Biomonitoring of Environmental Status

and Trends (BEST) Program: Selected

Methods for Monitoring Chemical

Contaminants and their Effects in

Aquatic Ecosystems

Information and Technology Report

USGS/BRD/ITR--2000-0005

U.S. Department of the Interior

U.S. Geological Survey

Biological Resources Division

Cl

Cl

Cl

Cl

Cl

O

O

Cl

Cl

Cl

Cl O

Cl

Cl

Cl

Cl

(a) 2,3,7,8-Tetrachlorodibenzo- p-dioxin (b) 2,3,7,8-Tetrachlorodibenzofuran

(c) 3,3 ́,4,4 ́,5-Pentachlorobiphenyl (PCB 126) (d) Benzo[a]pyrene
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Biomonitoring of Environmental Status

and Trends (BEST) Program: Selected

Methods for Monitoring Chemical

Contaminants and their Effects in

Aquatic Ecosystems

Editors :

Christopher J. Schmitt1

Gail M. Dethloff2

Contributors (in alphabetical order):

Vicki S. Blazer3

Nancy D. Denslow4

Gail M. Dethloff2

Timothy S. Gross5

Kelly K. McDonald5

Christopher J. Schmitt1

Donald E. Tillitt1

Jeff J. Whyte6

1

U.S. Geological Survey, Columbia Environmental Research Center, Columbia, MO

2

AScI, c/o U.S. Geological Survey, Columbia Environmental Research Center, Columbia, MO

3

U.S. Geological Survey, Leetown Science Center, National Fish Health Laboratory, Leetown, WV

4

University of Florida, Department of Biochemistry, Gainesville, FL

5

U.S. Geological Survey, Florida-Caribbean Science Center, Gainesville, FL

6

University of Missouri, Department of Fisheries and Wildlife, Columbia, MO
Suggested citation:

Schmitt, C. J. and G. M. Dethloff. editors. 2000. Biomonitoring of Environmental Status and Trends (BEST) Program:

selected methods for monitoring chemical contaminants and their effects in aquatic ecosystems. U.S. Geological Survey,

Biological Resources Division, Columbia, (MO): Information and Technology Report USGS/BRD-2000--0005. 81 pp.
i

CONTENTS

ABSTRACT...............................................................................................................................................................iv

PREFACE.................................................................................................................................................................vi

THE 1995 AND 1997 BEST PROJECTS. C.J. Schmitt, D.E. Tillitt, V.S. Blazer..........................................................1

SITE SELECTION..........................................................................................................................................1

ORGANISMS SELECTED FOR SAMPLING.........................................................................................................2

METHODS EVALUATED................................................................................................................................3

EROD ACTIVITY. J.J. Whyte, D.E. Tillitt...............................................................................................................5

Background ..............................................................................................................................................5

Performing the EROD Assay....................................................................................................................7

Factors That Can Afftect EROD Induction in Fish...................................................................................7

Value and Utility of EROD in the BEST Program....................................................................................8

H4IIE BIOASSAY. J.J. Whyte, D.E. Tillitt...............................................................................................................9

Background ..............................................................................................................................................9

Performing the H4IIE Bioassay...............................................................................................................10

Factors That Can Affect the H4IIE Bioassay...........................................................................................10

Value and Utility of the H4IIE Bioassay in the BEST Program ..............................................................12

CONDITION FACTOR AND ORGANO-SOMATIC INDICES. G.M. Dethloff, C.J. Schmitt............................................13

Background ............................................................................................................................................13

Measuring Condition Factor and Organo-somatic Indices......................................................................13

Factors That Can Affect Condition Factor and Organo-somatic Indices.................................................13

Value and Utility of the Condition Factor and Organo-somatic Indices in the BEST Program...............17

THE NECROPSY-BASED FISH HEALTH ASSESSMENT. V.S. Blazer.........................................................................18

Background.............................................................................................................................................18

Performing the Necropsy-Based Fish Health Assessment......................................................................18

Factors That Can Affect the Necropsy-Based Fish Health Assessment..................................................18

Value and Utility of the Necropsy-Based Fish Health Assessment in the BEST Program......................22

HISTOPATHOLOGIC ASSESSMENT. V.S. Blazer........................................................................................................22

Background ............................................................................................................................................22

Performing Histopathlogical Analyses...................................................................................................23

Factors That Can Affect Histopathological Biomarkers.........................................................................23

Value and Utility of Histopathological Biomarkers in the BEST Program.............................................25

IMMUNE SYSTEM INDICATORS. V.S. Blazer, G.M. Dethloff..................................................................................25

SERUM/PLASMA LYSOZYME ACTIVITY...................................................................................................................26

Background.............................................................................................................................................26

Performing the Lysozyme Assay.............................................................................................................26

Factors That Can Affect Serum/Plasma Lysozyme Activity...................................................................27
ii

Value and Utility of Serum/Plasma Lysozyme Activity in the BEST Program.......................................27

MACROPHAGE AGGREGATES.................................................................................................................................27

Background.............................................................................................................................................28

Measuring the Macrophage Aggregate Parameters................................................................................28

Factors That Can Affect Macrophage Aggregates..................................................................................29

Value and Utility of Macrophage Aggregates in the BEST Program......................................................30

REPRODUCTIVE INDICATORS. K.K. McDonald, T.S. Gross, N.D. Denslow, V.S.Blazer........................................30

SEX STEROIDS......................................................................................................................................................31

Background.............................................................................................................................................31

Performing Sex Steroid Hormone Assays...............................................................................................31

Factors That Can Affect Sex Steroid Hormones.....................................................................................31

Value and Utility of Sex Steroid Hormones in the BEST Program.........................................................32

VITELLOGENIN.....................................................................................................................................................34

Background.............................................................................................................................................34

Performing the Vitellogenin Assay.........................................................................................................34

Factors That Can Affect Vitellogenin.....................................................................................................35

Value and Utility of Vitellogenin in the BEST Program..........................................................................36

GONADO-SOMATIC INDEX (GSI) AND GONADAL HISTOPATHOLOGY.......................................................................37

Background.............................................................................................................................................37

Measuring the GSI and Performing Gonadal Histopathology.................................................................37

Factors That Can Affect the GSI and Gonadal Histopathology...............................................................39

Value and Utility of the GSI and Gonadal Histopathology in the BEST Program...................................39

SUMMARY...........................................................................................................................................................41

STABLE NITROGEN ISOTOPE RATIO ( δ

15N). C.J. Schmitt.....................................................................................42

Background.............................................................................................................................................42

Measuring Stable Nitrogen Isotopes.......................................................................................................43

Factors That Can Affect δ

15

N...............................................................................................................43

Value and Utility of the δ

15N in the BEST Program...............................................................................43

APPENDIX I: DEVELOPMENT OF THE BEST PROGRAM. C.J. Schmitt..................................................................45

THE NATIONAL CONTAMINANT BIOMONITORING PROGRAM...................................................................................46

Fish Sampling Under the NPMP and NCBP...........................................................................................46

Shortcomings of the NCPB Fish Network..............................................................................................46

Geographic, Programmatic, and Taxonomic Shortcomings.....................................................46

Chemical Shortcomings...........................................................................................................47

FEATURES OF THE BEST PROGRAM......................................................................................................................51

Weight-of-Evidence Approach...............................................................................................................51

Organisms and Habitats Sampled...........................................................................................................52

Monitoring Methods ..............................................................................................................................53

Analytical Chemistry................................................................................................................54

Bioassays and Toxicity Tests....................................................................................................54
iii

Biomarkers and Indicators of Organism Health........................................................................55

Population and Community Health Indicators..........................................................................55

ACKNOWLEDGEMENTS...........................................................................................................................................55

REFERENCES..........................................................................................................................................................56

TABLES

Table 1. Methods incorporated into the large rivers projects....................................................................4

Table 2. Monitoring and assessment strategy for PAHs andPHHs...........................................................5

Table 3. Investigations in which the HSI has been evaluated as a biomarker.........................................15

Table 4. Necropsy classificaton..............................................................................................................19

Table 5. Fish health assessment variables and substituted values...........................................................20

Table 6. Histological features of gonadal stages as defined for use in the BEST program.....................40

Table A-1. Contaminants of concern to the BEST program (not mutually exclusive)................................49

Table A-2. Primary and secondary guilds identified for sampling aquatic habitats....................................52

Table A-3. Habitat classification system used to identify habitats for sampling........................................53

FIGURES

Figure 1. Stations sampled durimg the BEST projects in 1995 and 1997..................................................2

Figure 2. Proposed mechanism of AhR-mediated toxicity.........................................................................6

Figure 3. Representative AhR ligands.......................................................................................................7

Figure 4. Ah receptor-mediated formation of DNA adducts and tumor initation......................................8

Figure 5. Application of the H4IIE cell line bioassay...............................................................................11

Figure 6. Hatching success in Great Lakes colonies of double-crested cormorants.................................12

Figure 7. Erosion of the caudal fin...........................................................................................................18

Figure 8. Ulceration of the skin................................................................................................................21

Figure 9. A missing or scarred-over eye..................................................................................................22

Figure 10. Parasites on the gill...................................................................................................................22

Figure 11. Tissue in fixative.......................................................................................................................23

Figure 12. Examples of histologic biomarkers in the liver.........................................................................24

Figure 13. Splenic tissue............................................................................................................................28

Figure 14. A macrophage aggregate in a tissue section stained by the Perl’s method................................29

Figure 15. The reproductive hormone pathway in teleosts........................................................................32 

Figure 16. Mean plasma steorid concentrations for male and female largemouth bass.............................33

Figure 17. The role of the ovaries and liver in vitellogenesis.....................................................................34

Figure 18. The standard curve in a vitellogenin assay................................................................................35

Figure 19. Mean plasma vitellogenin concentrations for male and female largemouth bass....................36

Figure 20. Histologic appearance of intersex fish......................................................................................39

Figure 21. Atretic eggs present in ovaries at two different stages of development....................................41

Figure 22. Mean (±s.e.) δ

15N for components of the Lake Ontario pelagic food chain............................42

Figure 23. Relationship between mean mercury concentration and mean δ

15N in lake trout...................43

Figure A-1. The network of stations for the National Contaminant Biomonitoring Program......................47
iv

Abstract: This document describes the suite of biological methods of the U.S. Geological SurveyBiomonitoring of Environmental Status and Trends program for monitoring chemical contaminants and their

effects on fish. The methods, which were selected by panels of experts, are being field-tested in rivers of the

Mississippi River, Columbia River, and Rio Grande basins. General health biomarkers include a health assessment index based on gross observation; histopathological examination of selected organs and tissues; condition factor; and the heptosomatic and splenosomatic indices. Immune system indicators are plasma lysozyme

activity and measures of splenic macrophage aggregates. Reproductive biomarkers include plasma concentrations of sex steroid hormones (17β-estradiol and 11-ketotestosterone) and vitellogenin, gonadal histopathology

(including reproductive stage and, in females, gonadal atresia), and the gonadosomatic index. Indicators of

exposure to polycyclic aromatic and polyhalogenated hydrocarbons are the H4IIE rat hepatoma cell bioassay

(performed on solvent extracts of composite fish samples) and hepatic ethoxyresorufin-O-deethylase activity.

Stable nitrogen isotope ratios are used to assess the trophic position of the fish and their exposure to sewage

and other animal wastes. For each indicator we describe endpoint(s) and methods, and discuss the indicator’s

value and limitations for contaminant monitoring and assessment. 

Keywords: Contaminants, biomarkers, fish health, aromatic hydrocarbons, histopathology, atresia, sex steroid

hormones, 17B-estradiol, 11-ketotestosterone, vitellogenin, GSI, HSI, SSI, HAI, EROD, H4IIE, condition factor, lysozyme, macrophage aggregates, nitrogen isotope ratios. 

Biomonitoring of Environmental Status and Trends (BEST) Program:

Selected Methods for Monitoring Chemical Contaminants and their

Effects in Aquatic Ecosystems

Christopher J. Schmitt and Gail Dethloff, editors
v
vi

PREFACE

The Biomonitoring of Environmental Status and Trends (BEST) program was initiated in 1991 by the U.S.

Fish and Wildlife Service (FWS) as a revision and expansion of the National Contaminant Biomonitoring

Program (NCBP). The NCBP originated as part of the National Pesticide Monitoring Program (NPMP), a

multi-agency effort through which temporal and geographic trends in persistent contaminant concentrations

were documented through the collection and chemical analysis of environmental media. The FWS participated in the NPMP by collecting and analyzing freshwater fish and European starlings (Sturnus vulgaris) from

nationwide networks of stations and by analyzing the wings of hunter-killed ducks from the four major U.S.

flyways (Johnson et al. 1967). In response to a growing need to address new classes of chemicals for which

the chemical analysis of animal tissues was not an effective monitoring method, and to focus its activities on

biological resources rather than on chemicals, the FWS re-named its monitoring program (to NCBP) in the

anticipation of adding biological monitoring and assessment components targeted towards lands and species of

concern to the agency. To a great extent, the BEST program is the result of FWS planning for this expanded

monitoring. In 1993, the BEST program, as well as curatorial responsibility for the historic NCBP data bases

and sample archives, were transferred to the National Biological Service (BEST 1996). The NBS became the

Biological Resources Division of the U. S. Geological Survey in 1996. Appendix 1 traces the development of

the BEST program in detail.

To test the feasibility of implementing selected components of the BEST program at the national and

regional scales, pilot projects based on the historic NCBP network of sampling stations were designed and initiated in 1995 and 1997 (Schmitt et al. 1995; Bartish et al. 1997). This report summarizes the scope and intent

of the projects relative to the continuing development of the BEST program and the monitoring methods being

evaluated.
THE 1995 AND 1997 BEST PROJECTS

Christopher J. Schmitt, Donald E. Tillitt and Vicki S. Blazer

The National Contaminant Biomonitoring Program

(NCBP) fish network was last fully sampled and analyzed in 1986 (Schmitt et al. 1999). Due to continued

interest in bioaccumulable contaminants, the BEST

program proposed to update selected parts of the

NCBP database, beginning with the Mississippi River

basin, in the fall of 1995 (Schmitt et al. 1995). These

investigations would also serve as platforms for

demonstrating and further evaluating some of the biomarkers and bioassays proposed for use in the program in a habitat class identified as important to the

BEST program — freshwater, flowing, large river

(BEST 1996) — at a large scale. The 1995 study also

included sites on smaller rivers and streams in selected basins sampled by the National Water Quality

Assessment (NAWQA) program of the U.S.

Geological Survey-Water Resources Division (USGSWRD). A second round of sampling was initiated in

1997 in the Rio Grande and Columbia River basins

(Bartish et al. 1997), in cooperation with the National

Stream Quantity Accounting Network (NASQAN II)

program of the USGS-WRD (Hooper et al. 1996).

The specific objectives of the 1995 and 1997 studies

were to: 

(1) Document and characterize the geographic distribution of chemical contaminants and their effects on

fish and wildlife in the large rivers of the Mississippi,

Columbia, and Rio Grande basins, and compare the

findings for bioaccumulable contaminants with those

of previous NCBP fish collections.

(2) Field test, evaluate, and optimize aquatic indicators selected for use in the BEST program.

(3) Evaluate and demonstrate the technical and logistic feasibility of implementing the BEST program

through partnerships with science centers, cooperative

research units, universities, and other monitoring programs and Department of the Interior agencies.

(4) Evaluate the compatibility of selected components

of the BEST program with the NAWQA and

NASQAN-II programs of the USGS-WRD. 

SITE SELECTION

The 1995 and 1997 projects were developed with the

cooperation of the NAWQA and NASQAN II programs of the USGS-WRD because these programs

collect a wide range of water quality and hydrologic

data. The NAWQA program also collects information

on land use and ecological conditions that may be

useful for evaluating the utility of the biomarkers

selected for use in the BEST program and for interpreting the results of the investigations. For 1995, the

38 NCBP stations located in the Mississippi River

drainage were targeted for sampling. These stations

are affected by a wide range of agricultural, industrial,

mining, and urban pollutants (Goolsby 1996). In

addition, the stations in the Mississippi basin were

most in need of updated information because extensive flooding in the Midwest during 1993 and 1995

greatly redistributed contaminants since the last

NCBP collection in 1986 (Rostad 1997). For budgetary reasons, three targeted sites were eliminated from

consideration — two in the upper Platte River system,

which had been sampled recently by the NAWQA

program (Goodbred et al. 1997; Tate and Heiny 1996;

Heiny and Tate 1997); and one on the Upper Missouri

that had historically yielded cold-water fishes, a guild

not targeted by our study. To integrate the study with

ongoing NAWQA investigations and thereby expand

it into more habitats, NAWQA sites on lower-order

streams in smaller watersheds within the Mississippi

basin were also included. A reference site (water supply reservoirs of the USGS Leetown Science Center

in Kearneysville, WV) was included. The final list of

1995 stations comprised 35 former NCBP stations, all

but one of which were sampled; a total of 13

NAWQA sites in two study units — four in the

Eastern Iowa Basins (Kalkhoff 1994) and nine in the

Mississippi Embayment (Mallory 1994) Study Units;

and the reference site (Fig. 1).

The 1997 study represented an extension of

the 1995 project, with the added specific objective of

testing the compatibility of the protocol and methods

evaluated in the 1995 Mississippi basin project with

the re-design of the NASQAN program (NASQAN II;

Hooper et al. 1996). The NASQAN II program monitors dissolved and suspended concentrations of many

water constituents (elemental contaminants,

hydrophilic pesticides, nutrients, etc.) in four large

river systems — the Mississippi, Columbia, Colorado,

and Rio Grande. The information on bioaccumulable

contaminants and contaminant effects generated by

the BEST program therefore complemented the chemical measurements of the NASQAN II program.

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 1
2 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Consequently, the 1997 project was also considered a

pilot for a possible national-scale combined network

for large rivers. Stations selected for sampling in

1997 included ten historic NCBP stations and four

NASQAN II stations in the Columbia basin; and four

NCBP and six NASQAN II stations in the Rio

Grande basin. An additional site in the Columbia

basin (Station 506) was added to provide collateral

information in support of an ongoing investigation of

contaminants and their effects on osprey (Pandion

haliaetus) reproduction (Henny unpub. data) (Fig. 1).

Like those in the Mississippi basin sampled in 1995,

the stations in the Columbia River and Rio Grande

basins are affected by a wide range of agricultural,

industrial, mining, and urban pollutants (Kelly 1996).

ORGANISMS SELECTED FOR SAMPLING

The 1995 and 1997 projects represented a bridge

between the NCBP and the BEST programs.

Consequently, the fishes selected for sampling in the

projects represented a blend of the historic NCBP fish

sampling protocol and the guilds identified for monitoring aquatic habitats by the BEST program (BEST

1996). For the projects, the chosen guilds were sediment-dwelling or invertebrate-eating fish and piscivorous fish, these representing a combination of the recommended primary and secondary BEST program

guilds (Appendix 1, Table A-2). The guild choices

were necessitated by the need for rapid implementation of the projects and a corresponding lack of onthe-ground capacity for sampling piscivorous birds at

the scale of the project.

Within the chosen guilds of fishes, preferred

taxa were identified for three habitat classes — cold-,

cool-, and warm-water — as they had been for the

NCBP (Schmitt et al. 1990; Schmitt et al. 1999). For

warm-water habitats, the preferred species were common carp (Cyprinus carpio) as the sediment-dwelling

species and largemouth bass (Micropterus salmoides)

or other black bass (i.e., Micropterus spp.) as the piscivore. These taxa were selected because they had

historically been the ones most frequently collected at

warm-water NCBP sites in the Mississippi basin (e.g.,

Schmitt et al. 1990; Schmitt et al. 1999), and their

continued collection facilitates analyses of temporal

trends, one of the stated objectives of the study and of

the BEST program (BEST 1996; Schmitt et al. 1995).

These are also species for which the chosen biomarkers have been validated, and there is consequently a

literature against which to compare the findings of the

studies.

Figure 1. Stations sampled during the BEST projects in 1995 and 1997. 
Within the chosen guilds and habitats

(Appendix 1, Tables A-2 and A-3), alternate taxa

were also identified for collection in the event that the

preferred species could not be found. Alternative

species for warm-water habitats were white bass

(Morone chrysops) or other percichthyid (Morone

spp.) for the piscivorous species and various suckers

(Catostomidae) and catfishes (Ictaluridae) for the sediment-dwelling, invertebrate-eating species. An

extensive body of literature exists for contaminant

concentrations and biomarker performance in these

taxa (e.g., Schmitt et al. 1990; Munkittrick et al.

1997; Schrank et al. 1997). In addition, suckers accumulate relatively high concentrations of a number of

metals, including lead (Schmitt et al. 1985; Schmitt et

al. 1993; Schmitt and Finger 1987). Alternative piscivores for cool-water habitats were the percids [walleye (Stizostedion vitreum) and sauger (S. canadense)]

and northern pike (Esox lucius), which are all welldocumented accumulators of mercury (Wiener and

Spry 1996). In addition, risk of polyhalogenated

hydrocarbons (PHHs) to wildlife has also been

assessed using walleye (Giesy et al. 1995; Giesy et al.

1997). Rainbow trout (Oncorhynchus mykiss) and

brown trout (Salmo trutta) were chosen as representative piscivores for cold-water habitats, and suckers

were chosen as the sediment-dwelling taxa in both

cool- and cold-water habitats. These taxa had also

been collected historically at their respective NCBP

sites, and the performance of many of the chosen biomarkers has been documented for them. In the

Columbia River system, the northern pikeminnow

(formerly northern squawfish, Ptychocheilus oregonensis) was permitted as an alternate piscivore; this

species had been targeted in many previous NCBP

collections and has been shown to accumulate mercury (Schmitt and Brumbaugh 1990).

METHODS EVALUATED

In the 1995 and 1997 projects, BEST program methods for aquatic habitats (BEST 1996) suited to fish

were evaluated (Table 1). In addition to analyses of

the fish carcasses for elemental and organic chemical

residues (following NCBP protocol) and for stable

isotopes of nitrogen (δ

15N), methods included condition indices and gross and histopathological examination incorporating a quantitative necropsy-based field

health assessment. Because of growing interest in the

effects of chemicals on the endocrine system (e.g.,

Colborn 1991; Colborn et al. 1993; Guillette et al.

1994), biomarkers diagnostic of reproductive health

and endocrine modulation or disruption in fish that

were tested in an earlier study (Goodbred et al. 1994;

Goodbred et al. 1997) were incorporated into the suite

of fish health indicators for evaluation. These included plasma concentrations of reproductive steroid hormones (Guillette et al. 1994) and vitellogenin (Folmar

et al. 1996). Furthermore, possible suppression of the

immune system was investigated using two parameters indicative of immune system function — splenic

macrophage aggregates (Blazer et al. 1994a; Blazer et

al. 1997) and plasma lysozyme activity (Blazer et al.

1994a). These latter methods, ranging from organism-level to subcellular, were used as indicators of

fish health. Health has been defined as the residual

capacity to withstand stress. Hence, the more stressed

(less healthy) an organism is, the less capable it is to

withstand further stress (Bayne et al. 1985).

Assessments of fish health attempt to integrate the

overall responses of an organism to environmental

stressors, including exposure to xenobiotics. 

Two additional assays, the H4IIE rat

hepatoma cell bioassay and hepatic ethoxyresorufinO-deethylase (EROD) activity, were used together

with analyses of fish carcasses for elemental and

organic chemical residues to provide qualitative and

quantitative information on the concentrations and

biological effects of PHHs [e.g., polychlorinated

biphenyls (PCBs), dioxins (PCDDs), furans (PCDFs)

and polycyclic aromatic hydrocarbons (PAHs)]. The

approach being evaluated assesses these complex

groups of structurally and toxicologically similar contaminants without resorting to costly high-resolution

analytical methods. In this approach, total PCB concentrations in whole fish were quantified along with

organochlorine pesticides by gas chromatography

with electron capture detection following NCBP protocol (Schmitt et al. 1990). Extracts from the composite samples were then screened for the presence of

aryl hydrocarbon hydroxylase (AHH)-active compounds by the H4IIE bioassay (Tillitt et al. 1991), a

sensitive in vitro method for documenting the cumulative concentrations of dioxin-like PHH residues. The

assay was scaled to a 2,3,7,8-tetrachloro-p-dibenzodioxin (dioxin) standard and results were reported in

units of total dioxin-equivalent concentrations. The

H4IIE assay also responds to AHH-active PAHs

(Willett et al. 1997), traces of which might occur in

the whole-fish extracts if environmental PAH concentrations are extremely high (Baumann et al. 1982).

