis. The number of females who had not entered 

vitellogenesis was significantly higher, however, in 

fish from the Duwamish Waterway and Eagle 

Harbor, than from the Port Susan site. They also 

found that sole with higher AHH activity were less 

likely to undergo gonadal recrudescence than those 

fish from the less contaminated areas. Johnson et al. 

(1988) concluded that contaminant exposure may 

interfere with ovarian development in female English 

sole. 

Beginning in 1994, a series of studies were conducted 

in the Hylebos Waterway in Puget Sound to deter

mine the effect of contaminants on flatfish, including 

an assessment of reproductive function in fish 

inhabiting this severely contaminated area (Collier et 

al., 1998). Contaminants in the Hylebos Waterway in 

central Puget Sound include PAHs, PCBs, 

hexachlorobutadiene, hexachlorobenzene, and a 

number of pesticides including DDTs and heptachlor. 

2

­

42
Winter flounder, Pleuronectes americanus collected from a 

number of sites in the northeast U.S., showed evidence of 

endocrine disruption. Image courtesy of NOAA/NMFS. 

Estuarine/Marine Studies 

As part of their investigation, gonadal development 

was assessed in female English sole. Collier et al. 

(1998) found precocious sexual maturation in 40-50 

percent of the juvenile female sole from the Hylebos 

Waterway as measured by fish with developing eggs. 

Approximately 20-30 percent of the adult females 

had inhibited gonadal development. Collier et al. 

(1998) found that precocious sexual maturation was 

associated with exposure to both chlorinated hydrocarbons and PAHs, while inhibition in the adults was 

related to PAH exposure. They concluded 

that the reproductive 

injuries observed in 

Hylebos Waterway fish 

presumably would 

reduce the number of 

eggs and larvae contributed by these fish to the

overall population of 

English sole. 

The effects on the 

juvenile female English 

sole found by Collier et 

al. (1998) appear similar 

to the results obtained 

by Janssen et al. (1997) 

after exposing juvenile 

European flounder to PCBs and PAHs in mesocosms 

containing dredge spoil. In the mesocosm study, 

Janssen et al. (1997) attributed precocious oocyte 

development to higher levels of E2

 in females exposed to the dredge spoil, possibly brought on by 

lower clearance rates of endogenous estrogen. 

Boston Harbor and Raritan Bay. Johnson et al. (1992) 

investigated the relationship between sediment 

contamination and a variety of reproductive parameters in prespawning winter flounder (Pleuronectes 

americanus), including ovarian developmental stage, 

ovarian atresia, GSI, plasma E2

 levels, fecundity and 

egg weight from 11 sites in the northeast U.S. Sediment contaminant concentrations ranged from 20­

50,000 ng/g dry weight for PAHs, and between 2 and 

1,400 ng/g dry weight for PCBs. Exposure to contaminants was assessed by measuring concentrations 

of fluorescent aromatic compounds in the bile, 

hepatic aryl hydrocarbon hydroxylase and concentrations of PCBs in various tissues of the flounder, 

including the ovaries. 

Decreased egg weights and an increase in atretic 

follicles were associated with high levels of PCBs or 

PAHs. However, no clear negative impact on gonadal recrudescence, GSI, plasma E2

, or fecundity in 

female winter flounder was found. Johnson et al. 

(1992) felt the apparent difference in effects in winter 

flounder compared with English sole (Johnson et al., 

1988), could be due to a number of factors. One 

possibility was interspecific differences in contaminant activation or detoxification. Another possibility 

was a noted difference in migration patterns. While 

English sole reside in the estuaries throughout the 

period of vitellogenesis 

before moving offshore 

to spawn, winter 

flounder remain 

offshore during early 

vitellogenesis, and then 

move into the estuaries 

prior to spawning. 

Johnson et al. (1992) 

hypothesized that 

winter flounder captured at the contaminated sites in Boston 

Harbor and Raritan Bay 

may not reside in these 

areas during the early 

stages of vitellogenesis, 

a time when contaminants could have a 

critical impact on steroidogenesis and subsequent 

ovarian development. 

Long Island Sound and Boston Harbor. Pereira et al. 

(1992) studied relative levels of vitellogenin in female 

winter flounder during the reproductive season at 

several contaminated sites (including Black Rock and 

New Haven harbors, Connecticut) and a relatively 

clean site in Long Island, as well as in Boston Harbor. 

Vitellogenin levels were assessed in both adult and 

juvenile flounder indirectly by measurement of ALP. 

Serum ALP values, indicative of vitellogenin, were 

actually higher in adult female flounder from the 

contaminated Black Rock Harbor (60.9 μg Pi/ml ) site 

than flounder from the relatively clean Shoreham 

(39.8 μg Pi/ml) site. There were no significant site 

differences in the juvenile females. Pereira et al. 

(1992) suggested that higher vitellogenin levels might 

be indicative of impeded uptake by the ovary, rather 

than higher E2

 levels, as higher levels of contamination have been associated with lower E2

 levels. They

also suggested the effect was likely associated with 

contaminant burden, as younger fish which exhibited 

no differences in ALP levels would not have accumu43
White croaker, Genyonemus lineatus collected in San 

Pedro Bay, California exhibited lower fecundity, possibly 

related to ovarian DDT concentrations. Image courtesy 

of NOAA/NMFS. 

Estuarine/Marine Studies 

lated the level of contaminants likely in the larger, 

older fish. 

The focus of the work in Boston Harbor was to 

discover if the presence of tumors in female winter 

flounder affected the level of vitellogenin in the fish. 

Pereira et al. (1992) found that mean ALP values in 

tumor-bearing fish (10.7 μg Pi/ml) were significantl

lower than nontumor-bearing fish (17.4 μg Pi/ml). 

They suggested it was not unreasonable that the 

potent mix of contaminants in Boston Harbor, 

including PCBs, PAHs and a variety of trace element

could have lowered vitellogenin levels in fish either 

directly by tumor-induced liver dysfunction or 

indirectly by interference with the production of E2

. 

San Pedro Bay, California. Hose et al. (1989) investigated reproductive 

impairment in female 

white croaker 

(Genyonemus lineatus) 

inhabiting waters off the 

Southern California 

coast. Croaker were 

collected from two sites, 

one a highly contaminated area in San Pedro 

Bay, and the other a 

reference site, 80 kilometers to the south. Heavy 

contamination of sediments with DDT in this 

area occurred from 1950 

to the 1970s and was the 

result of discharges from 

the Montrose Chemical Company via the Los Angeles County sewer system (Huh and Venkatesan, 

1998). The authors noted that because of a temporal 

relationship between chlorinated hydrocarbon 

discharges and the decline of several major sport and 

commercial fish stocks in coastal waters, it had been 

suggested that contaminants contributed to reduced 

fecundity or larval survival of sport fish. 

Mature female white croaker were collected from the 

sites, and following a 1 week acclimation period, 

were induced to spawn with injections (1 IU/g body 

weight) of human chorionic gonadotrophin (HCG). 

Females from San Pedro Bay produced significantly 

fewer eggs than fish from Dana Point, the reference 

site. Of the fish that did not spawn, examination of 

the ovaries revealed that spawning was imminent in 

all of the Dana Point females, but only in 16 percent 

of the San Pedro Bay females. The ovaries of the 

y 

s 

remaining San Pedro Bay fish contained only immature yolky oocytes that were unresponsive to injections of HCG. None of the fish that spawned contained ovarian DDT concentrations greater than 4 

μg/g. In addition, fish from San Pedro Bay exhibited 

lower fecundity (32 percent), and fertility (14 per

cent), leading the authors to suggest that contaminants, DDT in particular, might partially explain the 

population declines observed for a number of 

southern California species of fish since the 1940s. 

Chesapeake Bay. Pait and Nelson (in prep.) completed 

a project in the Chesapeake Bay to assess reproductive endocrine disruption in the killifish F. 

heteroclitus. Laboratory studies indicated that male F. 

heteroclitus were sensitive to 4-nonylphenol (Figure 

6), as well as to E2

. Seven field sites representing a 

variety of land use 

activities ranging from 

relatively pristine sites, to 

industrialized areas near 

Baltimore were established in the Chesapeake 

Bay and tributaries. 