To remove the PAHs and other labile compounds, the

extracts were first subjected to a reactive cleanup procedure (Schwartz and Lehmann 1982), leaving only

PHHs and other refractory compounds in the extracts

to be assayed. Used in this manner, the H4IIE assay

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 3
4 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Table 1. Methods incorporated into the large rivers projects. 

Method Description 

Tissue(s) 

examined Sensitivity 

Primary 

reference(s) 

Histopathology Microscopic examination 

for the presence of lesions; 

can provide early indication 

of chemical exposure 

Liver, gill, 

gonads, spleen, 

and kidney 

Overall 

organism 

health and 

contaminants 

Hinton et al. (1992); 

Hinton (1993); 

Goodbred et al. (1997) 

EthoxyresorufinO-deethylase 

(EROD) activity 

Enzyme induction by planar 

hydrocarbons 

Liver PCBs, PAHs, 

dioxins, and 

furans 

Pohl and Fouts (1980); 

Kennedy and Jones 

(1994) 

Lysozyme activity A disease resistance factor 

that can be suppressed in the 

presence of contaminants 

Blood plasma Overall 

organism 

health 

Blazer et al. (1994a) 

Macrophage 

aggregate analysis 

Macrophages are important 

in the immune system, 

serving as a first line of 

defense for the organism 

and as an antigen processing 

cell 

Spleen, 

hemopoetic 

kidney, and 

liver 

Multiple 

contaminants 

including 

PAHs and 

metals 

Blazer et al. (1994a); 

Blazer et al.(1997) 

H4IIE bioassay A screening tool to 

determine the presence of 

certain classes of planar 

halogenated compounds 

Whole fish 

(composites) 

PCBs, dioxins, 

furans, and 

PAHs 

Tillitt et al. (1991) 

Vitellogenin A precursor of egg yolk, 

normally synthesized in the 

liver of female fish 

Blood plasma Endocrine 

modulating 

compounds 

Folmar et al. (1996) 

Sex Steroids 

(estradiol and 

testosterone) 

Determine reproductive 

health and status 

Blood plasma Endocrine 

modulating 

substances 

Guillete et al. (1994); 

Goodbred et al. (1997) 

Chemical analyses Organochlorine chemical 

residues and elemental 

contaminants 

Whole fish 

(composites) 

Specific 

analytes 

Schmitt et al. (1999) 

Somatic indices The relative mass of some 

organs is often indicative of 

chemical exposure 

Gonads, spleen, 

liver 

Overall 

organism 

health 

Grady et al. (1992) 

Stable N isotopes 

(

14N and 15N) 

The ratio of 15N to 14N 

(δ15N) increases with 

trophic position and sewage 

pollution 

Whole fish 

(composites) 

Trophic 

position, 

nitrogen 

sources 

Cabana and Rassmussen 

(1996) 

Necropsy-based 

fish health 

assessment 

Visual assessment of 

external/internal anomalies 

(e.g., lesions, parasites, 

tumors), which may indicate 

contaminant-related stress 

All Overall 

organism 

health 

Goede (1988, 1996); 

Adams (1990); 

Adams et al. (1993) 
METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 5

provides quantitative information on the cumulative

concentration of AHH-active PCBs, chlorodioxins,

chlorodibenzofurans, and related compounds of great

concern, and thereby augments the information on

total PCBs provided by the chemical analysis of the

fish carcasses. To assess the cumulative exposure of

the fish to all AHH-active compounds, including

PAHs, the activity of the EROD enzyme in the livers

of the individual fish was measured (Pohl and Fouts

1980). By comparing EROD results with those from

the other two endpoints, the classes of chemicals

responsible for the cumulative effect on EROD could

be deduced (Table 2). 

The following reports summarize extant

information on the biological markers and evaluate

their value to the BEST program. More in-depth

reviews and evaluations of these methods are being

prepared as a corollary activity of the BEST program

(e.g., Whyte et al. in prep.). Analytical methods for

the measurement of elemental and organochlorine

chemical concentrations in fish are well documented

and are consequently not described in this summary.

EROD ACTIVITY

Jeff J. Whyte and Donald E. Tillitt

Measurement of ethoxyresorufin-O-deethylase

(EROD) activity in fish is a well-established in vivo

biomarker of exposure to certain planar halogenated

and polycyclic aromatic hydrocarbons (PHHs and

PAHs) and other structurally similar compounds

(Bucheli and Fent 1995; Stegeman and Hahn 1994).

EROD is a highly sensitive indicator of contaminant

uptake in fish, providing evidence of receptor-mediated induction of cytochrome P450-dependant

monooxygenases (the CYP1A subfamily specifically)

by xenobiotic chemicals. Numerous laboratory experiments, simulated field studies, and natural field studies have examined EROD induction in more than 150

species of fish. In addition to PHHs and PAHs, an

extensive list of individual contaminants and complex

environmental mixtures has been examined for teleost

EROD potential. Although EROD activity is best

viewed as an indicator of exposure, the relationship

between EROD and biological effects at higher levels

of organization is the subject of intense investigation.

It is becoming clear that the mechanism of CYP1A

induction is closely related to, if not directly involved

in, detrimental effects such as apoptosis and embryonic mortality seen in fish exposed to EROD-inducing

contaminants (Cantrell et al. 1996). Apart from xenobiotic induction, EROD activity can be influenced by

a large number of abiotic and biotic factors such as

water temperature, age and reproductive phase

(Andersson and Förlin 1992). An understanding of

these factors is critical to the design and interpretation

of field studies utilizing this biomarker.

Background

Over two decades have passed since the induction of

the cytochrome P450 1A subfamily of monooxygenases (CYP1A) was proposed as a biomarker of exposure to PHHs and PAHs (Payne 1976; Payne and

Penrose 1975). The foundations for this suggestion

originated from the extensive work on this enzyme

system performed in mammals dating back to the

mid-1960s (Conney 1967; Mason et al. 1965). The

cytochromes P450 are a diverse multigene family of

heme-containing proteins that oxidize, hydrolyze, or

reduce compounds through the insertion of an atom of

atmospheric oxygen to the substrate during the reaction cycle (Nebert et al. 1993; Nelson et al. 1996). In

Table 2. Monitoring and assessment strategy for

polycyclic aromatic and polyhalogenated hydrocarbons (PAHs and PHHs).

Contaminants

PCDDs &

Endpoint PCBs PCDFs PAHs+

GC-ECD1+- -

(carcass)

EROD2 activity * * *

(liver)

H4IIE assay * * -

(carcass)3

1 total PCBs by gas chromatography with electron-capture detection

2 7-ethoxyresorufin-O-deethylase 3 after reactive cleanup to remove AhR-active PAHs

* AhR-active isomers and congeners only

+ and other planar organic compounds
6 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

fish, these enzymes are concentrated mainly in the

liver, but have been detected in the kidney, gastrointestinal tract and gill tissue (Varanasi 1989).

Embedded in the smooth endoplasmic reticulum, they

metabolize both endogenous and exogenous compounds (phase I reactions), generally increasing the

water solubility of substrates, thereby enhancing their

elimination (Andersson and Förlin 1992). In this

way, cytochromes P450 such as CYP1A tend to

detoxify xenobiotic chemicals; however, the phase I

metabolites of some PAH and other contaminants may

be more toxic than the parent compound (Guengerich

and Liebler 1985).

The most useful aspect of CYP1A for biomonitoring purposes is the enzyme’s tendency to

increase in concentration upon chemical exposure.

Induction of CYP1A is mediated through the binding

of xenobiotics to a cytosolic aryl hydrocarbon receptor (AhR) (Fig. 2). AhR ligands generally have isoteric configurations and are similar in structure to

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

a model CYP1A inducer (Fig. 3). Receptor binding is

followed by a series of molecular events leading to

the expression of several genes (including CYP1A)

known as the “Ah-gene battery” (Nebert et al. 1993).

The toxic effects of PHHs and structurally similar

compounds are thought to be mediated through the

AhR, with induced proteins causing alterations in cellular homeostasis (DeVito and Birnbaum 1994). In

mammals, these effects include wasting syndrome,

tumor promotion and thymic atrophy (Poland and

Knutson 1982). In fish, early life stages appear to be

particularly sensitive to AhR ligands (Mehrle et al.

1988; Walker and Peterson 1991), and recent evidence indicates the involvement of CYP1A enzymes

specifically in this toxic response (Cantrell et al.

1996).

The use of CYP1A induction as an assessment technique has increased in recent years. This is

due mainly to the optimization of protocols for the

rapid and relatively inexpensive measurement of its

catalytic activity as EROD (Kennedy and Jones 1994;

Burke and Mayer 1974; Pohl and Fouts 1980).

EROD induction as a biomarker in teleost species has

several advantages. By indicating the induction of

CYP1A, EROD activity provides a fingerprint of the

presence of AhR-active compounds in fish.

Historically, assessing the degree of uptake of these

compounds was complicated by both their vast number and their varying degrees of bioavailability.

Signal

Transduction

Cascade

p37

Other

Proteins

CYP1A CYP1A

Nucleus

hsp90

hsp90

hsp90

AhR

p7 hsp90

AhR

O

O

Cl

Cl

Cl

Cl

AhR

ARNT

XRE

CYP1A Other

Genes

mRNA

XRE

?

Protein 

Translation

Endogenous Ligand

or

PHH

PHH-OH

Detoxification

Toxic 

Responses

p37

O

O

Cl

Cl

Cl

Cl

hsp90

AhR

p7 hsp90 p37

O

O

Cl

Cl

Cl

Cl

O

O

Cl

Cl

Cl

Cl

Figure 2. Proposed mechanism of AhR-mediated toxicity. Signal transduction by dioxin-like ligands is mediated by the AhR, which forms a

transcription factor complex with an aryl hydrocarbon nuclear translocator protein (ARNT). This heterodimer recognizes specific DNA

sequences, dioxin responsive elements (DRE), and leads to the induction of several genes (the Ah gene battery). The elevated levels of the

protein products are thought to be involved in the toxic action of AhR ligands. 
Although analytical measurements can provide the

identities and concentrations of organic contaminants

in fish tissues (Huestis et al. 1996; Firestone 1991),

they do not render a direct indication of biological

potency. Induction of EROD is an extremely sensitive

indicator of environmental alterations and is usually

one of the first detectable, quantifiable responses to

exposure (Stegeman 1992). In addition, EROD represents the cumulative impact of all inducing chemicals,

whether or not they are detected analytically. 

Performing the EROD Assay

EROD activity describes the rate of the CYP1Amediated deethylation of the substrate 7-ethoxyresorufin (7-ER) to form the product resorufin. The

catalytic activity towards this substrate is an indication of the amount of enzyme present and is measured

as the concentration of resorufin produced per mg

protein per minute (mol/mg/min) (Kennedy and Jones

1994). Because metabolism is generally highest in

hepatic tissue, the assay is typically conducted using

fish liver. Determination of EROD activity involves

two stages. The first stage is fish capture (typically 7-

10 individuals per site) followed by the excision and

cold preservation of the liver tissue. The excised tissue is homogenized and centrifuged to isolate fragments of the endoplasmic reticulum. This microsomal fraction contains the CYP1A enzymes of interest

(Pohl and Fouts 1980). Consistency in terms of liver

section removed, cryopreservation technique and

microsomal preparation are important means of

reducing individual sample variability (Pluta 1995;

Heinonen et al. 1996). The second stage, the actual

enzymatic assay, involves providing the microsomal

fraction with 7-ER and NADPH and fluorometrically

measuring resorufin production. Samples are then

standardized based on protein content of the liver

homogenate (Lorenzen and Kennedy 1993; Lowry et

al. 1951; Bradford 1976).

Factors That Can Affect EROD Induction in Fish

As with most biological phenomena, EROD in the tissues of an organism is influenced by a variety of

internal, external, and temporal factors (Bucheli and

Fent 1995). Biological factors that can influence

EROD activity include species (Addison et al. 1991;

Segner et al. 1995), fish size and age (Peters and

Livingstone 1995; Pluta 1993), and reproductive status (Campbell et al. 1976; Schreck and Hopwood

1974). The physical treatment of fish in both laboratory and field studies can also greatly affect EROD

measurements. Careful consideration should be given

to contaminant exposure route (James and Bend 1980;

Haasch et al. 1993), fish diet (Jimenez and Burtis

1988), and the use of anesthesia during capture

(Kleinow et al. 1986). Environmental variables such

as temperature and pH can drastically affect induction

of EROD and should be routinely measured throughout a study (Andersson and Koivusaari 1985;

Sleiderink et al. 1995; Willis et al. 1991).

Contaminant exposure period and study duration are

important in both laboratory and caged-fish studies

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 7

Cl

Cl

Cl

Cl

Cl

O

O

Cl

Cl

Cl

Cl O

Cl

Cl

Cl

Cl

(a) 2,3,7,8-Tetrachlorodibenzo- p-dioxin (b) 2,3,7,8-Tetrachlorodibenzofuran

(c) 3,3 ́,4,4 ́,5-Pentachlorobiphenyl (PCB 126) (d) Benzo[a]pyrene

Figure 3. Representative AhR ligands. The molecules demonstrate the general structure of compounds in the following classes: (a) polychlorinated dibenzo-p-dioxins (PCDDs), (b) polychlorinated dibenzofurans (PCDFs), (c) polychlorinated biphenyls (PCBs) and (d) polycyclic

aromatic hydrocarbons (PAHs). 
8 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

examining EROD (van der Weiden et al. 1994b;

Sleiderink and Boon 1996).

A variety of chemicals and chemical mixtures are known to inhibit the induction of EROD in

fish. These include organic, organometallic, and

metallic compounds such as specific polychlorinated

biphenyl (PCB) congeners (Gooch et al. 1989;

Newsted et al. 1995), and organotins (Bucheli and

Fent 1995). In addition to antagonistic chemicals,

there a variety of AhR agonists that are of biogenic

nature, including plant metabolites and biotoxins

(Takahashi et al. 1995). In short, the presence or

absence of EROD activity in fish from a site may not

always represent contamination by or lack of traditional AhR agonists, and study designs incorporating

a battery of tests for biological responses to contaminants will likely yield more concrete information.

Value and Utility of EROD in the BEST Program

As a monitoring tool, EROD activity provides a relatively rapid indication of toxic planar compound

uptake in fish. For this reason, EROD is often termed

an “early warning system” (Payne et al. 1987). Thus

far, hundreds of field studies have employed the

EROD assay to determine spatial and temporal trends

of contamination in aquatic systems (Kennedy and

Jones 1994; Balk et al. 1993; Teal et al. 1992; Adams

et al. 1996; Achazi and Leydecker 1992). These studies have detected EROD induction in many species of

fish from a broad range of habitat types. This extensive validation of EROD as a biomarker has led to its

use in several contaminant monitoring programs

worldwide [e.g., North Sea Task Force, (Stagg 1991;

Pluta 1995), Environment/Cellulose (Förlin et al.

1995), Mediterranean Pollution Network (Burgeot et

al. 1996), French National Observation Network

(Godefroy et al. 1996)]. 

Ideally, a biomarker will exhibit a relationship to toxicity in the organism(s) being examined.

Induction of CYP1A (estimated from EROD activity),

while not a toxic response per se, does indicate the

potential for AhR ligands to induce biochemical

change. The generation of reactive PAH intermediates by CYP1A has long been known as a source of

DNA adducts that can lead to carcinogenesis [Fig. 4,

Ah

Receptor

Induction of Ah

Gene Battery

Reactive

Intermediates

Perturbation of Differentiation

Apoptosis

Epithelial Hyperplasia

Clonal expansion

of initiated cells

Tumor Promotion

and Progression

Excretion of

Conjugated Products

DNA Adducts Mutation

Fixation

Oncogene

Activation

Tumor 

Initiation

Ah

Receptor

Induction of Ah

Gene Battery

Reactive

Intermediates

Perturbation of Differentiation

Apoptosis

Epithelial Hyperplasia

Clonal expansion

of initiated cells

Tumor Promotion

and Progression

Excretion of

Conjugated Products

DNA Adducts Mutation

Fixation

Oncogene

Activation

Tumor 

Initiation

Ah

Receptor

Induction of Ah

Gene Battery

Reactive

Intermediates

Perturbation of Differentiation

Apoptosis

Epithelial Hyperplasia

Clonal expansion

of initiated cells

Tumor Promotion

and Progression

Excretion of

Conjugated Products

DNA Adducts Mutation

Fixation

Oncogene

Activation

Tumor 

Initiation

Figure 4. Ah receptor-mediated formation of DNA adducts and tumor initation upon exposure to benzo[a]pyrene. Compounds such as

2,3,7,8-TCDD are also capable of perturbing differentation, affecting apoptotic balance and causing tissue-specific proliferation, another

hypothesized route of AhR-mediated carcinogensis (From Nebert et al. 1993; reproduced by permission). 
(Varanasi et al. 1989)]. In addition, a variety of

parameters in fish have been associated with the

induction of EROD activity including reproductive

effects (e.g. reduced serum steroid levels), increased

liver somatic index and mortality (Watson and Di

Giulio 1997; van der Weiden et al. 1994a; Levine et

al. 1995; Gagnon et al. 1994). At present the many

symptoms and effects that AhR ligands exert on a

wide variety of biochemical systems have made it difficult to identify the critical events that lead to toxicity. Until a more direct linkage between CYP1A

induction and detrimental effects in fish is established, EROD activity is best viewed as an indicator

of contaminant exposure rather than of effect. As

more information is gathered on the relationship

between EROD activity and detrimental effects in

fish, this biomarker may also serve as a predictive

tool for contaminant risk assessment. In the meantime, EROD induction serves as an economical

assessment tool for environmental managers, allowing

them to prioritize areas for examination using more

expensive techniques such as congener specific analysis of tissue and sediment.

By employing the teleost EROD induction

technique with other methods (e.g. H4IIE bioassay,

instrumental analysis), an integrated assessment of the

contaminant levels and identities in fish will be provided. This multilevel approach renders the best balance of information at the minimal cost.

H4IIE BIOASSAY

Jeff J. Whyte and Donald E. Tillitt

The H4IIE rat hepatoma cell line bioassay (H4IIE

bioassay) is an in vitro test used to detect and semiquantify specific contaminants and classes of contaminants chemically extracted from environmental

matrices such as sediment, water, and organisms

(whole or specific tissues). The H4IIE bioassay

measures the catalytic activity of cytochrome P4501A

(CYP1A), a mixed-function oxidase (MFO) enzyme,

as 7-ethoxyresorufin-O-deethylase (EROD) activity in

cultured rat liver cells exposed to environmental

extracts. EROD is induced by, and the H4IIE bioassay is consequently useful for characterizing, the presence of certain polycyclic aromatic hydrocarbons

(PAH) and related compounds (e.g., nitrogen heterocyclics and sulfur-, oxygen-, nitro-, amino,- and alkylsubstituted PAH) and polyhalogenated hydrocarbons

(PHH) in environmental samples. The PHHs include

the highly toxic and persistent polychlorinated dioxins

(PCDDs), dibenzofurans (PCDFs), biphenyls (PCBs),

and naphthalenes, as well as the brominated analogs

of these compounds. These classes of compounds

induce CYP1A and hence EROD activity in cells by

binding to the cytosolic aryl hydrocarbon receptor

(AhR). This AhR-mediated mechanism of EROD

induction is believed to be involved in many of the

toxic effects associated with PHHs and PAHs (Poland

and Knutson 1982) (Fig. 3). An overview of EROD

induction is presented in “EROD Activity.”

The H4IIE bioassay has advantages over traditional analytical chemistry techniques in that it

reveals the cumulative biological activity of numerous

structurally similar contaminants, each with differing

potencies. This assay can also reveal the potential

interactions that can occur between contaminants

when they are present in environmental samples as

complex mixtures. The H4IIE bioassay is valuable

for environmental monitoring purposes because it

enables the assessment and ranking of the potential

toxicity of samples based on their ability to induce

EROD as a surrogate for analytical determination of

specific compounds. When based on tissue samples,

cumulative H4IIE-derived potency estimates can be

used to assess the risk to the organism(s) from which

the extract was obtained. Such estimates can also be

used to estimate the contaminant burden or dose that

the organism could contribute either to higher trophic

levels (via the food chain) or to its progeny (via

maternal transfer). The H4IIE bioassay has a high

degree of sensitivity (detection limit < 10 femtomoles

2,3,7,8- tetrachlorodibenzo-p-dioxin [TCDD]) and

can be rapidly performed. These characteristics make

the H4IIE bioassay an ideal tool for evaluating the

toxic potential of the samples collected for monitoring

programs.

Background

The continuous cell line, H4IIE, was derived from the

Reuber Hepatoma H-35 (Reuber 1961) by Pitot and

coworkers (Pitot et al. 1964). A decade later, induction of CYP1A catalytic activity in the cell line was

demonstrated (Benedict et al. 1973; Niwa et al. 1975).

The H4IIE cells are well-suited for the examination of

EROD induction due to their excellent growth characteristics and the presence of low basal, but highly

inducible, CYP1A activity. The H4IIE cell line is

exquisitely responsive to 2,3,7,8-TCDD-based

CYP1A induction. These characteristics prompted

researchers from the U.S. Food and Drug

Administration to develop and characterize a contaminant detection bioassay based on the H4IIE cell line

(Bradlaw and Casterline 1979). This original assay

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 9
was used to screen for the presence of CYP1A-inducing chemicals in foodstuffs as indicated by aryl

hydrocarbon hydroxylase (AHH) activity, a catalytic

measure of CYP1A (Trotter et al. 1982; Casterline et

al. 1983). The assay was subsequently modified to

examine EROD activity rather than AHH activity

(Sawyer and Safe 1982) because the EROD assay

employs a non-toxic substrate.

Use of the H4IIE assay to rank the toxic

potency of individual chemicals based on the 2,3,7,8-

TCDD equivalency (TEQ) concept was first proposed

by Safe (1987). The TEQ concept was subsequently

expanded and used to demonstrate the cumulative

toxicity of mixtures of PAHs and PHHs in controlled

laboratory studies and in extracts of environmental

samples. TEQ values generated by the H4IIE assay

provided a relative toxicity estimate for individual

chemicals. The values can also be used together with

analytical chemistry to evaluate the potential interactions of mixtures of CYP1A-inducing chemicals in

biological systems. More recently, environmental

assessments using the H4IIE bioassay have become

more prevalent, mainly due to the systematic characterization of the assay by Tillitt et al. (1991).

Additional modifications that have improved the

bioassay (e.g., microtiter plates, use of live cells) have

also been introduced over the past decades (Tysklind

et al. 1994; Whyte et al. 1998; Bradlaw et al. 1982;

Donato et al. 1992; Munkittrick et al. 1993).

Performing the H4IIE Bioassay

Conduct of the H4IIE bioassay can be divided into

three general stages (Fig. 5): 1) Extraction, cleanup,

fractionation, and analytical characterization of contaminants in the tissue or other environmental matrix;

2) preparation of an extract dilution series with which

the cells are dosed; and, 3) measurement of EROD

activity and calculation of TEQs (expressed as pg

2,3,7,8-TCDD/g tissue). Extraction, cleanup, and

fractionation methods depend on the contaminant

classes of interest, but generally involve isolation of

hydrophobic planar compounds such as PHHs (Feltz

et al. 1995; Huestis et al. 1995). Whole extracts can

be examined for general inducing potency.

Alternatively, fractionated extracts can provide more

detailed information on the contribution of different

classes of contaminants (e.g., PCBs, PCDDs/Fs)

(Gale et al. in press). Analytical detection techniques

such as GC/MS or HPLC can identify specific chemicals in extracts potentially responsible for EROD

induction in H4IIE cells, but the cost of this analysis

is avoided if the bioassay is being used simply to

indicate and quantify the cumulative presence of

EROD-inducing compounds in extracts.