Male and female fish 

were sampled during the 

fall months after reproduction had ceased, and 

again in the spring/ 

summer, the peak reproductive period for this 

species, for a period of 2 

years. All spring/ 

summer sampling 

occurred within 2 days of the full moon, to coincide 

with the species’ lunar spawning cycle. Parameters 

measured included plasma vitellogenin, GSI, HSI, 

and some gonadal histopathology. Vitellogenin was 

analyzed by direct ELISA using a monoclonal 

antibody prepared by the University of Florida. 

The results of the plasma vitellogenin analysis in 

males is shown in Table 8. Eighty-six percent of the 

male fish sampled contained no detectable levels of 

vitellogenin as determined by the ELISA. In all cases 

except one, levels of vitellogenin were low, in the 

range of 0.01 - 0.09 mg/ml. In their study in freshwater locations around the U.S., Goodbred et al. (1997) 

detected vitellogenin in approximately 16 percent of 

male carp, similar to what was found for F. 

heteroclitus. In the U.K., others (Purdom et al., 1994); 

Sumpter and Jobling, 1995), have found elevated 

levels of vitellogenin in all fish sampled, albeit close 

­

44
Table 8. Vitellogenin in male Fundulus heteroclitus from Chesapeake Bay sites. 

Site/Characteristics Fall Spring 

Site Land use 

Vitellogenin (mg/ml) 

±SEM 

(Range) 

Number 

of samples 

Vitellogenin (mg/ml) 

±SEM 

(Range) 

Number 

of samples 

Annapolis Urban/suburban 0.00±0.00 

(0.00 - 0.02) 20 0.01±0.00 

(0.00 - 0.07) 20 

Back River Sewage treatment 0.00±0.00 

(0.00 - 0.04) 20 0.00±0.00 

(0.00 - 0.03) 20 

Beaverdam Creek Reference 0.00±0.00 

(0.00 - 0.00) 10 0.01±0.01 

(0.00 - 0.09) 20 

Choptank River Agricultural 0.00±0.00 

(0.00 - 0.02) 20 0.03±0.03 

(0.00 - 0.58) 20 

Fort Armistead Industrial/sewage 

treatment 

0.01±0.00 

(0.00 - 0.07) 19 0.00±0.00 

(0.00 - 0.04) 20 

Point Lookout Reference 0.00±0.00 

(0.00 - 0.00) 10 0.00±0.02 (0.00 - 0.03) 10 

Patapsco River Industrial/urban 0.00±0.00 

(0.00 - 0.04) 20 0.00±0.00 

(0.00 - 0.03) 20 

Note - no significant differences found in vitellogenin concentration between sites or seasons. 

SEM, standard error of the mean. Pait and Nelson (in prep.). 

Estuarine/Marine Studies 

to STPs. In the 

Chesapeake, 

there were also 

occasional 

positive detections of vitellogenin in males 

from the pristine 

sites. In their 

report, 

Goodbred et al. 

(1997) suggested 

that some male 

fish may contain 

low background 

levels of vitellogenin during 

some portion of 

the reproductive 

cycle. 

A second 

species of fish, 

the sheepshead 

minnow (Cyprinodon variegatus) was also collected at 

three sites where F. heteroclitus was taken (reference 

and two contaminated sites). An analysis of 

vitellogenin in male sheepshead minnows revealed 

no significant difference from F. heteroclitus. 

Although there were no apparent estrogenic effects, 

there was some evidence of negative effects at certain 

sites. The GSIs in male and female fish from Baltimore were significantly lower than fish from the 

reference site (Beaverdam Creek). Additionally, 

females from Baltimore had vitellogenin levels 

significantly lower than Beaverdam Creek. Data on 

sediment contaminant concentrations for the Baltimore locations as well as a reference site were 

obtained from published data bases. An analysis of 

the data revealed that lower GSIs in male F. 

heteroclitus were negatively and significantly correlated with sediment concentrations of PAHs and 

PCBs, and in females, lower vitellogenin levels and 

GSIs were negatively correlated with sediment PAHs. 

Histological examination of gonadal sections, however, did not reveal any ovotestes and also failed to 

reveal any major differences in the fish between 

sites. 

New York Harbor. McArdle et al. (1998) also investigated endocrine disruption in F. heteroclitus. Parameters measured included liver microsomal vitellogenin, GSI and HSI. 

Relative amounts of vitellogenin were assessed using 

Western blot and a polyclonal antibody for F. 

heteroclitus. The intensity of the vitellogenin band of 

each sample was compared with the microsomal 

protein band of a vitellogenic female from a control 

site (Flax Pond), loaded at three different concentrations (1, 5, and 10 μg) into the wells of the gel. 

Adult male and female F. heteroclitus were collected 

by minnow trap in Newark Bay. None of the 22 

adult males contained detectable levels of vitellogenin, while all five female F. heteroclitus contained 

levels of vitellogenin that were intermediate of the 1 

and 5 μg bands. 

Zhou et al. (2000) investigated thyroid disfunction in 

F. heteroclitus in New Jersey. Fish were collected 

using minnow traps from two locations, Piles Creek, 

which is contaminated with petroleum hydrocarbons 

and heavy metals, and from a reference site near 

Tuckerton, New Jersey. They noted that fish from 

Piles Creek have reduced growth and longevity 

along with behavior deficits compared to the refer

ence population. 

The thyroid glands were processed for histologic 

examination, and concentrations of the hormones 

thyroxine and triidothyronine were analyzed using 

­

45
The killifish, Fundulus heteroclitus, also known as 

the mummichog has been used in a number of 

laboratory and field investigations of endocrine 

disruption. From Goode, 1884. 

Conclusions 

radioimmunoassay. Zhou et al. (2000) found that fish 

collected from the contaminated Piles Creek site had 

abnormally large thyroid follicles with thickened 

epithelia. Although levels of triidothyronine were 

normal, levels of thyroxin were elevated. Zhou et al. 

(2000) noted the altered thyroid status may correlate 

with the sluggish behavior patterns observed in the 

Piles Creek fish, and that it was probably an example 

of environmentally related endocrine disruption.

 CONCLUSIONS 

Endocrine disrupting compounds or EDCs have the 

potential to interfere with a variety of life processes 

controlled or influenced by the endocrine system. A 

number of synthetic and natural chemicals and 

chemical classes including synthetic estrogens (e.g., 

17α-ethinylestradiol, EE2

), industrial intermediates 

(e.g., 4-nonylphenol, and bisphenol-A), PAHs (e.g., β

napthoflavone, benzo[a]pyrene, phenanthrene, and 

chrysene) pesticides (e.g., DDT, endosulfan, 

carbofuran and atrazine), metals (e.g., cadmium and 

lead), dioxins, and PCBs 

(e.g., Aroclor 1254 and 

Aroclor 1221) have been 

shown to adversely affect the 

endocrine system in fish. In 

addition, certain naturally 

occurring compounds, (e.g., 

β-sitosterol) can also interfere 

with normal endocrine 

function. The potency of 

most if not all EDCs is 

typically hundreds to 

thousands of times less than 

that of the endogenous 

hormones. 

To date, most of the research on endocrine disruption 

in fish has concentrated on reproductive effects. 

Other areas of the endocrine system in fish and 

higher vertebrates are also at risk. 

Significant environmental endocrine disruption has 

been seen in fish. Effects range from reduced levels 

of the reproductive hormones (i.e., estrogens and 

androgens), inhibited gonadal growth, the appear

ance of a female egg protein (vitellogenin) in male 

fish, to gonadal histopathologies and even intersex 

fish containing ovotestes. 

Overt endocrine disruption in fish, however, does not 

appear to be a ubiquitous environmental phenom­

­

enon. When detected, endocrine disruption appears 

more likely to be associated with higher levels of 

contamination; fish from less polluted areas or areas 

downstream of pollutant discharges or spills typically are less affected using current indicators. It 

remains to be seen, however, whether subtle effects 

may be occurring at very low levels of contamination, a concern that has been raised by some laboratory studies. 