Prior to dosing the H4IIE cells, isolated contaminants are transferred into a solvent carrier suitable for delivery to the cultured cells (e.g., isooctane

or DMSO) and a logarithmic dilution series is prepared (Tillitt et al. 1991; Smith et al. 1994). H4IIE

cells, cultured under conditions described by Tillitt et

al. (1991), are seeded in microtiter plates and grown

for 24 h. Cells are dosed with extracts or a 2,3,7,8-

TCDD standard, incubated (24 – 72 h) and assayed

for EROD using a modified method of Pohl and Fouts

(1980). The microtiter plate method allows for measurement of both EROD and protein in the same wells

(Kennedy and Jones 1994). At the final stage of the

assay, concentration-response curves are used to

determine relative extract potencies (TEQs) by comparing EC50 or slope values of the extract to those of

2,3,7,8-TCDD (Mason et al. 1985; Tillitt et al. 1993).

Factors That Can Affect the H4IIE Bioassay

Because the H4IIE bioassay is a laboratory assay,

many of the external modifying factors that can influence in vivo measurement of EROD in fish (e.g., sex,

species, ambient temperature) do not influence EROD

induction in H4IIE. Therefore, deviations from specified experimental conditions are the most likely

source of variability in the H4IIE bioassay. Similar to

the in vivo assay, variables such as reagent temperature and pH, and resorufin and ethoxyresorufin purity

may influence EROD measurements. Other factors

specific to cell culture (e.g., cell passage number,

mycoplasma contamination) may also influence

EROD measurements in H4IIE.

The assessment of environmental extracts

with the H4IIE bioassay may also be affected by the

presence of specific compounds in the mixture (e.g.,

certain PCBs). This is mainly caused by inhibition of

the catalytic activity of CYP1A in H4IIE and can lead

to erroneously reduced EROD measurements (Sawyer

et al. 1984). An elegant solution to this problem was

presented by Denison and coworkers (Garrison et al.

1996), who transfected a luciferase gene that is

expressed upon AhR-complex binding to DNA in

H4IIE cells. This results in induction of luciferase

activity, which produces luminescence, upon exposure

to CYP1A-inducing compounds. The luciferase

activity is not influenced by substrate inhibition,

resulting in enhanced sensitivity to AhR ligands and

greater confidence in induction measurements. It is

yet to be determined whether this modified technique

will replace the traditional H4IIE bioassay.

10 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
Excision of Tissue

Contaminant

Extraction

Analytical Detect

Serial Dilution of Extract

TCDDStandardsExtractDilutions

Cell Dosing and 

Incubation

N

C2H5O O O

HO 

7-Ethoxyresorufin 

7-Ethoxyresorufin-O

(CYP1A1

Fluorometric deter

of EROD Act

EROD Reac

2,3,7,8-TCDD

1 10 100

EROD Activity

0

20406080

100120

EC50

Tissue Extract

0 0.1 1

EROD Activity

0

20406080

100120

EC50

Dose-response Curve Comparison

For TEQ determination (pg TCDD/g tissue)

H4IIE Bioassay

Figure 5. Application of the H4IIE cell line bioassay to assess the toxic potency of environmental extracts. Extract EROD-inducing potency in the cells is compared to th

TCDD equivalent (TEQ) concentration (pg TCDD/g tissue). 
Value and Utility of the H4IIE Bioassay in the BEST

Program

The persistence and demonstrated toxic effects of

PHHs are of great concern to the U.S. Fish and

Wildlife Service and other natural resource management agencies (U.S. Fish and Wildlife Service 1993).

The H4IIE bioassay is a rapid and highly sensitive

test that can indicate the presence of PHHs and related compounds in environmental samples and in some

situations can estimate the potential cumulative toxicity of these compounds to organisms. The bioassay

can also provide information on the biological interactions of PHHs when used in concert with analytical

chemistry methods. The H4IIE bioassay is also of

value because of its ability to relate the EROD-inducing strength of PHHs to potential toxic impacts in

whole organisms. Relationships between CYP1A

enzymatic activities in H4IIE cells and deleterious

effects in live rats have been demonstrated for individual PCDD (Bradlaw et al. 1980), PCDF (Mason et

al. 1985; Bandiera et al. 1984), and PCB (Sawyer and

Safe 1982; Leece et al. 1985) congeners. The strong

correlation observed between AHH induction by these

congeners in cultured cells and weight loss or thymic

atrophy in whole organisms supports the use of the in

vitro bioassay as a tool for predicting toxic potency

(reviewed by Safe 1990).

Complex mixtures of PHHs have also been

examined using the H4IIE bioassay and have been

used to assess toxic impacts in wildlife. For example,

hatching success of double-crested cormorants

(Phalacrocorax auritis) in Great Lakes colonies was

strongly correlated with TEQ results from the H4IIE

bioassay (Fig. 6), whereas conventional contaminant

analysis correlated poorly with hatching success of

eggs from the same colony (Tillitt et al. 1992).

Mammalian studies have also revealed the utility of

the H4IIE bioassay to predict toxic effects. Mink

(Mustela vison) fed a diet containing increasing percentages of fish from a contaminated bay in the Great

Lakes showed reproductive effects that were related

to results from the H4IIE bioassay (Tillitt et al. 1996).

By taking into account the differing biological potencies of the chemical components in an extract, the

H4IIE bioassay is a good predictor of the toxic outcome of exposure to the extract.

The H4IIE bioassay is among the Tier I or

“workhorse methods” recommended for screening

environmental samples for dioxin-like activity in virtually all habitats and guilds likely to be sampled by

the BEST program (BEST 1996). The assay can

detect extremely low concentrations of contaminants

that are a high priority for the BEST program (PHHs

and PAHs; U.S. Fish and Wildlife Service 1993).

Used in conjunction with other methods (e.g., in vivo

measurement of hepatic EROD and instrumental

analysis of total PCB residues in the carcasses of the

fish), the H4IIE bioassay can provide information on

the class or classes of contaminants in the samples

inducing the MFO system without the expense of

high-resolution chemical analyses. It should be noted

that the in vivo and in vitro measures of EROD activity provide similar but distinctly different information.

Hepatic EROD in fish indicates the presence of compounds that have already interacted with the AhR,

whereas the H4IIE bioassay reveals contaminants

accumulated in tissues that have the potential to bind

to the AhR. Use of these two assays in concert and

with specific fractionation schemes can yield critical

information regarding the presence of CYP1A inducers that are easily metabolized or otherwise non-persistent. Perhaps most importantly, and in contrast to

many biomarkers, TEQs in biological samples as estimated by the H4IIE bioassay indicate the presence of

compounds that are known to exert toxic effects

through a similar mode of action. Although questions

have been raised about species-specific differences in

response to PHH and PAH congeners (Clemons et al.

1997), extant information suggests that the mechanism of AhR-mediated CYP1A induction is similar

among vertebrates (Stegeman and Hahn 1994), supporting the use of H4IIE in the assessment of toxic

potency of contaminants accumulated in organisms

such as fish. These attributes give the H4IIE bioassay

predictive power in terms of risk to organisms.

12 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Dioxin equivalents, parts per trillion (pg/g)

Egg mortality (%)

Figure 6. Hatching success in Great Lakes colonies of doublecrested cormorants versus PCBs in eggs as dioxin-equivalent concentrations, 1987-1988 (modified and redrawn from Tillitt et al. 1992

by permission); each point represents the mean for a colony.
CONDITION FACTOR AND ORGANOSOMATIC INDICES

Gail M. Dethloff and Christopher J. Schmitt

Measurements of condition factor, which relates

weight to length, and organo-somatic indices, which

indicate the proportional sizes of certain organs, are

standard procedures in fish physiology studies and in

fisheries biology. The condition factor is an organism-level response, with factors such as nutritional

status, pathogen effects, and toxic chemical exposure

causing greater-than-normal or less-than-normal

weights. Organo-somatic indices reflect the status of

organ systems, which may change in size due to environmental factors more rapidly than organism weights

and lengths increase or decrease. Both the condition

factor and organo-somatic indices are used as indicators of the well-being of individual organisms.

Background

Condition factor and organo-somatic indices have

been used extensively in fish health and population

assessments (reviewed by Goede and Barton 1990;

Hoque et al. 1998). Because it integrates many levels

of sub-organismal processes (e.g., molecular, cellular,

organ system), an index such as Fulton’s condition

factor (Carlander 1969) may signify the overall condition and nutritional status of individual fish (Adams et

al. 1992a). The size or weight of the liver, spleen,

and gonads relative to fish length or weight may also

signify overall health and reproductive status. The

organo-somatic indices are generally expressed as

percentages of total body weight. The hepato-somatic

index (HSI) is the weight of the liver expressed as a

percentage of total body weight; it is also known as

the liver somatic index. Gingerich (1982), in summarizing the extensive literature on the biology of the

fish liver, reported that the liver constitutes, on average, about 2% of body weight in mature teleost fishes. Alterations in liver size may reflect changes in the

metabolism and energy reserves of an individual fish

(Busacker et al. 1990). The spleno-somatic index

(SSI) is the weight of the spleen expressed as a percentage of total body weight. Alterations in this

index could indicate an abnormal condition in the

spleen such as necrosis or swelling due to infection

(Goede and Barton 1990). The gonado-somatic

index, the weight of the gonads expressed as a percentage of total body weight, is discussed in

“Reproductive Indicators.”

Measuring Condition Factor and Organo-somatic

Indices

Measurements of condition factor and organo-somatic

indices require only measuring boards and balances of

the appropriate capacity to measure length of whole

fish and to weigh whole fish and organs. To avoid

moisture losses and gains and corresponding weight

changes during storage (i.e., freezing or preservation),

weight determinations should be made on live or

freshly killed specimens (Busacker et al. 1990).

Condition factor is generally computed as

weight/length3. This equation reflects the expected

exponential gain in weight relative to length as fish

grow. Depending on the units of measurement, a constant may be included to bring the index values near

unity (Anderson and Gutreuter 1992). Relationships

between length and weight also may be estimated by

regression analysis (length-weight equations)

(Carlander 1969). Organo-somatic indices are often

calculated as (organ weight/body weight) * 100

(Busacker et al. 1990). Alternatively, these relationships may be documented statistically (i.e., as regression coefficients of the relationships) (Delahunty and

de Vlaming 1980; Grady et al. 1992). If ratios are

calculated, they can be problematic because they do

not conform to any common statistical distributions.

An alternate approach to the appraisal of relative

organ size and weight relative to length is through a

procedure such as the analysis of covariance, whereby

the nature of the relations between the variables is

examined (Delahunty and de Vlaming 1980; Grady et

al. 1992). However, the requirement of more-or-less

equal-size fish to meet the assumptions for analysis of

covariance may preclude this option for some comparisons.

Factors That Can Affect Condition Factor and Organosomatic Indices

In general, condition factor varies directly with nutrition (Tyler and Dunn 1976). A negative correlation

has been seen between disease and condition in fishes

(Möller 1985). Condition factor may vary in either

direction outside the normal range in response to

chemical exposure. Elevated condition factors have

been found in white sucker (Catostomus commersoni)

and redbreast sunfish (Lepomis auritus) at sites polluted with pulp mill effluents (Adams et al. 1992a;

McMaster et al. 1991). Decreased condition factors

have been seen in white sucker at sites with elevated

concentrations of metal mixtures and in Atlantic cod

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 13
(Gadus morhua) exposed to petroleum (Munkittrick

and Dixon 1988; Miller et al. 1992; Kiceniuk and

Khan 1987). Nutrition, disease and contaminants are

highly inter-related in terms of their effects on fish

condition. Insufficient nutrition can lead to higher

disease susceptibility (Klontz 1985) and thus altered

condition factor. There is evidence that contaminants

may potentiate disease outbreaks (Sindermann 1990;

Möller 1985) and that they can alter food resources

(Munkittrick and Dixon 1988). Incidence of disease

or poor food resources can then manifest as lowered

condition factor.

Condition factor may also vary seasonally

(Griffiths and Kirkwood 1995; Saborowski and

Buchholz 1996), possibly due to changes in food

availability or metabolism, and with changes in

gonadal status (Chellappa et al. 1995). As one might

expect, condition factor varies greatly among fish taxa

owing to their differential architecture, but condition

indices can also vary from location to location within

a species (Doyon et al. 1988; Fisher et al. 1996). A

final matter to consider when using condition factor

to assess fish health is that a decrease in weight due

to loss of energy stores can be offset by an increase in

body water (Goede and Barton 1990).

Because of the energy storage and metabolic

functions of the liver, alterations in liver size due to

environmental stressors are of interest. Evaluation of

the HSI must consider the role of both endogenous

and exogenous factors. The HSI varies with seasonal

cycles (Saborowski and Buchholz 1996; Delahunty

and de Vlaming 1980; Beamish et al. 1996; Slooff et

al. 1983). Because of the liver’s role in storage and

metabolism, nutritional quality and regimes also

affect relative liver size (Daniels and Robinson 1986;

Foster et al. 1993; Heidinger and Crawford 1977;

Swallow and Fleming 1969). The HSI can also vary

with sex and changes in gonadal status (Grady et al.

1992; Fabacher and Baumann 1985; Förlin and Haux

1990). In females, the HSI may change as the gonado-somatic index changes due to the liver’s role in

vitellogenesis (Scott and Pankhurst 1992).

Like the condition factor, the HSI is constrained by the allometry of the population, the

species, or both (Grady et al. 1992). The form of the

allometric relation between liver weight and body

weight varies widely among species, both positively

and negatively (Grady et al. 1992; Yakoleva et al.

1976). The fish liver may also store blood during

periods of quiescence (Gingerich 1982), which suggests that the activity of the fish immediately prior to

capture and the protocol used to procure the liver may

affect relative liver size. Factors that cause a disproportionate change in body weight will also affect the

HSI.

Of the organo-somatic indices, the HSI is the

one most often associated with contaminant exposure

(Adams and McLean 1985). Several investigators

have suggested that relative liver enlargement in fish

indicates exposure to environmental carcinogens or

other toxic chemicals (Table 3). Increased HSI has

been reported in brown bullheads (Ameiurus nebulosus) from sites polluted with polycyclic aromatic

hydrocarbons (PAHs) (Fabacher and Baumann 1985;

Gallagher and Di Giulio 1989), in rainbow trout

(Oncorhynchus mykiss), Atlantic cod, and winter

flounder (Pleuronectes americanus) exposed to waters

containing a mixture of PAHs and other pollutants

(Poels et al. 1980; Kiceniuk and Khan 1987; Fletcher

et al. 1982), and in redbreast sunfish exposed to

industrial discharge containing PAHs and polychlorinated biphenyls (PCBs) (Adams et al. 1989).

Additionally, striped bass (Morone saxatilis) exposed

to Hudson River water contaminated with PCBs,

hard-head catfish (Arius felis) from PAH-contaminated sites, and European plaice (Pleuronectes platessa)

collected from sites contaminated with sewage sludge

all had enlarged livers (Buckley et al. 1985; Everaarts

et al. 1993; Secombes et al. 1995). Fabacher and

Baumann (1985) and Gallagher and Di Guilio (1989)

concluded that increased xenobiotic metabolism in

fish was achieved by enlarging the liver (and therefore increasing the HSI) rather than by increasing the

specific activity of the detoxification enzymes.

In contrast with the studies described above,

a number of laboratory studies found that liver size

decreased following exposure to contaminants (Table

3). Exposure of rainbow trout to sodium pentachlorophenate (Hickie and Dixon 1987) caused a

reduction in the HSI, as did exposure of perch (Perca

fluviatilis) to a mixture of metals (Larsson et al.

1984); Atlantic salmon (Salmo salar) to cyanide

(Ruby et al. 1987); Asian redtail catfish (Mystus

nemurus) to hydrogen sulfide (Hoque et al. 1998);

and striped mullet (Mugil cephalus) to crude oil

(Chambers 1979). These decreases may have reflected glycogen loss in the liver as energy stores were utilized (Barton et al. 1987). Exposure to carbofuran

also decreased HSI in the green snakehead (Channa

punctatus); the decrease was linked to histopathological changes in the liver, including hepatocyte damage

and degeneration (Ram and Singh 1988). Field studies investigating bleached kraft mill effluent (BKME)

effects have found significantly lower HSI in exposed

fish, though seasonal cycles may also have contributed to the decrease in liver size (Adams et al.

1992a; McMaster et al. 1991). Adams et al. (1992a)

attributed lower HSI to altered carbohydrate metabo14 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
Table 3. Investigations in which the hepatosomatic index (HSI) has been evaluated as a biomarker of contaminant exposure. Species TSN1 Location Contaminant(s) Result Field StudiesBrown bullhead (Ameiurus nebulosus) 164043 Black River, Ohio PAH Increase 

North Carolina PAH Increase 

Hardhead catfish (Arius felis) 164165 Gulf of Mexico PAHs, cadmium Increase Bream (Abramis brama) 163666 European Rivers Multiple Increase Rainbow trout (Oncorhynchuss mykiss) 161989 Rhine River Multiple Increase 

Finland Paper mill effluent No effect 

White sucker (Castostomus commersoni) 163895 Canada Bleached kraft mill effluent Decrease Striped bass (Morone saxatilis) 167680 Hudson River Multiple (PCBs) Increase Redbreast sunfish (Lepomis auritus) 168131 East Tennessee PCBs, PAHs Increase 

East Tennessee Bleached kraft mill effluent Decrease 

European plaice (Pleuronectes platessa) 172902 Scotland Sewage sludge Increase Laboratory studiesRainbow trout (Oncorhynchuss mykiss) 161989 NA Acidity Decrease 

NA Sodium pentachlorophenate Decrease 

Atlantic salmon (Salmo salar) 161966 NA Cyanide Decrease 

Perch (Perca fluviatilis) 168470 NA Metals (mixture) Decrease 

Atlantic cod (Gadus morhua) 164712 NA Crude oil Increase 

Winter flounder (Pleuronectes americanus) 172904 NA Crude oil Increase 

Brown bullhead (Ameiurus nebulosus) 164043 NA Benzo(a)pyrene No effect 

Asian redtail catfish (Mystus nemurus) NA Hydrogen sulfide Decrease 

Striped bass (Morone saxatilis) 167680 NA Arsenic Mixed 

Striped mullet (Mugil cephalus) 170335 NA Crude oil Decrease 

Green snakehead (Channa punctatus) 166666 NA Carbofuran Decrease 

1Taxonomic serial number from the Interagency/International Taxonomic Information System (ITIS) database
lism. 

Spleen size is considered a useful diagnostic

factor because the spleen is a hematopoietic organ

(Anderson 1990) and dysfunction could have effects

at the whole-organism level. The SSI has not been as

thoroughly investigated as the HSI, but certain

endogenous and exogenous factors are known to

affect it. The range of spleen sizes varies among fishes (Anderson et al. 1982; Ruklov 1979) and among

populations of the same species (Lipskaya and

Salekhova 1980; Ruklov 1979). Relative spleen

weight may also differ with gender, age, size, gonadal

development, and growth rate (Krykhtin 1976;

Ruklov 1979). Seasonal changes also affect the SSI

(White and Fletcher 1985). Finally, as with the HSI,

factors that cause a disproportionate change in body

weight will affect the SSI.

Nonspecific stressors (e.g., hypoxia) can

result in altered spleen morphology. Studies on six

species of teleost fish found that transient hypoxic

conditions or severe exercise caused the spleen to

contract fully and then decrease in size and hemoglobin content (Yamamoto and Itazawa 1985; Yamamoto

1988). Acute stressors including increased temperature, exhaustive exercise, hypoxia, and simulated

transport led to spleen contraction and decreased

spleen mass in young European sea bass

(Dicentrarchus labrax), dab (Limanda limanda), and

in Trematomus bernachii, a benthic Antarctic teleost

(Hadj-Kacem et al. 1987; Davison et al. 1994;

Pulsford et al. 1994). Alterations in spleen morphology due to nonspecific stressors are paralleled by alterations at the cellular level: the release of erythrocytes

into circulation, a decrease in the total numbers of

intact white blood cells, an increase in the proportion

of atypical cells (erythrocytes and macrophage-like

cells), and enhanced red blood cell degradation in the

spleen (Yamamoto and Itazawa 1983; Maule and

Schreck 1990; Peters and Schwarzer 1985).

Histological changes (edema, necrosis, hyperaemia)

were seen in the spleens of channel catfish (Ictalurus

punctatus) held under conditions of sublethal hypoxia

for 24, 48, and 72 h (Scott and Rogers 1980).

Enlargement or swelling of the spleen, on the other

hand, is considered to be indicative of disease or

immune system problems (Goede and Barton 1990).

This condition may be due to the increased hypertrophy or proliferation of leukocytes (Anderson 1990). 

Chemical contaminants can also alter the

SSI. A trend of elevated SSIs was seen with chronic

BKME exposure of redbreast sunfish (Adams et al.

1992a). Decreased SSIs occurred in cunners

(Tautogolabrus adspersus) exposed to petroleum for

six months, in Atlantic cod exposed to Venezuelan

crude oil for 21 or more days, and in gobies

(Zosterisessor ophiocephalus) residing at a polluted

site (elevated PCBs, PAHs, metals) in the Venice

Lagoon (Payne et al. 1978; Kiceniuk and Khan 1987;

Pulsford et al. 1995). Juvenile rainbow trout exposed

for 24 h to a component of BKME experienced significant decreases in the SSI and hemoglobin concentration. A significant increase in the SSI and leukocrit, and significantly higher cumulative mortality

after disease challenge were seen after a 25-d exposure (Johansen et al. 1994). Histological data show

cellular changes occurring in the spleen with exposure

to contaminants, supporting the use of the SSI as a

relevant indicator of spleen dysfunction. Chronic

exposure of rainbow trout to bis(tri-n-butyltin) oxide

resulted in a concentration-related splenic lymphocyte

depletion. Reticuloendothelial cells proliferated in the

spleen, suggesting an increased need for phagocytes

to remove damaged blood cells, and increased erythrophagia was noted (Schwaiger et al. 1992).

Lymphoid cell depletion and cell necrosis in the

spleen have been recorded in rainbow trout exposed

to the fungicide triphenyltinacetate, redbreast sunfish

exposed to a mixture of PCBs and metals, and rainbow trout held in Rhine River water (elevated levels

of PAHs, chlorinated hydrocarbons, and metals)

(Schwaiger et al. 1996; Teh et al. 1997; Poels et al.

1980). Chronic tetrachlorodibenzo-p-dioxin exposure

caused splenic lymphoid depletion and overall splenic

atrophy in rainbow trout, though recovery seemed to

occur with removal to clean water (Fisk et al. 1997).

Certain contaminants can affect organs such as the

spleen directly (size and function) or they can suppress immune system functions (Anderson et al. 1989;

Hutchinson and Manning 1996a), increasing disease

prevalence and thus causing enlargement of the

spleen. 

Condition factor, the HSI, and the SSI appear

to be useful indicators of fish health; however, these

indices must be interpreted with caution. Potentially

confounding factors need to be recognized when one

uses such indices to compare groups of fish for the

effects of contaminants. A primary restriction is that

the condition factor and organo-somatic indices

should only be compared within a species or

between/among similar species. Also, the study protocol must be consistent and conservative in terms of

sampling, particularly since the SSI and HSI can be

altered within minutes by capture and holding stress.

The nature of the indices may also restrict their use.

For example, the HSI can only be measured on fish

with discrete livers. It is simply not practical to dissect and weigh all of the liver pieces for fishes with a

dispersed liver [e.g., common carp (Cyprinius car16 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
pio)] in the field. A solution to the problem of gonad

size varying with season and thus affecting other

indices was proposed by Grady et al. (1992), who

suggested that liver size was better expressed relative

to fish size exclusive of the gonads. This recommendation, which is also applicable to the condition factor

and SSI, was incorporated into BEST protocols being

evaluated in the Mississippi River, Columbia River,

and Rio Grande basins (Schmitt et al. 1995; Bartish et

al. 1997)

Value and Utility of Condition Factor and Organosomatic Indices in the BEST Program

Although quite crude compared to many biomarkers,

condition factor and organo-somatic indices are

important to the BEST program. As general indicators of the overall health and well-being of the fish,

alterations in these indices may indicate deleterious

effects resulting from exposure to chemicals or classes of chemicals. Some of these chemicals may not

even be recognized as threats because they remain

undetected by current water quality monitoring protocols. The indices also integrate, at the organ system

and organism level, the combined effects of multiple

contaminants and the combined effects of contaminants and other stressors. A decrease in condition

factor, the HSI, or both is considered a reflection of

depletion in energy reserves (Goede and Barton 1990;

Barton et al. 1987) because these indices are positively related to total muscle and liver energy content

(Lambert and Dutil 1997). A logical link then exists

between this depletion of energy reserves and potential health problems for fish. An increase in both the

condition factor and HSI can, however, also signal the

deleterious effect of a stressor. Although the general

interpretation is that a greater weight relative to

length indicates a healthier condition (for the individual organism), the presence of fewer, larger, and more

robust individuals may signify an out-of-balance or

abnormal condition at the population or community

level (Wege and Anderson 1978; Lehtonen and

Jokikokko 1995). An increase in the HSI suggests

increased metabolism of xenobiotics (Fabacher and

Baumann 1985) and may also be linked to pathological damage (Hinton and Laurén 1990). Depending on

the relation of the condition factor and HSI to the specific energy reserves in a fish species (e.g., muscle

protein, liver lipid), the condition factor and HSI may

respond at different rates to environmental conditions,

reflecting the different mobilization or accumulation

rates of different types of stored energy. The use of

condition factor and the HSI could thus provide

insight into shorter- and longer-term responses of

fishes to stressors (Lambert and Dutil 1997).