Endocrine disruption in fish appears more frequently 

associated with three types of land use: sewage 

treatment plants (STPs) (estrogenic effects), pulp and 

paper mills (antiestrogenic and probably 

antiandrogenic effects), and areas of high industrial 

activity/chemical contamination (antiestrogenic and 

possibly antiandrogenic effects). 

Some of the most significant examples of endocrine 

disruption in fish (e.g., ovotestes and degraded 

gonadal structure) have been found adjacent to STPs, 

particularly in the United Kingdom. Impacts such as 

these appear to be less common in the U.S. The 

reasons for this are still unclear, but may be related to 

higher population densities, smaller flow rivers in 

which effluents can be 

diluted, different treatment 

regimes, or perhaps simply 

a greater number of 

investigations in the U.K. 

Compounds thought 

responsible for a significant 

portion of the estrogenic 

effects observed in fish 

near STPs include the 

natural estrogens E2

, its 

metabolite estrone, and the 

synthetic estrogen EE2

, which are not entirely removed or degraded by treatment processes. To a 

lesser extent, the degradation products of alkylphenol polyethoxylate surfactants (APEs) have also been 

implicated as contributing to observed effects. High 

use and discharge of APEs, as has been seen downstream of STPs that receive influent from certain 

textile (ie., wool processing) operations, may in some 

cases contribute very significantly to the observed 

estrogenicity in a river. 

A second activity frequently associated with endocrine disruption in fish is the discharge of wastes 

from pulp and paper mills, which have been associated with both antiestrogenic and antiandrogenic 

­

46
Conclusions 

effects. One candidate compound that may be 

associated or contributing to the observed effects is 

the plant sterol, β-sitosterol, generated during the 

wood pulping process, although the possible role of 

the dioxins is still being debated. β-Sitosterol may 

contribute to observed effects by suppressing production of the reproductive hormones through 

inhibition of the enzyme systems involved in the 

conversion of cholesterol to estrogens and androgens. 

Effects include decreased levels of the steroid hormones, decreased gonadosomatic indices, and 

external masculinization of female fish. A number of 

studies in the U.S. and in northern Europe have 

detected these and other biochemical and morphologic effects on resident populations of fish. 

Endocrine disruption has also been seen in areas of 

high industrial activity and chemical contamination. 

Although somewhat variable, effects typically 

include reduced levels of estrogens and inhibition of 

gonadal growth in female fish. More work is needed, 

however, to assess the effects of contaminants, 

particularly PAHs and PCBs on the endocrine system 

in male fish. Compounds thought responsible for 

observed effects include PCBs, PAHs and possibly 

dioxins. A number of studies in the Puget Sound 

area, and also in Europe have indicated some impacts in flatfish. 

Although endocrine disruption in fish has been 

documented in a number of field studies, little is 

currently known regarding the effect of EDCs at the 

population level. Overall, it does not appear that 

EDCs are currently having serious effects on large 

populations of fish. However, several authors have 

noted that reproduction and population structure are 

almost certainly affected in areas where the greatest 

impacts on individual fish have been found, such as 

near certain STPs. To address this lack of information, the National Research Council (NRC, 1999) has 

recommended long term studies of populations 

subjected to EDCs in order to assess effects on 

population size, age structure and dynamics. 

Detecting changes in population structure, however, 

will be difficult in part due to the mobility of many 

fish species and the possible recruitment of new 

individuals from outside the affected population. In 

addition, detection of subtle effects, such as differences in levels of endogenous hormones, is difficult 

due to natural hormonal variability between fish, and 

by a lack of knowledge regarding what is “normal” 

for each species, and for each species in a particular 

area. 

As more laboratory and field research is conducted 

on the identity, presence, and effects of EDCs, a 

clearer picture will emerge of the risks to fish and 

higher vertebrates, as well as strategies to reduce or 

even eliminate the discharge and effects of these 

compounds. 

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