Although differences in the HSI may indicate changes

on a shorter time scale than condition factor, data

from a number of studies (Grady et al. 1992; Oikari

and Niittylä 1985; Holm et al. 1994) still support the

use of the HSI as an indicator of chronic rather than

acute or recent pollution.

Condition factor is also considered a useful

indicator in monitoring because baseline data are

available for comparison; length-weight equations and

condition factors for many of the North American

fishes likely to be sampled by the BEST program

have been tabulated (Carlander 1969; Carlander

1977). Data on condition factor can also be used to

compare growth of sampled fishes against established

empirical standards (Goede and Barton 1990).

Finally, condition factor is a recognized indicator of

fish health that is being used to monitor fish populations in other national programs (Bulger et al. 1995).

The SSI is of interest due to the spleen’s

hematopoietic function which also makes it an

immune system organ. Alterations in relative spleen

size could signal a dysfunction capable of affecting

individual health. Decreased size has often been seen

with acute, nonspecific stressors, but chronic exposure to a number of chemical contaminants also leads

to this effect. The decrease seems to be due to necrosis and perturbations in cell processing, both of which

could impact the overall condition of the individual

fish. An increased SSI, on the other hand, appears to

be linked to a diseased state. Disease incidence in the

population may thus be estimated with this parameter.

Information on alterations in all of the discussed

parameters may provide an early warning of an incipient or impending problem, one of the stated objectives of the BEST program (U.S. Fish and Wildlife

Service 1993). For these reasons, the National

Research Council (NRC), in its review of the BEST

program, strongly recommended that the program

include such general indicators (NRC 1995).

Although abnormal condition factors and

organo-somatic indices cannot be linked to specific

causal mechanisms due to the influence of confounding factors, they do indicate perturbations in biological systems at the organismal level. Changes in the

overall condition of organism(s) may corroborate

findings from other biomarkers and from chemical

analyses, and thereby document whole-organism and

higher-level effects for those contaminants that are

being measured as part of the weight-of-evidence

approach inherent in the BEST program. Correlation

among biomarkers at different biological and temporal scales should allow detection and quantification of

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 17
the biological impact of contaminants and may help

investigators in isolating causal mechanisms (Ham et

al. 1997).

THE NECROPSY-BASED FISH HEALTH

ASSESSMENT

Vicki S. Blazer

Two types of assessments have been directed toward

whole fish or gross (visible to the naked eye) observations: the incidence of gross external pathological

disorders and a more comprehensive necropsy-based

fish health assessment (internal and external).

Changes at this level represent an advanced stage, i.e.,

when a high incidence of skin abnormalities or hepatic tumors are found there may already have been a

significant impact on the population. Lesions

observed at this level also suggest that adaptive mechanisms – immunological, physiological, biochemical

– have been overwhelmed. However, it must be recognized that even lesions at this level can heal or be

resolved and are not necessarily life-threatening.

Background

The prevalence or percentage of fish with visible

pathological disorders has been used for many years

as a convenient and relatively easy indicator of environmental quality by fisheries managers and field personnel. Visible lesions generally include fin erosion

(Fig. 7), skin ulcers (Fig. 8), eye disorders (Fig. 9),

visible tumors and skeletal deformities. The index of

biotic integrity, which was designed to evaluate quality or condition of an aquatic ecosystem, includes

external abnormalities in its calculation. Three categories of fish community metrics, species abundance,

trophic composition, and health and abundance of

fishes, are used to reflect the condition of the fish

community and the environment in which it is found.

Health is determined by the proportion of fish with

disease or anomalies (Karr 1981).

Goede (1989) developed a systematic fish

health/condition or necropsy-based system for use by

fisheries personnel at the field level. It was developed to use minimal equipment to provide a rapid,

relatively inexpensive method in order to detect

trends in health and condition of fish populations

(Goede and Barton 1990). This method is a more

robust and sensitive indicator of fish health as it

includes both internal and external observations and

uses a computer program (AUSUM) to calculate a

variety of indices to compare populations either spatially or temporally. It has been used in hatchery

(Goede and Barton 1990; Novotny and Beeman 1990)

and field (Goede and Barton 1990) situations. This

original method did not provide a quantitative basis

for comparing populations statistically. Hence, a

health assessment index (HAI) and the AUSUM 430

program were developed by the Tennessee Valley

Authority (TVA) primarily for use in warm water

environments, and have been included in TVA’s standardized aquatic biomonitoring program. This modification of the original program gives numerical ratings to the observations. It has been used in a range

of reservoir types (Adams et al. 1993; Pritchard 1995)

as well as river systems (Coughlan et al. 1994; Sutton

et al. 2000).

Performing the Necropsy-Based Fish Health

Assessment

In Goede’s original method, field personnel are provided with a data sheet and accompanying description

of lesions (Table 4) in order to document observations

made on live fish. The information on the data sheet

is entered into the AUSUM computer program and a

summary of the necropsy is generated.

The quantitative health assessment index

(HAI) uses basically the same variables as Goede's

original method, with some additions, and substitutes

a numerical value as shown in Table 5. The values

are summed and an actual value is computed for each

fish. These values can then be used to calculate site

means and compare sites statistically. In the BEST

projects, a revised data sheet was used (Schmitt et al.

1999) and substituted values as described by Adams

et al. (1992b) were calculated. 

Factors That Can Affect the Necropsy-Based Fish

Health Assessment

Laboratory studies have demonstrated that gross

18 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Figure 7. Erosion (A) of the caudal fin.
METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 19

Table 4. Necropsy classification (from Goede 1989). Items in parentheses are entered onto the data sheet and

into the computer program. 

Length: Total length (L) in mm

Weight: Weight (W) in gm

CF: W X 105/L3

Eyes: Normal (N), Exopthalmia (E1, E2), Hemorrhagic (H1, H2), Blind (B1, B2), Missing (M1, 

M2), Other (OT)

Gills: Normal (N), Frayed (F), Clubbed (C), Marginate (M), Pale (P), Other (OT)

Pseudobranch: Normal (N), Swollen (S), Lithic (L), Swollen and Lithic (S&L), Inflamed (I), Other (OT)

Thymus: No hemorrhage (0), Mild hemorrhage (1), Severe hemorrhage (2)

Fins: No active erosion or previous erosion healed over (0), Mild active erosion with no bleeding 

(1), Severe active erosion with hemorrhage and/or secondary infection (2)

Opercles: No shortening (0), Mild shortening (1), Severe shortening (2)

Mesentary Fat: Internal fat is expressed with regard to amount present -

0 - None

1 - Little, less than 50% of each cecum covered

2 - 50% of each cecum is covered

3 - More than 50% of each cecum is covered

4 - Ceca are completely covered by large amount of fat

Spleen: Black (B), Red (R), Granular (G), Nodular (NO), Enlarged (E), Other (OT)

Hind gut: No inflammation (0), Mild inflammation (1), Severe inflammation (2)

Kidney: Normal (N), Swollen (S), Mottled (M), Granular (G), Urolithic (U), Other (OT)

Liver: Red (R), Light red (B), "Fatty" liver - "coffee with cream" color (C), Nodules in liver (D), 

Focal discoloration (E), General discoloration (F), Other (OT)

Bile: 0 - Yellow or straw color, bladder empty or partially full

1 - Yellow or straw color, bladder full and distended

2 - Light green to "grass" green

3 - Dark green to dark blue-green

Blood: Hematocrit - Packed cell volume of red blood cells (erythrocytes) expressed as percent of 

total blood volume. Leucocrit - Volume of white blood cells (leucocytes) expressed as per

cent of total blood volume ("buffy" zone of the packed cell column).
20 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Table 5. Fish health assessment variables and substituted values (as described in Adams et al. 1992b).

Variable Variable condition Field Substituted

Designation Value

Thymus No hemorrhage 0 0

Mild hemorrhage 1 10

Moderate hemorrhage 2 20

Severe hemorrhage 3 30

Fins No active erosion 0 0

Light active erosion 1 10

Moderate active erosion, some hemorrhage 2 20

Severe active erosion with hemorrhage 3 30

Spleen Normal: black, very dark red or red B 0

Normal: granular, rough appearance G 0

Nodular, containing fistulas or nodules D 30

Enlarged E 30

Other: aberration not fitting any above OT 30

Hindgut Normal, no inflammation or reddening 0 0

Slight inflammation or reddening 1 10

Moderate inflammation or reddening 2 20

Severe inflammation or reddening 3 30

Kidney Normal: firm, dark, flat N 0

Swollen: enlarged or swollen S 30

Mottled: gray discoloration M 30

Granular in appearance and texture G 30

Urolithiasis or nephrocalcinosis U 30

White-cream mineral deposits

Other: aberration not fitting any above OT 30

Skin Normal: no aberration 0 0

Mild skin aberrations 1 10

Moderate skin aberrations 2 20

Severe skin aberrations 3 30

Liver Normal: solid red or light red color A,B 0

"Fatty" liver, "coffee with cream" color C 30

Nodules or cysts in liver D 30

Focal discoloration E 30

General discoloration F 30

Other: deviation not fitting any above OT 30

Eyes No aberration, good, clear eyes N 0

Opaque eye (one or both) B 30

Swollen, protruding (one or both) E 30

Hemorrhaging (one or both) H 30

Missing one or both eyes M 30
lesions can be induced by exposure to contaminants

(Couch et al. 1977; Sindermann 1979; Capuzzo et al.

1988). Field studies have also shown that fish in

severely polluted areas have a higher frequency of

gross lesions than in similar, less polluted habitats

(Sindermann 1979; Malins et al. 1984; Malins et al.

1988; Couch 1985; O’Connor et al. 1987; Vethaak

and Rheinallt 1992; Fournie et al. 1996). The vast

majority of these studies have been conducted in estuarine or marine environments. The recent report by

Fournie et al. (1996) provides a regional-scale perspective on the prevalence of gross abnormalities.

The U.S. Environmental Protection Agency’s

Environmental Monitoring and Assessment Program

collected fish from 120 randomly selected estuarine

sites in the Virginian province (mid-Atlantic) and 220

sites in the Louisianian province (Gulf Coast) in 1991

and 1992. A total of 24,291 fish representing 143

species were examined. Skin lesions were the most

prevalent gross abnormalities in both provinces, followed by ocular abnormalities in the Virginian

province and branchial chamber abnormalities (gill

parasites, gill arch deformities; Fig. 10) in the

Louisianian province. Prevalence, in general, was

three-fold higher for demersal fishes than pelagic and

eight-fold higher at sites with high sediment contaminant concentrations.

Although contaminants alone can induce

grossly visible lesions in the laboratory, it must be

recognized that more often development of gross

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 21

Figure 8. Ulceration of the skin (A).

Other: deviation not fitting any above OT 30

Gills Normal: no apparent aberrations N 0

Frayed, ragged appearance F 30

Clubbed, swelling of tips C 30

Marginate: light discolored margin M 30

Pale, very light color P 30

Other OT 30

Pseudobranchs Normal, flat N 0

Swollen, convex in appearance S 30

Lithic, mineral deposits L 30

Swollen and lithic S&L 30

Inflamed, redness, hemorrhage I 30

Other OT 30

Parasites No observed parasites 0 0

Few observed parasites 1 10

Moderate parasite infestation 2 20

Numerous parasites 3 30

Table 5 (continued).
lesions requires a set of interacting conditions. These

include susceptible (or sensitive) hosts, stressors, and

infectious disease organisms. Hence, genetics,

species, food habits, contaminants, other water quality

parameters, weather, handling or capture methods,

trauma of other types, and the presence and number

of disease organisms play a role in the development

of gross lesions.

Value and Utility of the Necropsy-Based Fish Health

Assessment in the BEST Program

Necropsy-based assessments were developed to be

performed by field personnel with minimal training,

equipment, and expenses and to provide a rapid indication of the health of the population being studied.

They are not intended to be diagnostic but rather to

alert investigators to possible problems that should

then be investigated with more diagnostic (specific)

tests. It must be recognized that these methods are

not biomarkers but rather methods for documenting

lesions or changes that have advanced to the point of

being grossly visible. The value of the necropsybased HAI is: 1) it provides a systematic method for

field personnel to observe and document gross abnormalities; 2) it allows investigators to compare incidences of grossly observable lesions (those the public

is most concerned about) among sites; and, 3) it

allows investigators to document both spatial and

temporal trends. The drawback of this method, if

used alone, is it is not necessarily an early warning

system and does not generally provide an indication

of causes. Nevertheless, it does estimate cumulative

stress and, within the BEST program, it is used in

concert with other biomarkers to assess environmental

conditions.

HISTOPATHOLOGIC ASSESSMENT

Vicki S. Blazer

Histopathology is the study of lesions or abnormalities on a cellular level. Organs and tissues from fishes of any size, age or type can be examined. When

organs or tissues are properly fixed after use in the

necropsy-based health assessment, they can be stored

until later processing. Sectioning of these fixed tissues allows retention of in vivo relationships. More

than one tissue may be studied simultaneously to

determine biological effects of toxicants not only in

localized portions of certain organs but subsequent

derangements in tissues or cells at other locations.

This often allows for diagnoses of changes observed

grossly as well as indications of mechanisms of toxicity. In addition, cellular changes that occur prior to

development of grossly visible lesions can often be

detected. Macroscopic signs of toxicity are almost

always preceded by changes at the tissue, cellular or

molecular levels (Segner and Braunbeck 1990).

When the concentration of a toxicant(s) is not sufficient to cause acute death there may be sublethal or

adaptive changes which occur. When cell injury or

death of cells without the death of the organism

occurs, this is followed by cellular reactions and/or

host responses that can be described and sometimes

be diagnostic of cause(s).

Background

Histopathology has been used for many years to study

the cellular basis of infectious and noninfectious diseases. Fish respond to various insults in ways very

similar to mammals. Therefore, fish histopathology

utilizes knowledge gained over many years in human

and veterinary pathology. To perform histopathologic

22 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Figure 10. Parasites (A) on the gill.

Figure 9. A missing or scarred-over eye (A). 
METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 23

analyses, the pathologist must know the normal structure of fish tissues – their unique tissue types and

responses. Fortunately, there are a number of excellent descriptions available for selected species

(Ashley 1975; Grizzle and Rogers 1976; Groman

1982; Kubota et al. 1982; Yasutake and Wales 1983;

Ferguson 1989). Histopathologic changes observed in

infectious and noninfectious fish diseases have also

been documented in a number of sources (Ribeln and

Migaki 1975; Kubota et al. 1982; Ellis 1985;

Ferguson 1989; Roberts 1989; Sindermann 1990).

It has been recognized for many years that

fish respond to toxicant exposure, sometimes in fairly

specific ways (Braunbeck 1994), but often the same

response is elicited by a variety of chemicals or

chemical mixtures. There have been a number of

excellent reviews directed toward evaluation of

histopathologic changes as biomarkers of contaminant

exposure (Hendricks et al. 1985; Hinton et al. 1992;

Hinton 1993). Hinton et al. (1992) divided histologic

biomarkers into present and future. Present biomarkers are reliable, well-documented lesions evaluated

both in laboratory and field situations. Future or

potential biomarkers are those which may have strong

field relevance but have not been laboratory-validated

or which have been observed in laboratory exposures

but have not been evaluated in the field. It should be

recognized that as information is collected from both

field and laboratory studies future biomarkers may

become present biomarkers. For the organs routinely

examined in the BEST program only the liver and

reproductive tract have what are considered present

biomarkers by Hinton et al. (1992).

Performing Histopathological Analyses

Pieces of any organ (or whole small fish) are collected in the field and placed in fixative (Fig. 11). It is

very important to have sufficient fixative for the volume of tissue. If whole fish are fixed, it is very

important to slit open the abdominal cavity so fixative

can penetrate all organs. A variety of fixatives,

including neutral buffered formalin, NoTox,

Davidson’s (Dietrich’s) and Bouin’s (Luna 1992) have

been used. Each fixative type has advantages and disadvantages. Once properly fixed, tissues can be

stored for long periods of time at room temperature.

Routine processing of tissue involves trimming into

small pieces, dehydration through a series of alcohols

followed by an organic solvent, and infiltrating with

paraffin. Blocks of paraffin containing the tissues are

allowed to harden and then cut into 3 – 6 μm slices.

These sections are placed on glass slides, allowed to

dry, deparaffinized, and

stained. The most commonly used stain is

hematoxylin and eosin

(H&E). A variety of

other staining techniques

are used to demonstrate

infectious agents, cellular

components or specific

pathological responses

(Luna 1992).

Examination of

tissue sections should be

done by an experienced

histopathologist with

knowledge of normal fish

histology as well as an

understanding of pathology. Many histopathological studies rely on descriptive comparisons. Few, to date, have used quantitative methods that can then be compared statistically.

Reimschuessel et al. (1992) described a classification

system that allows for a statistical comparison among

groups. Lesions or observations are classified according to location, type of change (inflammatory, necrotic, growth), extent (focal, multifocal, diffuse) and

severity (rated on scale of 1-4 or 5). Hence, a hepatocarcinoma would be classified as DI.LI GC.NE.MA

M 3 (Digestive System.Liver Growth

Change.Neoplasia.Malignant Multifocal 3=moderate). If only one small focus of neoplastic tissue was

observed it may receive a 1 or 2 severity rating. Use

of this system allows comparisons of prevalence as

well as severity among sites.

Factors That Can Affect Histopathological Biomarkers

The liver is, by far, the most studied organ in terms of

toxic effects. The liver is the site of detoxification

mechanisms such as the mixed function oxidase system and metallothioein; it produces many regulatory

proteins; it is the site of energy storage – both lipid

and glycogen; and it synthesizes bile. Present hepatic

biomarkers include: hepatocellular coagulative necrosis, hyperplasia of regeneration, bile ductular/ductal

hyperplasia, hepatocytomegaly (hepatocellular hypertrophy), foci of cellular alteration, and hepatic and

bile duct neoplasms (Fig. 12). These have all been

well described by Hinton and Laurén (1990), Hinton

et al. (1992), and Hinton (1993), and contaminants

that cause these changes have been described by the

same authors. In laboratory exposures, species

(Braunbeck et al. 1990), sex (Braunbeck et al. 1989),

Figure 11. Tissue in fixative. 
24 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005 Normal liver, rainbow trout Basophilic foci (A); rainbow trout

Abnormal bile duct proliferation (A); rainbow trout Large tumor (A) compressing normal tissue; rainbow trout

Figure 12. Examples of histologic biomarkers in the liver (165x).
and age (Braunbeck et al. 1992) have all been shown

to affect hepatocellular reactions to chemicals.

Because the liver is also the site of energy storage —

i.e., glycogen and/or lipids — many other factors such

as nutritional status, season, and food habits will

affect the gross and histologic appearance of the liver.

The other present biomarker used in the

BEST program is oocyte atresia. Oocyte atresia is

degeneration and necrosis of developing ova. It is

normal for ovaries of fish to resorb ova that are not

released. However, it can become pathologic following exposure to certain contaminants. Oocyte atresia

is discussed in ‘Reproductive Indicators’.

The kidney has varied functions among fish

species. However, the kidney does filter large quantities of blood and produce urine, which is a major

route of excretion for some xenobiotics. In at least

some species of fish, as in mammals, the kidney can

regenerate after sublethal toxic damage

(Reimschuessel et al. 1990). Hence, development of

new nephrons has recently been suggested as a possible biomarker of exposure. Histologically these new

nephrons first appear as basophilic clusters. The

developing nephrons also stain more basophilic and

the number of these structures can be quantified per

square millimeter of tissue. These have been shown

to increase following laboratory exposures of goldfish

(Carassius auratus) to hexachlorobutadiene

(Reimschuessel et al. 1990), mercury (Reimschuessel

and Gonzalez 1992), and gentamicin (Reimschuessel

et al. 1991) and of rainbow trout (Oncorhynchus

mykiss) to tetrachloroethylene (Reimschuessel et al.

1994). Cormier et al. (1995) evaluated this biomarker

in field situations. They found significantly increased

numbers in Atlantic tomcod (Microgadus tomcod)

from a contaminated site (high levels of polychlorinated biphenyls and pesticides) when compared to

one reference site, but not a second reference site. In

brown bullheads (Ameriurus nebulosus) collected

from the Cuyahoga River (Ohio) with high levels of

polycyclic aromatic hydrocarbons, developing

nephrons were significantly higher than the reference

sites. These authors suggested that developing

nephrons may have potential as general indicators of

environmental condition but may have more utility

for detection of specific nephrotoxins. A significant

increase in developing nephrons was observed in

white suckers (Catostomus commersoni) collected in

a contaminated reach of the Sheboygan River

(Wisconsin) when compared to a reference area

(Schrank et al. 1997). Contaminants known to be

problems in this area include polychlorinated

biphenyls, polycyclic aromatic hydrocarbons and

heavy metals.

In the spleen, macrophage aggregates are a

histologic feature, and alterations in their number,

size, or both have been suggested as potential biomarkers. These structures are described in “Immune

System Indicators.”

Value and Utility of Histopathological Biomarkers in

the BEST Program

The use of histopathology of selected organs together

with the other fish health assessment methods allows

for a more comprehensive evaluation of abnormalities

than grossly visible alone. In many situations the

cause(s) of lesions observed grossly can be determined, and a judgment as to whether these are contaminant- or disease-related can be made. In addition,

changes not yet visible to the naked eye can be documented. Hence, histology offers a “very early warning” system for potential contaminant effects. In

addition, there have been few large-scale, regional or

river system type evaluations of histologic lesions.

Those that have been conducted have primarily

focused on estuarine or marine fishes (Sindermann

1990; May et al. 1987; Myers et al. 1994). Thus,

development of this type of database will provide

baseline comparisons for other studies, provide needed field information for validation of selected biomarkers, and provide additional information on

species, sex, and age as confounding factors.

IMMUNE SYSTEM INDICATORS

Vicki S. Blazer and Gail M. Dethloff

Immunotoxicology is a relatively new and emerging

branch of environmental toxicology. In the early

1970s it became apparent that chemicals known to be

present in the environment could compromise immunity in animals. In the 1980s it was confirmed that a

variety of environmental contaminants such as

toxaphene (Allen et al. 1983), lead, polychlorinated

biphenyls (PCBs) (Koller et al. 1983), and pentachlorophenol (Kerkvliet et al. 1982) produced

immunosuppression at dosages lower than those that

altered other commonly used toxicological indices

(Koller 1996). 

The immune system of fishes has been

shown to be as sensitive to a variety of environmental

contaminants (Weeks et al. 1992; Wester et al. 1994;

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 25
Zelikoff 1994) as that of homeotherms. For this reason there is strong interest in incorporating measures

of immunity and disease resistance for fish into the

overall organism health assessments conducted as part

of the BEST program. It is important to note, however, that this represents an emerging area of aquatic

toxicology, and both the methods and the significance

that can be attached to their results are developing

rapidly. Generally, assessment of immune system

function requires harvesting live cells (macrophages

or lymphocytes), maintaining them at optimal conditions in sterile liquid media, and performing functional assays within a short period of time. Most of the

functional assays require sterile techniques, sophisticated equipment, and specially trained personnel, and

often utilize radio-labeled substrates. Although these

types of assays have been used in field studies of limited scope, they are simply not feasible in large-scale

programs such as the BEST projects. Consequently,

Weeks et al. (1992) suggested a tiered approach for

use in screening or comprehensive analysis of

immunomodulatory effects of chemicals on aquatic

organisms. Organosomatic indices, histology of

spleen and lymphoid tissue, lysozyme activity, and

macrophage aggregate analyses were all included in

tier 1. These are methods that can be performed on

cryogenically frozen plasma or serum (lysozyme) or

with preserved tissue (histology of lymphoid tissue;

macrophage aggregate parameters) and were chosen

as biomarkers for the BEST program.

SERUM/PLASMA LYSOZYME ACTIVITY

Lysozyme is an enzyme with antibacterial and antiviral activity (Jolles and Jolles 1984) that acts on peptidoglycan (a protein in bacterial cell walls), causing

lysis of the bacteria (Chipman and Sharon 1969).

Gram-positive bacteria are most susceptible to its

action; however, rainbow trout (Oncorhynchus

mykiss) lysozyme has been shown to lyse a number of

gram-negative bacteria (Grinde 1989). It is believed

that the lysozymes of both fishes and mammals work

in conjunction with other proteolytic enzymes to lyse

gram-negative organisms (Neeman et al. 1974;

Hjelmeland et al. 1983). Lysozyme is an enzyme of

leucocytic origin considered to be an important component of the nonspecific humoral disease resistance

mechanisms of fishes (Yano 1996). It has also been

found to increase with enhanced phagocytic function

of macrophages in rats and therefore has been suggested as a marker for macrophage function

(Kokoshis and Di Luzio 1979). However, this has not

been validated with fishes.

Background

Lysozyme is a disease resistance factor that has been

studied for many years, particularly in relation to

infectious diseases and vaccination of cultured fishes.

Serum/plasma lysozyme levels have been shown to

rise in the early stages of coccidial infections of common carp (Cyprinus carpio), peaking at the point clinical lesions are evident and then declining slightly as

the infection progresses (Studnicka et al. 1986).

Common carp immunized with the bacteria

Aeromonas punctata (Vladimirov 1972) or experimentally infected with a mixture of A. punctata and

Pseudomonas alcaligenes (Siwicki and Studnicka

1987) exhibited a rise in circulating lysozyme. This

has also been shown for Atlantic salmon (Salmo

salar) inoculated with the bacterium Vibrio anguillarum (Muona and Virtanen 1993) or infected with

Aeromonas salmonicida (Møyner et al. 1993).

Recently, there has been some interest in using

lysozyme activity as an indicator of environmental

stress with the potential of linking alterations in

lysozyme to disease prevalence. However, the use of

lysozyme in either field studies or contaminant-related laboratory studies has been limited.

Performing the Lysozyme Assay 

Blood is collected from living fish and chilled, then

allowed to clot or centrifuged to obtain either serum

or plasma respectively, depending on study objectives

and protocol. Plasma or serum is quick-frozen in liquid nitrogen or dry ice and stored cryogenically 

(-80°C) until needed for analysis. Measurement of

lysozyme activity is based on the lysis of suspensions

of the bacteria Micrococcus lysodeikticus by

serum/plasma. Variations of a photometric method

(Litwack 1955) and of a lysoplate assay described by

Osserman and Lawlor (1966) have both been used

with fish serum. The pH optima for the conduct of

the assay vary with taxon and should be determined

in advance for the species under consideration (Möck

and Peters 1990; Blazer et al. 1996). Lysozyme from

the whites of chicken (Gallus gallus) eggs is generally used to produce a standard curve. However, the

optimum of hen egg white lysozyme is pH 7.5 (Möck

and Peters 1990) while fish serum optima generally

range from 5.5 (Möck and Peters 1990; Blazer et al.

1996) to 8.0 (Kusuda et al. 1987). Investigators have

recognized this difference in pH optima among fishes;

however, there is no consistency in approaches to deal

26 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
with the difference. Many investigators use the same

buffer system for both standard and unknowns (Ellis

1990; Tahir et al. 1993), while others have suggested

the use of an internal standard (Möck and Peters

1990; Røed et al. 1993). We chose to use a kinetic

method, in which groups of fish are compared using

mOD/min or the change in absorbance over a specified time.

The microplate method developed by Tahir

et al. (1993) and modified by Blazer et al. (1996) was

used to analyze the plasma of fish from the BEST

projects (Schmitt et al. 1995; Bartish et al. 1997). A

0.075% suspension of dried M. lysodeikticus is prepared in the appropriate buffer. Serum or plasma (25

μl) is added (triplicate determinations for each fish) to

wells of a 96-well, flat-bottomed micro-titer plate. A

175-μl aliquot of the above suspension is then added,

and the plate is immediately shaken and read on a

kinetic microplate reader at 450 nm every 15 seconds

over a 5 minute period.

Factors That Can Affect Serum/Plasma Lysozyme

Activity

As noted above, disease can alter lysozyme levels.

Water temperature appears to have a positive correlation with lysozyme (Balfry et al. 1997; Aranishi et al.

1998). Lysozyme activity has also been found to be

modulated by certain dietary factors (Kiron et al.

1995; Roberts et al. 1995). In addition, seasonal, sexual, species, strain and age-dependent variations in

plasma lysozyme activity have been reported (Balfry

et al. 1997; Holloway et al. 1993; Fletcher and White

1976; Fletcher et al. 1977; Studnicka et al. 1986; Lie

et al. 1989; Røed et al. 1993). Fish species that

undergo smoltification experience altered lysozyme

activity during this developmental stage (Muona and

Soivio 1992).

General stressors such as handling and transport can affect lysozyme activity (Möck and Peters

1990; Fevolden et al. 1994). However, the effects

appear to depend on the type or duration of the stress.

Möck and Peters (1990) reported that 30 min of handling caused either a decrease or increase in rainbow

trout lysozyme activity. Demers and Bayne (1997)

reported an increase in the lysozyme levels of rainbow trout with short-term air exposure. Transport

stress lasting 2 hours or acute exposure to un-ionized

ammonia levels of 0.450mg/l, however, caused a significant decrease in lysozyme activity in rainbow

trout (Möck and Peters 1990). Differing responses to

short- and long-term stressors were also seen in common carp (Hajji et al. 1990) and Atlantic salmon

(Røed et al. 1993). In dab (Limanda limanda), transport stress (approximately 60 minutes) resulted in significantly decreased serum lysozyme activity

(Hutchinson and Manning 1996b).

Investigations have recently begun into the

effects of contaminants on lysozyme activity. In laboratory studies, lysozyme activity of dab was reported

to decrease after exposure to oil-contaminated sediments from drilling sites (Tahir et al. 1993). Also,

common carp exposed to the organophosphate insecticide trichlorphon (Siwicki et al. 1990) had decreased

lysozyme activity. Field studies have found reduced

lysozyme activity in common carp exposed to sewage

effluent (Price et al. 1997) and in European plaice

(Pleuronectes platessa) collected along a sewage gradient (Secombes et al. 1995). However, dab exposed

to sewage sludge in the laboratory for 12 weeks did

not display significantly altered lysozyme activity

(Secombes et al. 1991).

Value and Utility of Serum/Plasma Lysozyme Activity

in the BEST Program

Despite its limited use in environmental toxicology,

serum/plasma lysozyme activity remains attractive for

use in the BEST program because it is among the few

logistically reasonable markers of immune system

function. For this reason, lysozyme activity has been

included among the indicators of fish health as a general indicator of exposure to a wide variety of factors

(including contaminants) not otherwise accounted for

by more specific measures. However, it must be recognized that a great deal more field and laboratory

data need to be generated to validate this biomarker.

MACROPHAGE AGGREGATES

Pigment-bearing macrophages are a prominent feature

in fish spleen, kidney and sometimes liver (Agius

1980). In advanced teleosts they form discrete aggregations called macrophage aggregates (MA) or

melanomacrophage centers (Fig. 13). Macrophage

aggregates are believed to be functional equivalents

of the germinal centers, active in the centralization of

foreign material and cellular debris for destruction,

detoxification or reuse, the storage of exogenous and

endogenous waste products, the immune response,

and iron storage and recycling (Ferguson 1976; Ellis

et al. 1976). The subject of MAs has recently been

thoroughly reviewed by Wolke (1992) and by Blazer

et al. (1997).

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 27
Background

Macrophage aggregates have been recognized as normal structures of fish spleen, kidney and liver since

the 1920s (Jolly 1923; Yoffey 1929). Roberts (1975)

coined the term melanin-macrophage or melanomacrophage centers and suggested these structures

should be considered part of the reticulo-endothelial

system and hence part of the defensive system against

infectious disease. It has been shown that particulate

matter (such as carbon particles or bacteria) injected

intraperitoneally or intravascularly eventually accumulates within these aggregates or centers (Mackmull

and Michels 1932; Ellis et al. 1976). It is also known

that bacteria and parasites may be phagocytized and

taken to these centers (Roberts 1975; Ferguson 1976;

Herraez and Zapata 1987). Agius was the first to

study these structures in depth (reviewed in Agius

1985). He examined their phylogenetic development

(Agius 1980), their ontogeny and age-related changes

(Agius 1981), the effects of starvation on these structures (Agius and Roberts 1981), their role in iron storage (Agius 1979), and the development of pigments

within these centers (Agius and Agbede 1984).

Wolke et al. (1981, 1985a) first suggested

pigmented macrophage accumulations as potential

monitors of fish health. Numerous studies had documented an increase in their number, size or hemosiderin content in fish collected at contaminated sites

when compared to those collected at reference sites.

For this reason, MAs have been suggested as potentially sensitive biomarkers of contaminant exposure.

Although they are structures observed histologically,

it has been suggested that MAs may be immunotoxicologic biomarkers (Weeks et al. 1992; Blazer et al.

1997).

Measuring Macrophage Aggregate Parameters 

Macrophage aggregates are measured in preserved tissues. They can be found in spleen, kidney and liver

depending on the fish species. Pieces of spleen, liver

or kidney are routinely processed for histology,

embedded in paraffin, cut into 5-μm sections with a

microtome, mounted on glass slides, and stained with

hematoxylin and eosin (H & E). A special staining

procedure called the Perl’s method (Luna 1992) is

used to increase the ease in visualizing these structures and also allows observation of all the pigments

within the MA. With this stain, melanin, the

melanosome pigment derived from tyrosine metabolism, is black; hemosiderin, a protein-bound iron pigment, is blue; and ceroid/lipofuscin, lipogenic pigments arising from the oxidation of unsaturated lipids,

are yellow-tan (Fig. 14). Macrophage aggregate

parameters are measured using a computer-based

image analysis system. The number of aggregates in

2 sq. mm of tissue and their mean size are measured.

From this information the % of tissue occupied by

MAs can be calculated. Currently, qualitative observations are made on pigment content. However, the

relative measurement of pigments is possible with

true color image analysis. A number of studies have

compared the utility of splenic and hepatic MAs as

indicators of contaminant exposure, and splenic MAs

28 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Figure 13. Splenic tissue (165x). Top panel: High density of

macrophage aggregates. Bottom panel: Low density of

macrophage aggregates. 
have been found to be the most responsive (Blazer et

al. 1994b). Therefore, in the BEST projects splenic

MA parameters were measured.

Factors That Can Affect Macrophage Aggregates

Occurrence of MAs may vary depending on the size,

nutritional status, or health of a particular fish species

(Agius 1979; Agius 1980; Agius and Roberts 1981;

Wolke et al. 1985b). Larger fish, fish with nutritional

deficiencies, or fish in poor health tend to have more

or larger MAs. In addition, the number and/or size of

MAs increase with age (Brown and George 1985;

Blazer et al. 1987).

Relatively few studies on the response of

fish MAs to contaminants under controlled laboratory

conditions have been reported. The majority have

found that MAs increase with exposure to contaminants. Exposure of goldfish (Carassius auratus) to

phenylhydrazine caused extensive hemolysis and

resulted in significant increases in MA number and

size (Herraez and Zapata 1986). Laboratory exposure

of Atlantic cod (Gadus morhua) to the water soluble

fractions of two types of crude oil caused an increase

in MAs (Khan and Kiceniuk 1984). An increased

number of MAs was also seen in dab exposed to

sewage sludge (Secombes et al. 1991) and common

carp exposed to sediments contaminated with polychlorinated dibenzodioxins/dibenzofurans (PCDD/Fs)

and PCBs (van der Weiden et al. 1993). European

plaice exposed to water-borne potassium dichromate

had an increased MA number but a decrease in the

mean size of the aggregates. In addition, there was an

increase in the density of the melanin pigment (Kranz

and Gerken 1987). In striped bass (Morone saxatilis)

exposed to dietary arsenic, splenic MA number

increased with increasing concentrations of arsenic.

In addition, there was a striking increase in hemosiderin in fish fed higher levels (200 ppm) of arsenic

(Blazer et al. 1997). In contrast, juvenile yellowfin

sole (Pleuronectes asper), rock sole (Pleuronectes

bilineatus) and Pacific halibut (Hippoglossus

stenolepis) experienced decreased hepatic MA numbers with exposure to hydrocarbon-contaminated sediments; the decrease in MA numbers may, however,

have been linked to significantly decreased growth in

the exposed fishes (i.e. metabolism) (Moles and

Norcross 1998).

A larger number of studies have noted

changes in MA number in liver, spleen or kidney of

fish from polluted waters (Poels et al. 1980; Bucke et

al. 1984; Khan and Kiceniuk 1984; Kranz and Peters

1984; Wolke et al. 1985b; Spazier et al. 1992). Very

consistently, MA numbers have been reported to

increase at contaminated sites versus reference sites.

There are a few studies that have suggested a

decrease or no significant effect. Payne and Fancey

(1989) reported a decrease in hepatic MAs of winter

flounder (Pleuronectes americanus) after 4 months of

laboratory exposure to hydrocarbon-contaminated

sediments, when compared to nonexposed. However,

there was no information on fish size/age, condition

prior to exposure, growth or weight loss during the

experiment. Haaparanta et al. (1996) compared perch

(Perca fluviatilis) and roach (Rutilus rutilus) from

four lakes with a gradation of water quality. Three of

the lakes were polluted by effluent from a paper and

pulp mill. These authors reported no observed patterns of splenic, hepatic or kidney MAs in terms of

rankings of water quality. However, again methodology may play a role, as age of fish was not considered

and size distribution of captured fish was different

among the lakes. Conversely, Couillard and Hodson

(1996) found increased density of MAs in white suckers (Catostomus commersoni) to be a useful marker of

bleached-kraft mill effluent. In winter flounder from

eight coastal embayments, the splenic area occupied

by MAs was correlated with chemical contamination

of sediments: PCBs, polycyclic aromatic hydrocarbons, and trace metals measured at these sites all correlated with MA parameters and the number of MAs

and their size increased with increasing levels of

benzo(a)pyrene (Benyi et al. 1989; Gardner et al.

1989). A study comparing a number of biochemical

and physiological indicators of contaminant stress to

population level responses of redbreast sunfish

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 29

Ceroid/lipofuscin

Hemosiderin

Melanin

Figure 14. A macrophage aggregate in a tissue section stained by

the Perls method demonstrating all three pigments that may be

present (1680x).
(Lepomis auritus) sampled four sites along a PCB and

mercury gradient in a stream receiving point-source

discharges. At the sites sampled along this gradient,

MAs were more sensitive at indicating contaminant

effects than cytochrome P450 and NADPHcytochrome c reductase levels, liver-somatic index,

fecundity, and population abundance and were as sensitive as serum triglyceride levels and the liver

RNA:DNA ratio (Adams et al. 1992b). It is interesting to note that in a subsequent study at these same

sites, macrophage function was also significantly

depressed at the first three sites (Rice et al. 1996).

Blazer et al. (1994a) also found an increase in MA

number and percent of tissue occupied by MAs in

species from contaminated sites when compared to

less contaminated sites. MAs of gafftopsail catfish

(Bagre marinus) were the best indicators of elevated

tissue contaminants whereas MAs of both gafftopsail

catfish and spot (Leiostomus xanthurus) were good

indicators of elevated sediment contaminants.

Value and Utility of Macrophage Aggregates in the

BEST Program

Macrophage aggregates have long been recognized as

potentially useful biomarkers (Wolke et al. 1985a) but

intrinsic and extrinsic factors may confound investigations into the role of contaminants. Hence, it has

been recognized that continuing field and laboratory

studies are needed to fully understand the relationship

of MA number, size, and pigment content to various

environmental contaminants (Hinton et al. 1992;

Wolke 1992; Blazer et al. 1997). Nevertheless, this

histologic as well as potential immune system biomarker, although quite general, has been shown

through both field and laboratory studies to respond

to exposure of fish to a variety of contaminants of

concern. In addition, the assay is logistically reasonable as part of the suite of health assessment indicators because it is based on preserved tissues collected

as part of an overall procedure. 

REPRODUCTIVE INDICATORS

Kelly K. McDonald, Timothy S. Gross, Nancy

D. Denslow, Vicki S. Blazer

To predict the potential of a fish or wildlife population to persist within a given ecosystem, it is necessary to evaluate the reproductive health and capacity

of the individuals within that population. Multiple

definitions are available for “reproductive success,”

including the number of eggs in a clutch, number of

viable offspring, or number of viable offspring capable of reproducing. Whereas reproductive success

can be a reliable indicator of a population’s reproductive health and potential, it is often difficult to evaluate, especially in aquatic environments where fish and

other wildlife are not easily contained for convenient

monitoring. Consequently, the development of techniques for measuring other reproductive indicators

(e.g., sex steroid hormones, vitellogenin, gonadosomatic indices, and gonadal histopathology) has aided

researchers in assessing the reproductive health of

many fish species. These reproductive indicators, or

biomarkers, provide quantifiable measures of biochemical, physiological, or histological changes that

occur naturally throughout the reproductive cycle.

These indicators may also be altered by exogenous

factors and may reflect reproductive impairment.

Reproductive biomarkers are useful for (1) examining

the effects of environmental stressors, such as chemical contaminants, eutrophication, and temperature

fluctuations, on an individual or population, (2) predicting future reproductive trends and population

abundance by serving as early indicators of sublethal

effects of environmental stressors, and (3) providing

insight into the causal relationships between reproductive failure and environmental stressors. For an

accurate understanding of the reproductive health of

an individual or population, it is necessary to measure

a variety of responses using multiple methods.

The complex process of vertebrate reproduction includes sexual differentiation, embryonic development and birth or hatching, sexual maturation,

gametogenesis, mating, and fertilization. Fish display

a wide repertoire of reproductive strategies; however,

strict hormonal and environmental regulation is

required for the reproductive processes of all species.

Hormonal regulation is the function of the endocrine

system, a collection of cells, tissues, and organs that

produce and secrete hormones that influence virtually

every stage of the lifecycle (reviewed by Norris 1997;

van der Kraak et al. 1998). The coordinated reproductive efforts of the hypothalamus, pituitary, and

gonads have led to the designation of the hypothalamo-hypophysial (pituitary)-gonadal axis. The hypothalamus exerts ultimate control over the reproductive

process by synthesizing and releasing gonadotropinreleasing hormones (GnRHs) that stimulate the

release of several gonadotropin hormones from the

pituitary (Fig. 15). In nonmammalian vertebrates,

including several orders of teleosts, two gonadotropin

hormones, GTH-I and GTH-II, have been characterized (Kawauchi et al. 1989; Swanson et al. 1991) and

30 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
their functions compared to the mammalian folliclestimulating hormone and leutinizing hormone, respectively (Redding and Patino 1993). In fish, GTH-I is

important for vitellogenesis, the process of yolk protein synthesis, and early gonadal development. GTHII, on the other hand, is secreted late in gonadal

development and plays a role in the final maturation

and release of mature gametes (ovulation in females,

spermiation in males). As mentioned earlier, a variety

of environmental factors, including photoperiod, temperature, salinity, dissolved oxygen concentration,

water flow, and turbidity, influence each stage of

reproduction (reviewed by Donaldson 1990; Munro et

al. 1990). Some of these environmental factors will

be addressed for each parameter.

SEX STEROIDS

The sex steroids, a class of hormones derived from

cholesterol and synthesized by the gonads in response

to circulating levels of GTH-I and GTH-II (Fig. 15),

collectively control the development of the gonads

and gametes, secondary sexual characteristics, and

reproductive behavior such as pheromonal attraction,

spawning, and parenting (Liley and Stacey 1983;

Fostier et al. 1983). The reproductive hormones can

be divided into two major categories: (1) the androgens, produced by the testes and ovaries; and (2)

estrogens and progesterone, which are primarily produced by the ovaries but are often synthesized in

smaller amounts by the testes. Although species and

maturational stage largely influence the type of reproductive hormones synthesized by the gonads, the

major androgens in fish include testosterone, 11-

ketotestosterone, and androstenedione. The predominant estrogens are 17β-estradiol and estrone. Sex

steroids in immature fish probably influence gonadal

differentiation, whereas these same hormones play an

important role in gametogenesis, ovulation, and spermiation in mature fish (Redding and Patino 1993;

Barry et al. 1990; Patino and Thomas 1990a; Patino

and Thomas 1990b).

Background

Measuring reproductive hormones in plasma may

yield biochemical information concerning the reproductive status of an individual, as well as provide a

method for detecting potential reproductive injury.

Although this technique has been used to gain information regarding early development and the reproductive cycles of healthy individuals (Johnson and

Casillas 1991; Freund et al. 1995), monitoring sex

steroid concentrations has primarily been utilized to

study the endocrine-disrupting effects of various environmental chemicals. Early evidence from studies of

fish-eating birds suggested that some chlorinated

hydrocarbons impaired normal cyclical changes in

plasma concentrations of estrogen, progesterone, and

testosterone (Peakall et al. 1981). Over the last few

decades, sex steroid hormones have evolved as convenient biomarkers for detecting contaminant-induced

biochemical alterations, and fish have emerged as

favorable models for examining the effects of environmental pollutants in aquatic ecosystems.

Performing Sex Steroid Hormone Assays

Due to its facility and inexpensiveness, the method

utilized by the BEST program involves the analysis of

sex steroid hormones (e.g., 17β-estradiol, testosterone, 11-ketotestosterone) in serum or plasma samples by radioimmunoassay (Goodbred et al. 1997).

Using this technique, a given hormone concentration

in a serum or plasma sample can be determined. The

samples are extracted twice with an organic solvent

(e.g., ethyl ether) to isolate lipophilic compounds such

as sex steroids. The reaction solution, comprised of

the serum or plasma extract, a radiolabeled hormone

(e.g. 3H-estradiol), and a corresponding hormone-specific antibody, is allowed to equilibrate overnight.

During this time the unlabeled hormone from the

extract and a constant concentration of the corresponding radiolabeled hormone compete for the same antibody binding sites. The non-antibody bound, radiolabeled hormone is then removed from solution by

addition of charcoal dextran followed by centrifugation. The resulting solution, containing bound radiolabeled hormone, is measured using scintillation spectrophotometry. Known concentrations of unlabeled

hormone, prepared in media or buffer, are subjected

to the same experimental procedure in order to generate a standard curve. Quantification of the steroid

level in the sample is ascertained by aligning values

of the inhibition curve, generated using stable concentrations of radiolabeled hormone, with those of the

standard curve. Each sample is measured in duplicate

for the selected hormones and corrected for an extraction efficiency. Cross-reactivities of the hormones

being tested with other steroid hormones are measured and reported.

Factors That Can Affect Sex Steroid Hormones

Circulating sex steroid levels are subject to normal

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 31
variation due to differences in sex, age, geographical

location, species, and season (Goodbred et al. 1997;

Denslow et al. 1998; Chang and Chen 1990; Barry et

al. 1990; Down et al. 1990; Bromage et al. 1982; So

et al. 1989). Several studies document the greatest

variance in sex steroid concentrations during spawning, yet hormone levels from individuals at the same

site have been shown to vary up to 30-fold during

gonadal recrudescence (Chang and Chen 1990; Down

et al. 1990; Folmar et al. 1996). There is also evidence that the stress of collecting, holding, and

obtaining blood from fish may affect hormone concentrations (Jardine et al. 1996; McMaster et al. 1994;

van den Heuvel et al. 1995; Barton and Iwama 1991;

Magri et al. 1982), so conservative sampling protocols should be used.

In terms of xenobiotics, numerous studies

demonstrate that exposure to a variety of structurally

unrelated contaminants, including bleached kraft mill

effluent (BKME) (McMaster et al. 1991; Munkittrick

et al. 1991; Munkittrick et al. 1992; Munkittrick et al.

1994), agricultural pesticides (Goodbred et al. 1997;

Gross et al. 1997; Singh et al. 1994; Singh and Singh

1987; Singh and Singh 1991), industrial chemicals

(Sivarajah et al. 1978; Spies et al. 1996), and heavy

metals (Allen-Gil et al. 1993; Thomas 1988), can lead

to alterations in plasma sex steroid concentrations in a

variety of fish species. For example, Grady et al.

(1998) found that environmentally relevant concentrations of atrazine affected hormone concentrations in

both male and female largemouth bass (Micropterus

salmoides) (Fig. 16). There is substantial evidence to

suggest that pollutants can alter sex steroid levels by

interfering at multiple sites along the hypothalamopituitary-gonadal axis. Mukherjee et al. (1991)

reported that phenol and sulfide, two components of

BKME, disrupt steroidogenesis by inhibiting the conversion of cholesterol to sex steroids in the gonads.

Heavy metals, on the other hand, may reduce sex

steroid levels by stimulating the production of cortisol, which subsequently accelerates the metabolism of

steroids in the liver (Hansson 1981). It is also possible that multiple endocrine pathways are affected

simultaneously (Singh et al. 1994), especially if

wildlife populations are exposed to combinations of

chemicals. Consistent with this theory, van der Kraak

et al. (1992) reported that in addition to increasing the

metabolism of sex steroids in the liver, BKME interfered with the release of gonadotropins from the pituitary and suppressed steroid synthesis in the ovaries.

Value and Utility of Sex Steroid Hormones in the

BEST Program

Current risk assessment strategies are generally

designed to detect and characterize acutely hazardous

chemicals that pose potential health risks such as cancer and birth defects. However, the effects of environmental stressors on the endocrine and reproductive

systems are often subtle, making them difficult to

32 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Figure 15. The reproductive hormone pathway in teleosts. GnRH =

gonadotropin-releasing hormone, GTH-I and -II = gonadtropin hormones. 
METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 33

identify using traditional protocols. Measuring reproductive hormone levels provides one method for monitoring subtle physiological alterations that may lead

to long-term, detrimental population effects. Reviews

by several investigators correlate reduced sex steroid

levels, in response to contaminant exposure, with

reproductive impairment (Colborn and Clement 1992;

Mayer et al. 1992), although the issue of determining

cause-and-effect from sex steroid hormone studies

remains somewhat controversial.

Although several caveats exist, measuring

sex steroids is rapid and relatively inexpensive and

can provide valuable biochemical and physiological

information. With regard to natural variance, it is

important to determine the base-line values and consider the environmental factors that may influence the

reproductive physiology of the species being investigated. In general, sex steroids as biomarkers for

reproductive status and health serve as convenient

early indicators of stress and are most reliable when

Figure 16. Mean plasma steroid concentrations for male (blue) and female (red) largemouth bass following a 20-d exposure to atrazine. 
field experiments are validated with laboratory experiments and species-specific physiology and environmental factors are considered. Other methods which

examine reproductive capacity at different organizational levels (i.e., community, species, tissue, cellular,

and subcellular) should also be employed to aid in

interpretation of field data particularly since, with the

exception of a few well-documented studies

(Munkittrick et al. 1991; Sepulveda et al. 1998),

information regarding ecological implications of hormonal modulations in fish exposed to harmful chemicals is limited.

VITELLOGENIN

Vitellogenesis usually occurs during the prespawning

season in oviparous fish (Scott and Sumpter 1983;

Lamba et al. 1983). GTH-I, released from the pituitary, stimulates the production of 17β-estradiol in the

ovaries (reviewed by Redding and Patino 1993).

High levels of circulating 17β-estradiol stimulate the

liver to synthesize and release vitellogenin (vtg), a

glycolipophosphoprotein egg yolk precursor (Fig. 17).

Evidence for this estrogen-dependence comes from

several studies in which high concentrations of estrogens have been reported prior to, or concurrent with,

the onset of vitellogenesis in a number of fish including European flounder (Platichthys flesus) (Emmersen

and Petersen 1976), stinging catfish (Heteropneustes

fossilis) (Sundararaj et al. 1982), and chum salmon

(Oncorhynchus keta) (Ueda et al. 1984). In addition,

administering estradiol to males and immature (nonvitellogenic) females has been reported to induce vtg

production in a number of studies (Emmersen and

Petersen 1976; Campbell and Idler 1980; Korsgaard

et al. 1983). After being released by the liver into the

bloodstream, vtg is delivered to the ovaries where it is

recognized by high-affinity receptors on the surface

of the oocyte and subsequently internalized by receptor-mediated endocytosis (Fig. 17) (Specker and

Sullivan 1994). In the final step, vtg is enzymatically

cleaved within the oocyte to form the yolk proteins

that serve to nourish the developing embryo.

Depending on the species of fish and the experimental

technique used, GTH-I, GTH-II, or both influence the

uptake of vtg by the developing oocyte. In rainbow

trout (Oncorhynchus mykiss), in vivo and in vitro

methods have documented the involvement of GTH-I,

but not GTH-II (Tyler 1991), whereas experiments

using vitellogenic goldfish (Carassius auratus)

reported enhanced vtg uptake following exposure to

both gonadotropins (Rodriguez and Peter, personal

communication).

Background

Vitellogenin is an estrogen-inducible protein that is

normally synthesized by the liver of nonmammalian

female vertebrates during oocyte development. The

discovery that many structurally diverse chemicals

[e.g., chlorobenzene and dichlorodiphenyltrichloroethane (DDT)] that are released into the environment possess estrogenic properties (Arcand-Hoy

and Benson 1998; McLachlan 1993) has encouraged

the development and utilization of bioassays that

evaluate estrogenicity. However, the development of

a universal assay has been challenging because some

chemicals that bind the estrogen receptor do not elicit

an estrogenic response and, conversely, some compounds evoke an estrogenic response without interacting with the receptor. This unpredictable structurefunction relationship demands an assay based on the

bioactivity of potential environmental estrogens. The

fact that vtg synthesis is primarily regulated by circulating estrogens has made vtg an attractive indicator

of potential estrogen action (Palmer and Palmer 1995;

Palmer and Selcer 1996).

Performing the Vitellogenin Assay

Vitellogenin concentrations in plasma or serum can be

determined using several different immunological

methods. Sodium dodecylsulfate polyacrylamide gel

electrophoresis and immunoblotting (Western analysis) can be used to detect vtg protein with an anti-vtg

antibody, and relative protein concentrations can be

quantified using densitometry. Alternatively, the

enzyme-linked immunosorbent assay (ELISA) is less

expensive and has become the screening method of

choice for the BEST program (Folmar et al. 1996).

Several vtg antibodies have been developed and well34 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Figure 17. The role of the ovaries and liver in vitellogenesis. 
characterized (Heppell et al. 1995; Denslow et al.

1996), including the anti-carp vtg monoclonal antibody, Mab HL 1147 (2D3-3a9) (Folmar et al. 1996).

The following summarizes the ELISA protocol adapted from Folmar et al. (1996). Microtiter

plates (96-well) are coated with purified anti-carp vtg

monoclonal antibody and incubated overnight. The

next day, plates are washed with Tris Buffered

Saline/Tween 20 (TBST) solution and incubated with

bovine serum albumin to block non-specific antibody

binding. After thoroughly washing with TBST, plasma samples (diluted from 1:500 to 1:5000) are added

in duplicate to the plates and incubated overnight.

Standard curves are constructed by adding serial dilutions of purified vtg to male control plasma and

processed according to the same method. The following day, the plates are washed with TBST and incubated with a rabbit anti-vtg polyclonal antibody for 2

h. The rabbit antibody binds to the vtg captured by

the monoclonal antibody in the first step. The polyclonal antibody is in turn bound by a goat anti-rabbit

IgG conjugated to alkaline phosphatase, which is

applied to the wells and allowed to incubate for 2 h.

After a final series of washes with TBST, p-nitrophenyl phosphate in carbonate buffer is added to each

well and incubated for 30 min. The p-nitrophenyl

phosphate serves as a substrate for the alkaline phosphatase and this reaction generates a yellow color that

can be quantified by measuring the absorbance at 405

nm using an automated ELISA reader (Fig. 18).

Vitellogenin concentrations are determined by subtracting values obtained from male control serum and

comparing to standard curves.

Factors That Can Affect Vitellogenin

It is clear from the literature that the vitellogenic

response is dependent on a number of intrinsic factors, including species, sex, and maturation/reproductive stage, as well as extrinsic factors, such as water

temperature, season, and chemical composition of the

aquatic environment (Wallaert and Babin 1994;

Korsgaard et al. 1986). There is evidence for the regulation of vtg synthesis by circadian rhythms in catfish (Lamba et al. 1983), photoperiod in rainbow trout

(Bromage et al. 1982), and winter sea temperatures in

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 35

Figure 18. The standard curve in a vitellogenin assay (run in triplicate in columns 1-3, rows A-H) illustrating the gradation of color change.

Columns 4-6 contain unknown plasma samples run in triplicate.
small-spotted catshark (Scyliorhinus canicula) (Craik

1978). In two independent studies, female Atlantic

salmon (Salmo salar) and rainbow trout were monitored for vtg levels following exposure to cyanide.

Plasma levels of vtg were reduced in rainbow trout

(Ruby et al. 1986), whereas an increase in the plasma

vtg concentration was reported in Atlantic salmon (So

et al. 1987). The opposite response to the same environmental contaminant is not too surprising considering the differences in species, season, and exposure

(concentration/duration). Interpretation becomes even

more challenging when considering the natural variation in vtg synthesis and circulation that occurs

between fish of the same species, sex, and geographic

location. Sumpter and Jobling (1995) reported plasma vtg inductions ranging from 500-fold to over

50,000-fold in rainbow trout maintained in effluent

from several neighboring sewage treatment sites.

Recent reports document enhanced vtg levels

in both male and female fish exposed to various manmade and naturally occurring environmental agents.

Some of the most compelling evidence for chemicalinduced vtg production comes from field studies

investigating the effects of sewage effluent and other

chemicals on male and female fish residing downstream of treatment facilities (Bevans et al. 1996;

Purdom et al. 1994; Sumpter and Jobling 1995). In

one report, a 1000-fold increase in plasma vtg was

observed in fish exposed for only three weeks to

sewage effluent (Purdom et al. 1994). Corresponding

laboratory experiments verified the induction of vtg

following brief exposure to extremely low levels (0.1

ng/l) of 17β-ethinylestradiol, a potent estrogen and

common constituent of sewage effluent (Purdom et al.

1994). Elevated plasma vtg levels have also been

reported in response to a number of other environmental contaminants, including chlordecone, o,p’-

DDT, and o,p’-dichlorodiphenylethylene (DDE)

(Donohoe and Curtis 1996), dioxin (von der Decken

et al. 1992), polluted harbor dredged spoil (Janssen et

al. 1997), and the pesticide atrazine (Fig. 19) (Grady

et al. 1998). Although the potentially harmful effects

of environmental estrogens have received considerable attention, alterations in endocrine and reproductive functions have also been documented in fish

exposed to a number of metals and inorganic molecules and compounds. Reductions in plasma vtg concentrations were reported in brook trout (Salvelinus

fontinalis) exposed to various combinations of acid,

aluminum, and calcium (Mount et al. 1988), winter

flounder (Pleuronectes americanus) treated with cadmium (Pereira et al. 1993), and walking catfish

(Clarias batrachus) injected with lead, zinc, and mercuric acetate (Panigrahi et al. 1990).

Value and Utility of Vitellogenin in the BEST Program

The measurement of vtg levels in female and male

fish provides an additional biomarker for determining

the stage of maturation, assessing reproductive health,

and predicting the estrogenicity of various compounds; however, the researcher should be cautious,

as a number of studies have reported considerable

variability in both sexes. The literature also documents that the vitellogenic response is dependent on a

number of intrinsic and extrinsic factors.

Although female vtg production can be

affected by contaminants, plasma concentrations are

so variable that it is difficult to determine the significance. Thus, vtg production by male fish became the

focus of a number of investigations. Various studies

led to the assumption that healthy males would not

produce vtg and, therefore, the detection of plasma

vtg in males would serve as a reliable biomarker of

exposure to environmental estrogens (Palmer and

36 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Figure 19. Mean plasma vitellogenin concentrations for male (blue)

and female (red) largemouth bass following a 20-d exposure to

atrazine.
Selcer 1996; Heppell et al. 1995; Sumpter and Jobling

1995). According to the literature, most males do not

produce vtg in measurable quantities; nevertheless,

vtg receptors have been detected in testes, muscle,

and spermatocytes (Bidwell and Carlson 1995; Tao et

al. 1996). Males treated with estrogen can produce

significant concentrations of plasma vtg (Bromage

and Cumaranatunga 1988; Christiansen et al. 1998),

yet studies have also detected various levels of plasma vtg in untreated male fish from reference sites

(Goodbred et al. 1997; Denslow et al. 1998). Given

that the biological significance of vtg in males is currently unknown, correlating vtg production with

endocrine disruption or feminization is risky in the

absence of data corroborating such conditions (e.g.,

sex steroid measures, histopathology). Detrimental

effects on wild populations cannot currently be

ascribed to the unnatural or untimely production of

vtg in adult males. Evidence of injury from vitellogenesis has only been seen in juvenile male rainbow

trout that suffered renal problems, and ultimately

death, after vitellogenesis was induced by pharmacological doses of 17β-estradiol (Herman and Kincaid

1988). 

The development of a universal assay for

detecting vtg by immunological methods has been

complicated by the fact that the primary amino acid

sequence is not highly conserved among species of

fish (Campbell and Idler 1980; So et al. 1985; Benfey

et al 1989; Lee et al. 1992). Nevertheless, monoclonal antibodies have been developed that cross-react

with a wide range of species. For instance the monoclonal antibody, mAb 2D8 (Heppell et al. 1995)

showed wide taxonomic specificity and cross-reacted

with vtg from all species tested. But, because it was

an IgM and of relatively low affinity, its usefulness

for an assay was limited. Other monoclonal antibodies of the IgG class have been prepared that are of

high affinity and very specific for vtg (Denslow et al.

1996; Denslow et al. 1997). These monoclonal antibodies are not “universal” but they do cross-react

with vtg of fishes from diverse families, making them

useful for evaluating vtg induction in the wild. In

summary, vtg assays are sensitive and reasonably

inexpensive; however, interpretation of findings, especially when less than 25% in a population have low

amounts (<10 μg/ml), remains somewhat challenging.

As is true of many emerging biomarkers, a better

understanding of factors that either influence or inhibit vtg production in both males and females would

improve the utility of vtg as an indicator of chemical

exposure and reproductive health.

GONADO SOMATIC INDEX (GSI) AND GONADAL

HISTOPATHOLOGY (INCLUDING GONADAL STAGING)

The GSI and gonadal histopathology fall into a category of indicators that provide structural, rather than

functional, information about gonadal health and maturational stage. The GSI is one of several organosomatic indices, including the HSI and SSI, which

establishes a ponderal relationship between the organ

and the entire body. There is substantial evidence that

most animal species undergo reproductive cycling

and, frequently, dramatic variation in gonadal size is

observed throughout the cycle (de Vlaming et al.

1981). Consequently, calculating gonadal weight as a

percentage of body weight has routinely been used to

determine reproductive maturity, as well as assess

gonadal changes in response to environmental dynamics (e.g., seasonal changes) or exogenous stresses

(e.g., contaminant exposure).

Gonadal histopathology is often utilized

alone, or in conjunction with the GSI, to confirm

gonadal phenotype, determine the state of sexual

development, and investigate reproductive impairment. Although gonadal histopathology is routinely

used to detect higher level responses expressed as

morphological abnormalities, such as the presence of

ovotestes (Fig. 20) or multinuclear ova, this method is

capable of providing information at multiple levels of

biological organization (i.e., distribution of molecules;

distribution, number, volume, morphology of

organelles, cells, and organs). Observed alterations in

cells and tissues are often reflective of previous biochemical and physiological modifications. 

Background

The utilization of the GSI as a reproductive biomarker

was first reported in 1927 in a study describing the

yearly variations of female yellow perch (Perca

flavescens) ovaries (Meien 1927). Years later,

Nikolsky (1963) endorsed this method on the premise

that “...the effects of fish size on gonadal weight are

eliminated by expressing gonadal weight as a percentage of body weight.” There is significant evidence

that exposure to various environmental pollutants can

result in gonadal alterations such as a decreased GSI,

morphological changes, or both.

Measuring the GSI and Performing Gonadal

Histopathology

Measuring the GSI is a convenient and frequently

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 37
employed method that demands little time and few

instruments. The GSI is calculated as (gonad

weight/body weight) * 100. The only requirement is

that the measurements be made on live or freshly

killed specimens to avoid weight fluctuations induced

by storage conditions (i.e., moisture loss/gain).

Evaluating the condition of the gonads by histopathology is more labor intensive and requires experienced

personnel. Briefly, isolated gonads are fixed with

Bouins’ or Smith’s solution, a glutaraldehydeparaformaldehyde solution, or some other preservative (e.g., NoTox or 10% neutral buffered formalin).

Dehydration of the samples is performed by transferring them through a series of graded ethanol solutions

(50-99%). Testes are cut longitudinally and ovaries

are cut transversely before embedding in paraffin and

thin-sectioning (0.5 μm). Prior to evaluation by light

microscopy, the samples are treated with one or more

staining compounds, such as hematoxylin, which

stains nuclear elements, and eosin, which labels general cytoplasmic and intracellular materials.

To evaluate gonadal stage several classification schemes based on histological examination have

been described and applied to the evaluation of

gonadal development and the seasonal activity

changes that occur in fish. The classification system

adopted by Goodbred et al. (1997) describes four maturational stages for females and males. For females,

stage 0 describes inactive ovaries containing primarily

perinucleolar (primary or secondary) oocytes with no

developing ova. The previtellogenic stage, or stage 1,

contains a mixture of perinucleolar and cortical alveolar oocytes. Slightly to moderately enlarged ova containing vacuoles or lipid droplets, but few or no

vitelline granules, are representative of this stage.

Ovaries classified as early vitellogenic (stage 2) contain oocytes of various sizes and development. Few

(or no) fully developed ova exist, although moderate

numbers of vitelline granules may be present. The

vitelline granules are storage compartments composed

of lipid-bound vitellogenin fragments. Ovaries in the

final stage of sexual development, classified as late

vitellogenic (stage 3), contain ooyctes approaching

maximum size that consist of numerous densely

packed vitelline granules. Although stage 3 is typical

of fish approaching spawning, stage 1, 2, and 3 are all

representative of sexually mature female fish.

Wallace and Selman’s version of teleostian oocyte

maturation is similar to that above, although they

describe six stages of development: the chromatinnucleolus phase, perinuclear phase, cortical alveoli

phase, vitellogenic phase, maturation phase, and ovulation phase (Wallace and Selman 1990). 

The classification system adopted by the

BEST Program for common carp (Cyprinus carpio)

and black basses is similar to that of Wallace and

Selman (1990) in that it recognizes six developmental

stages for females and four maturational stages for

males. The stages are based on the size and developmental status of the oocytes (ovaries) and spermatozoa (testes) contained within the reproductive organs

(Table 6). In females, stage 0 (immature) ovaries

contain pre-vitellogenic oocytes in the chromatinnucleolus and early perinucleolus stages. Stage 1

ovaries contain oocytes in early development; more

than 90% are pre-vitellogenic and the remainder are

early to mid-vitellogenic. The early vitellogenic

oocytes are in the late perinucleolus through cortical

alveolar stages and are slightly larger than the previtellogenic oocytes. The mid-vitellogenic oocytes

contain yolk vesicles, which appear as globules

around the periphery of the cytoplasm. The chorion

is uniformly apparent and the oocyte is larger. Stage

2 ovaries contain oocytes in mid development. In this

stage the majority of developing follicles are early to

mid-vitellogenic. In late development, or Stage 3, the

majority of developing follicles are late vitellogenic,

which is characterized by increased oocyte diameter

and chorion thickness and yolk globules distributed

throughout the cytoplasm. Stage 4 is the late development/hydrated stage in which the majority of developing follicles are late vitellogenic and significantly

larger than stage 3 follicles. Stage 5, or the post-ovulatory stage, is characterized by the presence of spent

follicles marked by remnants of the theca externa and

granulosa. Although stage 3 is typical of fish

approaching spawning, stages 1, 2, and 3 are all representative of sexually mature female fish. It should

be noted that the classification schemes for females

described here are relatively general, and all stages

may not be evident in females of every species.

For males, the gonadal stages recognized in

fish collected by the BEST program (Table 6) are for

all practical purposes identical to those used by

Goodbred et al. (1997). The stages are based on the

maturity of the predominant stage of spermatogenesis

(Nagahama 1983). Stage 0 (immature) in males is

recognized by the absence of spermatogenic activity

in the germinal epithelium and the presence of primarily spermatocytes. No spermatozoa are present in the

tubules. Testes in early spermatogenesis (stage 1) are

characterized by mostly thin germinal epithelium and

the presence of primarily immature cells (spermatocytes to spermatids); however, some spermatozoa are

also present. In mid-spermatogenesis (stage 2) the

germinal epithelium is moderately thick and some

proliferation and maturation of sperm can be

observed; spermatocytes, spermatids, and spermato38 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
zoa are present in roughly equal proportion. In late

spermatogenesis (stage 3) the germinal epithelium is

thick. Although all cell types are represented, spermatozoa predominate in stage 3. Stages 1 through 3

are characteristic of sexually mature fish, with the

least activity occurring in off-season (stage 1) and the

most activity taking place immediately prior to and

during the spawning season (stage 3).

Factors That Can Affect the GSI and Gonadal

Histopathology

As noted above, both the GSI and gonadal

histopathology are affected by season, which controls

reproductive cycling. The effect of age on the GSI

and gonadal histopathology is reviewed in a recent

article by Patnaik et al. (1994). Gender influences the

GSI with males experiencing less gonadal weight gain

during recrudescence than females. Males also have

less well-defined stages of maturation (Kime 1995).

Variability in these parameters also exists among

species. For example, goldfish and other small

cyprinids are fractional spawners (Delahunty and de

Vlaming 1980), depositing eggs during isolated

episodes over the course of several months (Yamazaki

1962). Unlike species that spawn once a year, or

once during a lifecycle, the GSI and gonadal structure

change more frequently as the fractional spawners

undergo repeated oocyte maturation and release.

Pollutants may also cause alterations in these

two indicators. A reduction in the GSI and impaired

gonadal development (growth and structural pathologies) have been reported in response to environmentally relevant doses of dietary mercury in juvenile

walleye (Stizostedion vitreum) (Friedmann et al.

1996), organophosphate insecticides in female striped

catfish (Mystus vittatus) (Choudhury et al. 1993), and

metacid-50 and carbaryl in climbing perch (Anabas

testudineus) (Haider and Upadhyaya 1985). In a field

study, fathead minnows (Pimephales promelas) were

subjected to acidic water conditions and oocyte atresia was detected histologically (McCormick et al.

1989). Oocyte atresia, as defined by an involution or

resorption of unfertilized eggs by the ovaries, is a normal physiological event in all fish, but it has become

a pathological condition noted in fish after exposure

to certain environmental contaminants (Fig. 21)

(Johnson et al. 1988; Cross and Hose 1988; Cross and

Hose 1989; Kirubagaran and Joy 1988). The ability

to detect increased degeneration or necrosis of developing oocytes by histological examination has

inspired the use of oocyte atresia as a biomarker of

reproductive impairment. 

Value and Utility of the GSI and Gonadal

Histopathology in the BEST Program

Over the years, the accuracy and reliability of the GSI

measurement have been scrutinized by some, due to

the variability observed between relative ovarian

weights even within the same or related species (de

Vlaming et al. 1981) and the inability of some field

studies to correlate tissue levels of contaminants or

plasma sex steroid levels with the GSI (Johnson et al.

1994; Monosson et al. 1994; Sepulveda et al. 1998).

Others suggest that a correlation between all levels of

organization is not necessary and that a negative

response from one marker (e.g., the GSI) does not

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 39

Figure 20. Histologic appearance of intersex fish. Left panel: Low magnification (165x) showing oocyte (A) development alongside sperm

(B) within maturing testicular tissue. Right panel: A higher magnifcation view (412x).
negate or lessen the significance of meaningful data

collected from a nonequivalent bioassay (e.g., sex

steroid analysis) (Huggett et al. 1992). 

Despite conflicting opinions regarding its

reliability, the GSI is easy and inexpensive to measure, and has remained a commonly used criterion for

evaluating the reproductive status and health of fish.

As with the evaluation of sex steroid hormones and

vitellogenin, interpretations of GSI measurements rely

on a thorough understanding of the natural variability

between fish of the same age, sex, and species, as

well as the environmental influences, behavioral patterns, and reproductive strategies that may complicate

or confound data. For example, in one study it was

shown that winter flounder remain offshore during

most of early vitellogenesis and move into contaminated estuaries prior to spawning, whereas English

sole (Parophrys vetulus) reside in contaminated estuaries during vitellogenesis and move offshore to

spawn. The migratory behaviors of the two species

influences the duration and timing of exposure to pollutants and, therefore, can have an indirect impact on

gonadal recrudescence (Johnson et al. 1994).

Many biomarkers vary seasonally with the

reproductive cycle. To interpret these biomarkers,

one must know the reproductive status of each fish,

which can be achieved by determining its gonadal

stage or condition histologically. Histological examination alone is also a valuable technique that has been

used to evaluate reproductive status, determine sex,

and assess the gonadal health of male and female fish.

In addition to the advantages described in the intro40 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Table 6. Histological features of gonadal stages as defined for use in the BEST program. Adapted from

Treasurer and Holiday, 1981, Nagahama 1983, Rodriquez et al. 1995 and Goodbred et al. 1997.

Stage Testis characteristics (male) Ovary characteristics (female)

0 Undeveloped (immature): Undeveloped:

Little or no spermatogenic activity in germinal Pre-vitellogenic oocytes observed exclusively;

epithelium; immature states of spermatogenesis oocyte diamter <250 μm: cytoplasm stains basophilic 

(largely spermatocytes); no spermatozoa observed. with H&E.

1 Early spermatogenic: Early development:

Mostly thin germinal epithelium with scattered >90% pre-vitellogenic, remaining oocytes early to

spermatogenic activity; spermatocytes to spermatids mid-vitellogenic; oocytes slightly larger (up to 300

predominate; few spermatozoa observed. μm); late perinucleolus through cortical alveolar 

stages.

2 Mid-spermatogenic: Mid-development:

Germinal epithelia are of moderate thickness; Majority of observed follicles are early and midmoderate proliferation and maturation of vitellogenic; oocytes larger, 300-600 μm diameter,

spermatozoa and equal mix of spermatocytes, and containing peripheral yolk vesicles;

spermatids and spermatozoa present. globular and uniformly thick chorion (5-10 μm in 

black basses, 10-20 μm in common carp); cytoplasm 

is basophillic, yolk globules eosinophillic.

3 Late spermatogenic: Late development:

Thick germinal spithelium; diffuse regions of Majority of developing follicles are late vitellogenic;

proliferation and maturation of spermatozoa; all oocyte diameter is 600-1000 μm; eosinophillic yolk 

stages of development are represented, but globules distributed throughout the cytoplasm;

spermatozoa predominate. chorion thickness is 10-30 μm in black basses, 

40-50 μm in common carp.

4 N/A Late development/hydrated:

Majority of developing follicles are late vitellogenic;

follicles are much larger (>1000 μm).

5 N/A Post-ovlulatory:

Spent follicles, remnants of the theca externa and

granulosa.
duction to this section, histopathology can often provide information regarding the magnitude, or occasionally the source, of toxic impairment. In addition,

this method may be used to study alterations in animals that are too small for standard biochemical

analyses. For accurate interpretation of histological

data, the investigator must have extensive knowledge

of the gross and microscopic anatomy of the specimen under investigation and must be aware of normal

variations in the anatomy of the species being examined, as well as differences that may occur due to

gender, age, and seasonal variations. In addition, the

investigator must be able to differentiate between

lesions of different origins, including exposure to

toxic chemicals, infectious diseases, congenital anomalies, and stress from handling. Using control animals in experimental studies will help distinguish

between normal morphology and lesions derived from

the experimental treatment. Although the subjectivity

of the examiner can also lead to variance, microcomputers and associated software programs now exist

that allow for the quantitative analysis of morphological data. To date, oocyte atresia is the only histological biomarker that has been sufficiently validated; that

is, lesions in laboratory studies have been correlated

to chemical exposure, and these same lesions detected

in fish from contaminated sites (Hinton et al. 1992).

Also, a potential correlation has been found between

oocyte atresia and reproductive success (McCormick

et al. 1989). A number of other lesions (e.g., testicular atrophy, sperm reduction) have potential as

histopathological biomarkers; however, field confirmation of laboratory results is needed before these

indicators can be used with confidence.

SUMMARY

At the biochemical level, the sex steroids respond

quickly to both intrinsic (physiological) and extrinsic

(environmental) stimuli, making these indicators rapid

and relatively easy to measure. They are often the

first detectable responses to an environmental change

or stress and can serve as indicators of both exposure

and effect. Furthermore, effects at the biochemical

level are highly sensitive and often underlie changes

observed at more complex levels of organization and,

thus, may be predictive of perturbations at the organ

or organism level. Although the high sensitivity of

the sex steroids can complicate data interpretation

when considering the number of factors (e.g., temperature, photoperoid, handling and drawing blood) that

can affect circulating levels, many of these problems

can be overcome in both laboratory and field studies. 

Vitellogenin is another highly sensitive biomarker with the potential to provide information

regarding exposure and effect at the biochemical

level. It was originally hypothesized that vtg production in males was indicative of estrogen exposure and

would serve as an effective biomarker of endocrine

disruption. Although this may be the case, recent

studies suggest that even healthy males may produce

low levels of vtg, and there is currently no evidence

that trace concentrations affect either general health

or reproduction. Vitellogenin has proven to be a valuable biomarker in carefully controlled laboratory

experiments as well as isolated field studies.

However, the successful application of vtg as a biomarker for males in large-scale field studies is contingent on understanding the normal patterns of vtg production in males. Until then, vtg may be a more reliMETHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 41

Figure 21. Atretic eggs (A) present in ovaries at two different stages of development. Left panel, 412x, right panel, 165x.
able biomarker in females. For instance, in males, vtg

has no known function, whereas in females, it is a

critical component of the developing oocyte.

Therefore, identifying females with low or undetectable vtg may be as important as detecting males

with vtg. Since all female oviparous fish produce vtg,

usually in the prespawning period, a lack of vtg may

indicate serious reproductive problems.

The GSI and histological examination provide structural information concerning the gonads and

appear to be advantageous methods for identifying

effects of long-term contaminant exposure. The GSI

is easy and inexpensive, although definitely crude in

comparison to the more informative, yet more laborintensive, histological exam. Both techniques require

knowledge of the variations due to age, species, and

season, and histology requires extensive knowledge

of the gross and microscopic anatomy of the specimen under investigation. Although it requires a qualified pathologist, information gained from gonadal

histopathology, such as sex and reproductive stage, is

essential for interpreting many of the other biomarker

responses. In addition, histopathology can provide

unique information from the visual examination of

gonadal tissue, including the assessment of oocyte

atresia and detection of ovotestes.

In addition to the reproductive biomarkers

discussed in this review, information regarding a variety of other parameters including fertility, clutch size,

hatchability, and sexual behavior are necessary for a

complete understanding of the reproductive mechanisms in healthy and compromised wildlife populations, which cannot be assessed at the geographic

scale represented by the BEST program. Currently,

the optimal approach for studying the reproductive

system at this scale involves using a suite of biomarkers to gain information at different levels of organization. Each of the above indicators has advantages and

disadvantages, but analyzing them together may help

to overcome the individual weaknesses of any one

test. Individually, the reproductive biomarkers can

provide only limited information regarding the effects

of environmental stressors. However, when used

together, the biomarkers described here can provide a

more holistic account of an animal’s reproductive

health, allowing investigators to evaluate whether a

species or population is at risk. 

STABLE NITROGEN ISOTOPE RATIO (δ

15

N)

Christopher J. Schmitt

The ratio (δ

15N) of the abundance of naturally occurring stable (non-radioactive) isotopes of nitrogen (15

N: 14N) in organisms reflects both nitrogen sources to the

ecosystem and trophic relationships within the

ecosystem. δ

15N increases predictably upward

through aquatic food chains at the rate of about 3-4%

per trophic level (DeNiro and Epstein 1978; DeNiro

and Epstein 1981; Minagawa and Wada 1984;

Peterson and Fry 1987; Kling et al. 1992; Cabana et

al. 1994; Michener and Schell 1994; Hobson and

Welch 1995; Vander Zanden and Rasmussen 1996;

Vander Zanden and Rasmussen 1999; Vander Zanden

et al. 1997; Vander Zanden et al. 1998) (Fig. 22). In

addition, nitrogen sources (i.e., fertilizers vs. animal

wastes) commonly differ in isotopic composition and

can be differentiated on the basis of their isotopic signature (e.g., Heaton 1986; Kendall 1998).

Background

Under ideal circumstances, stable nitrogen isotopes

offer a direct means of source identification because

the two major sources of nitrate in many agricultural

areas, fertilizer and manure, generally have isotopically distinct δ

15N values. Hence, under favorable condi42 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005

Figure 22. Mean (±s.e.) δ

15

N for components of the Lake Ontario

pelagic food chain. Myis relicta, pelagic forage fish (smelt, alewife,

sculpin) and lake trout are represented by, respectively, square, circle, diamond, triangle, and filled circle (From Cabana and

Rasmussen 1996; reproduced by permission). 
tions, the relative contributions of these two sources

to groundwater or surface water can be estimated by

simple mass balance. Soil-derived nitrate and fertilizer nitrate commonly have overlapping δ

15N values,

preventing their separation using δ

15N alone. 

Animal waste (including sewage) almost

always has a higher δ

15N than other nitrogen sources

[e.g., fertilizers and atmospheric deposition (Heaton

1986)]. Animal-derived nitrogen in biological systems is indicated by a relatively high δ

15N in organisms at all trophic levels as sewage nutrients enter and

pass through the food chain. Because of this property, δ

15N has been used to trace sewage-impacted

groundwater inputs into a coral reef in Barbados

(Cabana unpub. data). In this latter study, the indicator organisms were large sedentary invertebrates

(mussels, chitons) and non-migratory fishes - guilds

of organisms chosen for monitoring aquatic habitats

by the BEST program (BEST 1996).

Measuring Stable Nitrogen Isotopes

14N and 15N (and thus δ15N) can be readily measured

by mass spectrometry. Samples of dried, homogenized muscle or whole fish (about 2-3 mg dry weight)

are pulverized and introduced in small tin boats prior

to flash combustion, purification, and determination

of isotopic ratios by mass spectrometry (Cabana and

Rasmussen 1996).

Factors That Can Affect δ

15

N

An early attempt to use natural δ

15N values to determine sources of nitrate in surface waters (Kohl et al.

1971) received a highly critical response (Hauck et al.

1972). This was partly because the use of the δ

15

N

values of fertilizer and animal waste to trace their relative contributions to groundwater is complicated by

several reactions (e.g., ammonia volatilization, nitrification, denitrification, ion exchange, and plant

uptake) taking place within the hydrologic system that

can significantly modify the δ

15N values. Moreover,

mixing of point and non-point sources along shallow

flow paths makes determination of sources and extent

of denitrification very difficult. Because of these

problems, attempts to use δ

15N for tracing the source

and fate of nitrate in ground waters and surface

waters often have only limited success, despite the

moderately good separation of δ

15N values. 

In addition to the problems cited above,

other factors may also contribute to variability and

uncertainty in the use of δ

15N as a nitrogen source

tracer (Battaglin et al. 1997; Kendall 1998). These

include variation in the isotopic composition of different sources of N as well as post-depositional recycling N (e.g., volatilization of ammonia, nitrification,

denitrification, assimilation, etc.). In addition, there is

the obvious confounding effect caused by trophic

dynamic differences among ecosystems. Cabana and

Rasmussen (1996) showed that in Ontario (Canada)

lakes, δ

15N in organisms increases with trophic position and is heavily influenced by land use, in particular by inputs of sewage (a correlate of human population density). 

δ

15N may also change over the life of an

organism, reflecting changes in feeding habits and

trophic position. In smallmouth bass (Micropterus

dolomieu), for example, δ

15N in eggs were similar to

adult levels, reflecting the trophic position of the parent. Levels decreased steadily from hatching through

swim-up and metamorphosis, then began increasing

after metamorphosis as the fish began to feed on progressively larger organisms (Vander Zanden et al.

1998). 

Value and Utility of δ

15

N in the BEST Program

Because it indicates the trophic position of the organisms being collected, δ

15N can be used as a corollary

variable for interpreting otherwise unexplainable differences in organism concentrations of biomagnifying

contaminants (e.g., mercury; polychlorinated

biphenyls; Fig. 23). This method therefore allows

one to account for site-specific differences in trophic

position within and among species [i.e., between

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 43

Figure 23. Relationship between mean mercury concentration (wet

weight) and mean δ

15

N in lake trout in seven Ontario and Quebec

lakes (From Cabana and Rasmussen 1996; reproduced by permission).
smallmouth bass and largemouth bass (Micropterus

salmoides) at the same and different sites] and to

reduce the confounding effect of collecting different

taxa at stations (Kiriluk et al. 1995; Kidd et al. 1995).

Many studies (Purdom et al. 1994; Goodbred et al.

1997; Folmar et al. 1996; Bevans et al. 1996) have

also shown that endocrine-modulating substances

(e.g., alkylphenols, pharmaceuticals, phytoestrogens)

are released from sewage treatment plants. In addition, recent evidence (Halling-Sørensen et al. 1998)

suggests that pharmaceuticals are also released to surface waters from confined animal feeding operations

(i.e., feedlots and pork and poultry production facilities). By serving as a means of differentiating animal

waste (including human sewage) inputs from fertilizer

inputs, δ

15N may be useful for interpreting the results

of reproductive biomarkers. δ

15N may also suggest

causal mechanisms behind certain fish kills and disease outbreaks; animal wastes are believed to be

involved in outbreaks of Pfisteria piscicida and other

toxic dinoflagellates responsible for recent, massive

fish kills in Mid-Atlantic and Southeastern estuaries

(Burkholder et al. 1995). δ

15N also represents a surrogate method for identifying and evaluating nutrient

sources, a class of contaminants of great current concern in the Mississippi River basin because agriculturally derived nitrogen is believed to be responsible for

the eutrophication of the Gulf of Mexico and the

occurrence of anoxic zones (Rabalais et al. 1996;

Battaglin et al. 1997). Excess nutrients are also often

involved in nuisance blooms of cyanobacteria

(Nodularia, Anabaena, Aphanizomenon, Microcystis)

and dinoflagellates (e.g., Protogonyaulax temorensis)

that produce potent hepatotoxins and neurotoxins

known to affect fish and wildlife (Gosselin et al.

1989; Kotak et al. 1993; Carmichael 1994). Elevated

δ

15N in aquatic organisms therefore indicates animalderived nitrogen. The confounding effects of trophic

position must be understood and accounted for, but

the measurement of δ

15N represents a potential tool

for identifying pollution impacts on water quality and

for explaining temporal and geographic variability in

monitoring studies. 

44 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
APPENDIX I: DEVELOPMENT OF THE BEST PROGRAM

Christopher J. Schmitt
THE NATIONAL CONTAMINANT BIOMONITORING PROGRAM

The publication of Silent Spring (Carson 1962) called

public attention to the environmental threat represented by dichlorodiphenyltrichloroethane (DDT) and,

later, other persistent toxins that have the potential to

bioaccumulate and harm wildlife. Member agencies

of the Federal Committee on Pesticides responded by

implementing the National Pesticide Monitoring

Program (NPMP) to obtain information on concentrations and distributions of these substances (Geary

1967). Participating agencies designed and operated

monitoring networks based on those segments of the

environment for which each had responsibility. The

U.S. Fish and Wildlife Service (FWS) participated in

the NPMP by collecting and analyzing freshwater fish

and European starlings (Sturnus vulgaris) and by analyzing the wings of hunter-killed ducks (Johnson et al.

1967) under its National Contaminant Biomonitoring

Program (NCBP). Although the environmental matrices collected and analyzed differed among agencies,

all components of the program were designed with

the objective of monitoring temporal and geographic

trends in the concentrations of specific contaminants,

including DDT and mercury. Among the analytes

added later were other organochlorine pesticides, such

as dieldrin; polychlorinated biphenyls (PCBs) and

other toxic industrial chemicals; and inorganic contaminants, including arsenic and selenium. 

The NCBP, along with the NPMP-derived

networks of other agencies, was highly successful in

achieving the primary program objective of documenting the distribution of the monitored contaminants in space and time, thereby providing positive

feedback to the regulatory process. Collectively, the

networks had demonstrated that by the 1980s restrictions on the use and discharge of accumulative toxins

had generally reduced concentrations of organochlorine pesticides, PCBs, and some inorganic contaminants in biota (Prouty and Bunck 1986; Bunck et al.

1987; Wade et al. 1988; Schmitt and Brumbaugh

1990; Schmitt et al. 1990; Sericano et al. 1990a, b;

O’Connor 1991; Turgeon and Robertson 1995;

Schmitt and Bunck 1995; Schmitt et al. 1999), but

that there were still areas of the U.S. where these contaminants could represent a threat to fish and wildlife.

Fish Sampling Under the NPMP and NCBP 

The fish component of the NPMP-NCBP was

designed to monitor spatial and temporal trends in the

concentrations of bioaccumulable contaminants over

large expanses of the U.S. Fish were collected periodically from a national network of stations located at

or near key points, such as at the confluences of

major tributaries or below dams, in the larger rivers of

the 50 states and in the open waters of the Great

Lakes (Fig. A-1). Station locations were fixed only to

the extent that the general area of the water body to

be sampled was identified. Because there are no truly

ubiquitous fishes that can be sampled throughout the

United States, cooperators were instructed to collect

representative species from two guilds—bottomdwelling/feeding and piscivorous (Johnson et al.

1967). Species were selected from a list of preferred

taxa stratified by aquatic habitat type — -cold-, cool-,

or warm-water. Samples comprised 3-5 adult fish of

a single species. Because of the recognized need to

evaluate the transfer of organochlorine chemicals,

mercury, and other bioaccumulable contaminants to

higher trophic levels (i.e., piscivorous birds and mammals), fish were analyzed whole; for reasons of economy, the 3-5 individual fish comprising a sample

were composited for chemical analysis. The overriding assumption was that periodic collections and

analyses of whole fish from key locales would integrate large expanses of time and space with respect to

the concentrations of the measured contaminants.

Results could then be generalized to taxa other than

those collected and used to assess the risk to piscivorous fishes and wildlife represented by the contaminants measured in the fish.

Shortcomings of the NCBP Fish Network

By the mid-1980’s the FWS perceived that the NCBP

was no longer meeting its growing need for contaminant-related information. Although the program provided data and information on the general distribution

and concentrations of contaminants in the environment, the information was not focused on the waters,

lands, and species of greatest concern to the FWS.

The information produced by the program was generally perceived to be scientifically authoritative and

academically interesting, but only marginally relevant

to policy-makers due to the geographic, programmatic, taxonomic, and chemical shortcomings outlined

here.

Geographic, Programmatic, and Taxonomic

Shortcomings

As a result of government-wide reorganizations in the

early 1970’s, the regulation and management of hazardous substances in the environment became the

responsibility of the newly created U.S.

46 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
Environmental Protection Agency (EPA).

Concomitantly, the contaminant-related concerns of

the FWS became more focused on the occurrence and

effects of a larger array of contaminants in and on

trust resources, especially those specific lands and

organisms for which the FWS had received stewardship responsibility. These included the National

Wildlife Refuge System; federally-listed threatened

and endangered species, including the bald eagle

(Haliaeetus leucocephalis); migratory birds, including

waterfowl; and certain marine mammals, and anadromous and inter-jurisdictional fishes, such as Atlantic

salmon (Salmo salar) and Great Lakes lake trout

(Salvelinus namaycush). During the 1980’s, contaminant concerns were further focused on Trust

Resources by well-publicized wildlife kills, oils spills,

and the acquisition of contaminated lands and other

contaminant-related problems within the National

Wildlife Refuge System (U.S. Fish and Wildlife

Service 1986; U.S. General Accounting Office 1987;

U.S. Department of the Interior 1990). The FWS

believed that a comprehensive monitoring program

would provide contemporaneous information on background contaminant and ecological conditions of trust

resources, a necessary prerequisite to the documentation of contaminants and their effects.

Chemical Shortcomings

By the mid-1980’s, many of the bioaccumulable contaminants that the NCBP and its predecessor were

designed to monitor had been regulated, replaced, or

both, and these changes were reflected by the fish and

starling networks. As environmental concentrations

of organochlorine chemical residues, mercury, and

other contaminants declined in many media, concerns

increased about the risks posed by other classes of

chemicals, many of which had not been monitored

under NCBP and were not amenable to traditional

monitoring methods (Table A-1). Among these were

short-lived, second- and third-generation pesticides,

which were being used in large quantities in U.S.

agriculture. These concerns resulted in part from

increasingly frequent reports of fish and avian

wildlife kills related to the use of highly toxic, but

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 47

Figure A-1. The network of stations for the National Contaminant Biomonitoring Program.
short-lived, organophosphate, carbamate, and synthetic pyrethroid insecticides (Glaser 1995). In addition,

the so-called soft pesticides are highly toxic to valuable non-target insects, including honey bees (Apidae)

(Pimental et al. 1992). Herbicides such as atrazine

were also being used in unprecedented amounts and

had become widely distributed in the surface and

ground waters draining agricultural areas (Paulson et

al. 1993; Coupe et al. 1995). Herbicides have been

identified as potential threats to important non-target

plants, including the aquatic and terrestrial vegetation

upon which wildlife depends (Kemp et al. 1983;

Kemp et al. 1985; Fairchild et al. 1998). In addition,

some widely used herbicides have been implicated as

endocrine disruptors (Colborn et al. 1993; Goodbred

et al. 1997) and others are either acutely toxic to

aquatic organisms, contain toxic byproducts, or can

be converted to toxic metabolites (e.g., Pothuluri et al.

1991).

In addition to agricultural chemicals, industrial and consumer chemicals not accounted for by the

NCBP were also recognized as potential threats to

fish and wildlife. For example, oil has long been a

significant environmental pollutant (Schmitt 1999).

In addition to the obvious direct toxic effects of oil

spills, microliter quantities of oil are highly toxic to

developing avian embryos (Albers 1977; Hoffman

1979; Hoffman 1990). Much of the toxicity of oil

stems from the polycyclic aromatic hydrocarbons

(PAHs) it contains. PAHs, which are created by the

combustion of fossil fuels and by many industrial

processes, are believed responsible for epizootics of

tumors and other lesions in fish from many industrialized waterways (Malins et al. 1984; Baumann et al.

1991; Black and Baumann 1991). Like soft pesticides, PAHs and related compounds are difficult to

monitor by traditional analytical methods. The most

environmentally significant PAHs are either rapidly

metabolized following their uptake by vertebrates

(Baumann et al. 1982) or dissipate into the atmosphere, and chemical residues do not persist.

Several other classes of contaminants of

potentially great environmental significance are difficult to monitor and assess by traditional analytical

methods. Among these are phenols, alkylphenols,

and diphenylalkanes, some of which have been identified as endocrine disruptors (Colborn et al. 1993).

These groups include compounds such as phenol, a

common and widespread industrial intermediate;

cresol, a major component of creosote; the bisphenols

(including bisphenol-A), which are components of

most plastics (Perez et al. 1998); and the alkylphenols, which are surfactants discharged in large quantities from paper mills and sewage treatment facilities

(Naylor et al. 1992; Nimrod and Benson 1996).

Sewage discharges and agricultural activities also

release large amounts of nutrients (i.e., nitrogen and

phosphorous), which can stimulate nuisance blooms

of algae and other microorganisms (such as Pfisteria

spp. and Protogoneaulax spp.) that ultimately represent threats to human and ecological health (Rabalais

et al. 1996; Gosselin et al. 1989; Kotak et al. 1993;

Carmichael 1994; Burkholder et al. 1992; Burkholder

et al. 1995). Bacteria can also convert excess nitrogen to ammonia, which can be acutely toxic to aquatic organisms (Wildhaber and Schmitt 1996).

Munitions and chemicals associated with their manufacture contaminate some waterways and lands formerly associated with military activities. More

recently, pharmaceuticals used in human and animal

medicine have been found in public waterways and

identified as environmental contaminants of potentially great significance (Steger-Hartmann et al. 1997;

Hartmann et al. 1998; Halling-Sørensen et al. 1998;

Buser et al. 1998).

Despite heightened concerns about new-generation pesticides and other classes of contaminants,

questions about the ecological risks of regulated contaminants remain. Although environmental concentrations of many persistent elemental and organochlorine contaminants had declined by the 1980’s, a growing body of information indicates that levels of these

regulated contaminants are still sufficiently high in

some areas to harm fish and wildlife in many important ecosystems, such as the Great Lakes (e.g.,

Colborn 1991). Organochlorine pesticide residues

remain evident in some ecosystems where these compounds were used heavily in agriculture and in other

applications, and in areas contaminated by sites of

synthesis or formulation (Schmitt and Bunck 1995;

Schmitt et al. 1999). Plasticizers such as di-n-butylpthatlate and bisphenol-A are ubiquitous environmental contaminants that can interfere with the control of

reproduction by the endocrine system (Colborn and

Clement 1992; Colborn et al. 1993; Perez et al. 1998).

In addition, elemental contaminants (i.e., metals and

metalloids) still constitute a threat from past and present mining activities, the use of organo-metallic pesticides and defoliants, and the continued growth of irrigated agriculture (Schmitt and Bunck 1995; Schmitt

et al. 1999; Schmitt 1999). Lead remains a global

pollutant from its historic (in the U.S.) and continued

(elsewhere) use as a motor fuel and paint additive,

and from mining, smelting, and other metallurgical

activities (Settle and Patterson 1980; Nriagu 1990).

Complex mixtures of polyhalogenated hydrocarbons

(PHHs), such as PCBs, chlorinated dibenzo-p-dioxins

(dioxins) and dibenzofurans (furans), and the insecti48 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
Contaminants

Common Mixtures, Congeners,

Compounds, Elements Use/Occurance 

PHHs

Organochlorine pesticides DDT and its metabolites, dieldrin, 

aldrin, endrin, mirex, chlordane, toxaphene, kelthane, HCB, BHC

Insecticides; fungicides; acaracides; rodenticides

Toxicendoc

PCBs PCBs 77, 118, 126, 153, 169;

Aroclors 1016, 1242, 1248, 1254, 1260

Dielectric, heat transfer, and hydraulic fluids; lubricants; plasticizers; copy papers

Cancedisrupdevel

Halogenated diphenyl ethers Ugilec-141 PCB replacement compounds Cance

disrupdevel

PCDDs, PCDFs 2,3,7,8-TCDD, 2,3,7,8-TCDF Impurities; combustion products; 

chlorine bleaching process

Cancedisrupdevel

Phthalates Di-n-butyl phthalate Plasticizer Toxic

endoc

Elemental Contaminants

Heavy metalsOther toxic metalsMetalloids

Hg, Cd, PbAl, SnAs, Se

Natural constituents of the earth's crust; elevated concentrations occur as a result of mining and smelting; agriculture; metal manufacturing, fabrication, and finishing; pesticides and other chemicals; fossil fuel combustion 

All canecesconcemediu

Herbicides Atrazine, 2,4-D, Dacthal®, simazine, 

cyanazine, propanil

Most heavily-used class of agricultural chemicals in the U.S.

Contamay dmetabdisrupnon-t

Table A-1. Contaminants of concern to the BEST Program (not mutually exclusive).
Table A-1 continued.

Contaminants

Common Mixtures, Congeners,

Compounds, Elements Use/Occurance Biological 

New-generation Pesticides

OrganophosphatesCarbamatesSynthetic pyrethroids

Parathion, diazinonCarbofuran, methyl carbamatePermethrin, esfenvalerate

Insecticides; fungicides Avian and wildlife 

highly toxic to invaquatic organisms

PAHs Chrysene, benz(a)pyrene, 

benz(a)anthracene, naphthalene, fluorene

Present in petroleum creosote; formed from combustion of fossil fuels

Carcinogenic; teramutagenic; some atoxic

Phenols

Nonylphenol,Polyethoxylates (APEs),Cresol, phenol

Surfactants; discharged from publiclyowned treatment works (POTWs) and paper or textile mills; creosote; industrial intermediates; plasticizers

Endocrine disrupt

Diphenylalkanes

Bisphenols Bisphenol A Plasticizer Endocrine disrupt

Pharmaceuticals

Synthetic estrogensAntibioticsAnalgesicsChemotheraputic agentsOthers

Ethenyl estradiolFluoroquinelonesIbuprofenCyclophosphamideClofibric acid

Used in human and veterinary medicine; detected in sewage treatment plant effluent and receiving waters

Endocrine disrupt

Munitions Nitro-aromatic explosives and 

derivatives (TNT, DNT), white P, RDX

Military installations; production, testing, and use of weapons/explosives

Highly toxic; muta

Nutrients and biotoxins Nitrogen, phosphorus Agriculture (fertilizers, manure); 

sewage; manufacture of explosives

Stimulate nuisancalgae and other aqcyanobacteria, and dinoflagellates thabiotoxins and caube converted to hicompounds (nitroammonia, sulfur to sulfide)

Hydrogen ions Mostly sulfuric acid Sulfur (as SO2) released to the 

atmosphere from the combustion of fossil fuels forms sulfuric acid, which is responsible for acid rain. Acid mine drainage from coal and metals mining has also resulted in significant biological consequences.

Highly toxic; signin community comand related variabimportant modulacontaminant toxicfor metals.
cides toxaphene and chlordane, are especially problematic. The components of these mixtures differ

greatly in toxicity and persistence, and there is evidence that some of the most toxic constituents and

metabolites are also among the most resistant to

degradation (Gooch and Matsamura 1987; Tillitt et al.

1992). Moreover, no large-scale monitoring program

has yet incorporated the periodic assessment of the

dioxins and furans owing to the great expense and

difficulty of their analysis. Local-, regional-, and

national-scale investigations have shown that these

and other highly toxic and accumulative PHHs, as

well as other structurally and toxicologically similar

compounds, are present in many U.S. ecosystems at

concentrations either actually or potentially harmful

to fish and wildlife (Colborn 1991; Kubiak et al.

1989; Keuhl et al. 1989; Tillitt et al. 1992; Mac and

Edsall 1991; U.S. Environmental Protection Agency

1992). In addition, mercury concentrations in some

remote areas are rising (Monteiro and Furness 1998),

ostensibly from the atmospheric transport of mercury

released during the combustion of coal (U.S.

Environmental Protection Agency 1997). For these

reasons, concentrations of bioaccumulative contaminants in fish have been proposed as an indicator of

sustainable economic development by the U.S.

Council on Environmental Quality (U.S. Council on

Environmental Quality 1997) and remains an important component of most environmental monitoring

programs (e.g., Hirsch et al. 1988; Messer et al.

1991).

FEATURES OF THE BEST PROGRAM

The intent of the BEST program is to build on the

success of the NCBP while addressing some of its

identified shortcomings by monitoring for a broad

array of contaminants and their effects and framing

the findings in terms of Department of the Interior

(DOI) issues. Because of the large number of contaminants not amenable to monitoring by traditional

analytical methods, the basic premise of measuring

only chemical concentrations in selected environmental media is now perceived as inadequate. In addition, chemical concentrations alone are not useful for

assessing the cumulative effects of multiple contaminants or the cumulative effects of contaminants and

other factors such as sedimentation, eutrophication,

and disease. It is widely known that contaminants

and other stressors tend to co-occur as a consequence

of land-use and industrial practices (e.g., Schmitt

1999). Assessing the cumulative effects of multiple

contaminants and differentiating contaminant-induced

effects from those caused by other factors, information that is necessary to initiate and evaluate management actions, requires more than just the concentrations of a relatively small number of contaminants in

habitats and organisms.

Weight-of-Evidence Approach

The BEST program documents temporal and spatial

trends in the exposure of organisms and ecosystems

to contaminants and the effects of exposure on selected organisms through the application of chemical and

biological monitoring methods. The monitoring

methods comprise scientifically well-founded and relatively inexpensive techniques chosen to maximize

the number of samples that can be analyzed and,

therefore, the number of sites that can be monitored.

The results of these monitoring activities may stimulate research targeted towards confirming and diagnosing the nature, extent, and significance of the findings using more robust, sensitive, and expensive

methods; however, this research is beyond the scope

of the BEST program. Through this approach, the

program hopes to overcome some of the geographic

and chemical shortcomings of the NCBP by sampling

a wider variety of habitats using screening methods

that are sensitive to many classes of chemicals, and

restricting the application of more definitive methods

to focused, follow-up research. “Workhorse” methods that can be applied to many taxa and guilds of

organisms in a wide variety of habitats, and at spatial

scales ranging from local to regional and national, are

employed. By using methods consistently across

ecosystems, guilds, and spatial scales, the findings

can be aggregated for the analysis and interpretation

of spatial and temporal trends. Unlike the NCBP,

which relied solely on the measurement of contaminant concentrations in fish and wildlife, the BEST

program incorporates complementary biological

methods spanning many levels of biological organization, as well as information from extant sources. This

strategy, referred to collectively as a weight-of-evidence approach, represents a holistic, integrated

assessment of the interactions between physical and

chemical factors and living organisms. The National

Water Quality Assessment Program of the USGS

Water Resources Division (Hirsch et al. 1988) and the

Environmental Monitoring and Assessment Program

of the U.S. EPA (Messer et al. 1991), both of which

evolved contemporaneously with the BEST program,

are also based on weight-of-evidence approaches. So

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 51
is the Sediment Quality Triad (Chapman et al. 1991;

Chapman 1992), a procedure in which sediments are

evaluated through chemical analysis, toxicity testing,

and benthic macroinvertebrate community composition. Through the application of this approach and

the minimization of redundancies among the methods

employed, the BEST program can monitor a wide

array of habitats and ecosystems for many classes of

contaminants in an efficient and cost-effective manner.

Organisms and Habitats Sampled

To remedy some of the taxonomic and programmatic

deficiencies of the NCBP, the BEST program bases its

measurements on ecosystems, habitats, and organisms

of interest to DOI agencies. Sampling of both organisms and habitats ensures that the exposure of mobile

species (especially birds and fishes) is documented.

For chemical analyses and the application of organism- and population-level methods, guilds of organisms were selected for sampling on the basis of the

following criteria:

(1) They have a high potential for exposure and

a measurable response (including the accumulation of tissue residues) to contaminants

of concern.

(2) They have a territory or home range that 

overlaps the area to be monitored.

(3) They are sufficiently large and abundant to 

permit easy enumeration and sampling.

(4) They represent, either in actuality or as a surrogate, a guild or species of programmatic 

importance to the DOI.

Guilds, which are groups of ecologically similar taxa,

were selected in lieu of particular species because no

fishes or birds that meet the above requirements have

distributions that cover all lands or waterways monitored under the BEST program. The guilds have been

selected in such a manner as to ensure that the maximum number of chemical classes and methods can be

accommodated with the smallest number of taxa and,

hence, collections and analyses. The identification of

guilds rather than species also allows substitutions for

species of programmatic importance to DOI agencies

in the event that a species of concern cannot be sampled [i.e., ospreys (Pandion haliaetus) or cormorants

(Phalacrocorax spp) substituted for eagles; common

carp (Cyprinus carpio) for rare cyprinids; rainbow

trout (Oncorhynchus mykiss) for rare salmonids].

The guilds selected for sampling have been

chosen to ensure that all classes of bioaccumulable

contaminants will be accounted for should they occur,

but that the BEST program will also be sensitive to

the occurrence of other contaminant classes that

might not accumulate. Two pairs of guilds—primary

and secondary—have been identified for the application of analytical chemistry, biomarker, and population health indicators in aquatic and estuarine habitats

(Table A-2), with the secondary guild to be sampled

in the event that no taxa representing the primary

guilds are present at a site.

Table A-2. Primary and secondary guilds identified for sampling aquatic habitats. From BEST Program (1996).

Primary Secondary Major 

guild guild contam.

classes

Piscivorous Piscivorous Bio-1

bird fish accum.

Sediment- Filter-feeding PAH, 

dwelling fish invertebrate other

non-bioaccum.2

1

PHHs, mercury, selenium, arsenic

2

PAHs and elemental contaminants that do not bioaccumulate; and

sediment-bound organic compounds not accounted for analytically 

These guilds were selected because of the

wealth of information that continues to show that piscivorous fishes and wildlife remain at greatest risk to

contaminants and their effects in aquatic ecosystems.

Because of the recognized importance of the PAHs as

environmental pollutants and the fact that most toxic

PAHs are rapidly metabolized by vertebrates and do

not accumulate, sediment-dwelling organisms are

included in both levels of the matrix. Other classes of

contaminants can be accounted for by sampling different media and by other methods, such as toxicity

tests and community analyses.

The aquatic habitats selected for sampling by

the BEST program have been classified hierarchically

(Table A-3). The classes were selected primarily on

the basis of the guilds of organisms likely to be found

52 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
and the methods that would be used to sample and

test them (BEST 1996).

Table A-3. Habitat classification system used to identify

habitats for sampling. From BEST Program (1996).

Estuarine

Open water

Vegetated (wetland)

Freshwater

Flowing

Large river (i.e., not wadeable)

Wadeable stream

Standing

Open water

Vegetated wetland

Marine 

(open water only habitat considered)

Monitoring Methods

The BEST program employs a suite of monitoring

and assessment methods chosen to address many

classes of contaminants and their effects. The suite

includes traditional chemical analyses of selected

media (animal carcasses, sediments) for contaminants

that persist sufficiently long and accumulate to concentrations that permit their periodic measurement at

reasonable cost; and biological indicators spanning

many levels of biological organization — from

molecular through population and community — that

integrate the effects of multiple contaminants and

other stressors. The biological indicators can also

document the exposure of organisms and ecosystems

to new or previously unrecognized chemicals as well

as chemicals for which analytical methods are either

not currently available or that are too expensive to be

incorporated into a large monitoring program.m.

Both general and specific indicators have

been selected so that the BEST program can provide a

degree of early-warning and can assess the cause(s) of

biological effects. Methods specific to individual

contaminants or specific classes of contaminants are

more sensitive and respond more rapidly than general

methods, and can consequently provide a warning of

impending or incipient problems from particular contaminants. In addition, their specificity is of obvious

value for the assessment of causation. Conversely,

general methods can provide information about contaminants for which specific methods either do not

exist or are not useful in a monitoring context.

General methods such as those indicative of organism, population, or community health can also be

used to assess the cumulative effects of multiple contaminants and the combined effects of contaminants

and other factors such as sedimentation, eutrophication, and disease. Such methods may be the most

effective way to monitor the exposure of organisms or

ecosystems to ephemeral contaminants.

To meet its stated objectives, the BEST program considers methods acceptable if they meet the

following criteria, which were derived from the DOI’s

Damage Assessment Guidance Document (U.S.

Department of the Interior 1987):

1. The method is well documented in the peerreviewed scientific literature.

2. The method has been shown to be well correlated with contaminants, other stressors, or

both across a gradient of concentrations —

both in the laboratory and in the field — and

has been shown to yield reproducible and

verifiable results.s.

3. The method has been used to a great enough

extent that data are available with which to

statistically document its performance.

4. Given 3, the method must be able to detect

relevant differences (between sites, timeperiod, taxa, etc.) at the desired level of significance (P <0.05) with financially realistic

sample sizes and sampling frequencies.

5. Confounding or interfering variables that

affect the performance of the method are

known, and the intended application of the

method takes these variables into account.

6. The application of the method has been documented in the scientific literature for the

species, medium, habitat, guild, or community for which its use is proposed.

The specific monitoring and assessment

methods being evaluated for use in aquatic habitats by

the BEST program have been drawn from three main

categories: analytical chemistry, toxicity tests and

bioassays, and biomarkers and indicators of organism

health. The BEST program also uses structural and

functional indicators of population and community

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 53
health, which are being evaluated through partnerships and leveraging with other investigations and

programs. Within each of these categories, methods

were proposed by working groups comprising experts

from the U.S. Geological Survey, other agencies, universities, and the private sector (BEST 1996).

Attributes of the categories are summarized here.

Analytical Chemistry

Assessing concentrations of toxic chemicals in environmental media and how their distribution changes

over time and space is the most common environmental monitoring technique. Although it is particularly

useful for those classes of contaminants that bioaccumulate or have long residence times, analytical chemistry alone cannot provide information on contaminant exposure and effect, for a number of reasons.

First, analyte concentrations in abiotic media may not

accurately portray risk because of bioavailability constraints. Consequently, the measurement of total metals or organic chemical residue concentrations in

some media—especially soils and sediments—may

have little relation to actual or even potential toxicity.

Similarly, contaminant concentrations in biotic media

do not necessarily provide direct measures of toxicity

because contaminants may be sequestered in fats or

other tissues, stored as non-toxic forms, or otherwise

immobilized and rendered non-toxic. Contaminants

may also be incorporated in recalcitrant tissues (bone,

scales, etc.) not amenable to digestion, or may occur

as solid-phase material either on the outside of organisms (Schmitt and Finger 1987) or inside, as gut contents (Brumbaugh and Kane 1985). These materials

may not be harmful to either the organism itself or to

higher trophic level organisms.

Although there is a growing body of information with which to assess the cumulative effects of

some classes of contaminants (e.g., Sprague and

Ramsay 1965; Enserink et al. 1991; Ahlborg 1989;

van den Berg et al. 1998) and complex mixtures in

sediments (Long and Morgan 1991; Long et al. 1995;

Swartz et al. 1995; Ingersoll et al. 1996; Wildhaber

and Schmitt 1996; Wildhaber and Schmitt 1998), the

ecological risk assessment (U.S. Environmental

Protection Agency 1992) of chemical residue data

remains centered largely at the level of the individual

contaminant. Due to differing requirements for sample handling and storage, the suite of analytes to be

considered must be pre-determined, which presumes

forehand knowledge of contaminants of concern. The

availability of analytical standards also tends to limit

analyses to parent compounds and their most abundant and persistent metabolites. Unfortunately, the

number of analytes for which methods and standards

are available represents only a small fraction of the

potentially toxic commercial and industrial substances

— as well as their precursors, metabolites, and

byproducts — released into the environment. Despite

these limitations, analytical chemistry remains attractive and necessary for three primary reasons: (1) The

sensitivity and relative precision of most analytical

chemistry enables the detection and periodic quantification of analytes at sublethal levels in a variety of

media, which provides important information for

managing and regulating impending or incipient problems; (2) As noted previously, the quantification of

contaminants in prey species and various environmental media provides information with which to assess

risk, albeit for individual or small numbers of contaminants; and (3) In combination with other evidence,

analytical chemistry forms the basis of cause-effect

analysis and risk assessment.

Bioassays and Toxicity Tests

Bioassays and toxicity tests provide direct evidence of

cumulative contaminant effects on the survival,

growth, behavior, or reproduction of living organisms

while controlling for extraneous confounding factors.

The results of such tests can also be extrapolated to

the population and community level and can be used

to differentiate contaminant effects from those

induced by other perturbations. Additionally, toxicity

tests and bioassays can provide cost-effective, rapid

assessments of cumulative toxicity, albeit at the

expense of chemical specificity. Toxicity tests and

bioassays can be performed in vitro, with cells or tissues from a variety of organisms (Fabacher 1982;

Tillitt et al. 1991; Zabel et al. 1995), or in vivo, with

whole organisms ranging in size from bacteria

(Johnson 1998; Johnson and Long 1998) to vertebrates (Finger and Bulak 1988; Hall et al. 1993).

Endpoints can range from a biochemical signal of

pollutant exposure or genetic damage (Fabacher 1982;

Tillitt et al. 1991; Johnson 1998) to the death or loss

of motility of organisms in a short-term test (Finger

and Bulak 1988; Hall et al. 1993; Wildhaber and

Schmitt 1996). Bioassays and toxicity tests, especially those conducted with whole organisms, are typically quite general with respect to the contaminants eliciting the response; however, they can also provide

more specific information on the nature and identity

of the substances involved when multiple tests are

conducted with organisms of differing susceptibility

to certain contaminants (Ingersoll et al. 1992;

Wildhaber and Schmitt 1996). Greater resolution can

also be obtained when toxicity tests are combined

54 INFORMATION AND TECHNOLOGY REPORT USGS/BRD/ITR 2000-0005
with reductionist approaches such as Toxicity

Identification and Evaluation (Mount 1988; Mount

and Anderson-Carnahan 1988) or by selectively sampling (Huckins et al. 1993; Johnson 1998; Johnson

and Long 1998) or fractionating (Tessier et al. 1979;

Tessier et al. 1984; Hansen et al. 1996) the test medium either prior to testing or after.

Biomarkers and Indicators of Organism Health

Biomarkers include biochemical, physiological, morphological, and histopathological responses of organisms signifying chemical exposure (Melancon 1995).

Biomarkers can sometimes be used to estimate or predict chemical exposure, effects, or both at higher levels of organization, thereby enabling evaluation of

cause-effect linkages between environmental exposure and biological responses (Kloepper-Sams et al.

1994). At the molecular level, biomarkers can provide early warning of potential higher level effects

that may not be obtainable through chemical analyses

or other methods of investigation. Some indicators

are especially useful for documenting exposure to and

the effects of ephemeral toxins as well as those that

do not accumulate in environmental media. Although

some are specific, many biomarkers are quite general.

Responses can be unique to one contaminant or a relatively small group of structurally similar chemicals,

or they may be general indicators of organism, population, community, or ecosystem health that respond

to a wide variety of chemical and other stressors. 

Biomarkers are attractive alternatives to

more traditional measures because they integrate the

cumulative effects of adverse conditions at the suborganismal level to indicate whether any precede

whole-organism or population-level effects (Adams

1990). Specific biomarkers are useful diagnostic

tools for assessing cause-effect relationships. In addition, because the toxicological response to a chemical

is caused by the interaction between the toxicant and

a cellular or extracellular component (e.g. protein,

nucleic acid), these responses may occur before any

are observed at higher levels of organization. As

noted by Mayer et al. (1992), the most useful biomarkers are those that are closely related to individual

fitness (survival, growth and reproduction) because

they represent potential surrogates for populationlevel indicators.

Population and Community Health Indicators

Although not being evaluated in the ongoing projects

(Schmitt et al. 1995; Bartish et al. 1997), populationand community-level responses and indicators constitute an important part of the BEST program. Indices

of population and community structure and function

reflect differences in the relative abundance, size, biomass, or age of organisms at different points in space

and time, and can integrate the effects of multiple

stressors across individuals and species. Structural

metrics have long been recognized as being among

the most sensitive indicators of environmental quality,

and have a rich history in the assessment and regulation of water quality (Washington 1984). Because

most population- and community-level indicators are

non-specific and may consequently yield little information regarding the nature of the stress, they are

often used together with other, more specific indicators such as environmental chemistry and toxicity

tests. Some community-level metrics, such as the

Ephemeroptera-Plecoptera-Trichoptera index (Plafkin

et al. 1989), Hilsenhoff’s (Hilsenhoff 1987) and

Lenat’s (Lenat 1993) benthic macroinvertebrate biotic

indices, and the Index of Biotic Integrity (Karr 1981;

Karr et al. 1986; Karr et al. 1987; Bramblett and

Fausch 1991) for fishes, reflect the differential susceptibilities of certain taxa to contaminants and other

stressors. At the population level, the documentation

of growth, reproduction, and survival rates of selected

fishes, birds, and mammals can be important, albeit

general, indicators of contaminant exposure.

Population- and community-level indicators, indices

and metrics are based on the relative abundance and

size of organisms in taxon and age groups. As such,

they largely reflect injury that has already occurred,

but by integrating the effects of a wide variety of contaminants and stressors they are important for environmental monitoring.

ACKNOWLEDGEMENTS

The editors would like to thank T. Bartish, S. Finger,

S. Jones, C. Kendall, and J. Lazorchak for their careful reviews of all or a part of this document. We

would also like to thank R. Lipkin for document layout and photo editing, A. Donahue and R. Young for

the formatting of tables and figures, B. Wissel for

software support, S. Rhyne for final edits, and C.

Densmore for reviewing this document and helping to

develop the gonadal staging criteria. Additionally, D.

Tillitt and J. Whyte would like to thank S. Birke, D.

Nicks, and J. Zajicek of the Biochemistry and

Physiology Branch of the CERC for their contributions to the continued development, refinement, and

improvement of the EROD and H4IIE methods. 

METHODS FOR MONITORING CONTAMINANTS AND THEIR EFFECTS 55
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