Unclassified ENV/JM/MONO(2006)22

Organisation de Coopération et de Développement Economiques 

Organisation for Economic Co-operation and Development 31-Jul-2006 

___________________________________________________________________________________________

English - Or. English 

ENVIRONMENT DIRECTORATE 

JOINT MEETING OF THE CHEMICALS COMMITTEE AND 

THE WORKING PARTY ON CHEMICALS, PESTICIDES AND BIOTECHNOLOGY 

 

OECD SERIES ON TESTING AND ASSESSMENT 

Number 55 

DETAILED REVIEW PAPER ON AQUATIC ARTHROPODS IN LIFE CYCLE TOXICITY TESTS 

WITH AN EMPHASIS ON DEVELOPMENTAL, REPRODUCTIVE AND ENDOCRINE DISRUPTIVE 

EFFECTS 

JT03212400 

Document complet disponible sur OLIS dans son format d'origine 

Complete document available on OLIS in its original format 

Unclassified ENV/JM/MONO(2006)22 English - Or. English 
ENV/JM/MONO(2006)22 

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 ENV/JM/MONO(2006)22 

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OECD Environment Health and Safety Publications 

Series on Testing and Assessment 

No. 55 

DETAILED REVIEW PAPER ON 

AQUATIC ARTHROPODS IN LIFE CYCLE 

TOXICITY TESTS WITH AN EMPHASIS ON DEVELOPMENTAL, 

REPRODUCTIVE AND ENDOCRINE DISRUPTIVE EFFECTS 

Environment Directorate 

ORGANISATION FOR ECONOMIC CO-OPERATION AND DEVELOPMENT 

Paris 

July 2006 
ENV/JM/MONO(2006)22 

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Also published in the Series on Testing and Assessment: 

No. 1, Guidance Document for the Development of OECD 

Guidelines for Testing of Chemicals (1993; reformatted 1995, 

revised 2006)

No. 2, Detailed Review Paper on Biodegradability Testing 

(1995)

No. 3, Guidance Document for Aquatic Effects Assessment 

(1995) 

No. 4, Report of the OECD Workshop on Environmental 

Hazard/Risk Assessment (1995)

No. 5, Report of the SETAC/OECD Workshop on Avian 

Toxicity Testing (1996)

No. 6, Report of the Final Ring-test of the Daphnia magna 

Reproduction Test (1997)

No. 7, Guidance Document on Direct Phototransformation of 

Chemicals in Water (1997) 

No. 8, Report of the OECD Workshop on Sharing Information 

about New Industrial Chemicals Assessment (1997)

No. 9, Guidance Document for the Conduct of Studies of 

Occupational Exposure to Pesticides during Agricultural 

Application (1997)

No. 10, Report of the OECD Workshop on Statistical Analysis of 

Aquatic Toxicity Data (1998)

No. 11, Detailed Review Paper on Aquatic Testing Methods for 

Pesticides and industrial Chemicals (1998)

No. 12, Detailed Review Document on Classification Systems 

for Germ Cell Mutagenicity in OECD Member Countries (1998)

No. 13, Detailed Review Document on Classification Systems 

for Sensitising Substances in OECD Member Countries 1998)

No. 14, Detailed Review Document on Classification Systems 

for Eye Irritation/Corrosion in OECD Member Countries (1998)

No. 15, Detailed Review Document on Classification Systems 

for Reproductive Toxicity in OECD Member Countries (1998)

No. 16, Detailed Review Document on Classification Systems 

for Skin Irritation/Corrosion in OECD Member Countries (1998)
 ENV/JM/MONO(2006)22 

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No. 17, Environmental Exposure Assessment Strategies for 

Existing Industrial Chemicals in OECD Member Countries (1999)

No. 18, Report of the OECD Workshop on Improving the Use of 

Monitoring Data in the Exposure Assessment of Industrial 

Chemicals (2000)

No. 19, Guidance Document on the Recognition, Assessment 

and Use of Clinical Signs as Humane Endpoints for Experimental 

Animals used in Safety Evaluation (1999)

No. 20, Revised Draft Guidance Document for Neurotoxicity 

Testing (in approval)

No. 21, Detailed Review Paper: Appraisal of Test Methods for 

Sex Hormone Disrupting Chemicals (2000)

No. 22, Guidance Document for the Performance of Out-door 

Monolith Lysimeter Studies (2000)

No. 23, Guidance Document on Aquatic Toxicity Testing of 

Difficult Substances and Mixtures (2000)

No. 24, Guidance Document on Acute Oral Toxicity Testing 

(2001)

No. 25, Detailed Review Document on Hazard Classification 

Systems for Specifics Target Organ Systemic Toxicity Repeated 

Exposure in OECD Member Countries (2001)

No. 26, Revised Analysis of Responses Received from Member 

Countries to the Questionnaire on Regulatory Acute Toxicity Data 

Needs (2001)

No 27, Guidance Document on the Use of the Harmonised 

System for the Classification of Chemicals Which are Hazardous 

for the Aquatic Environment (2001)

No 28, Guidance Document for the Conduct of Skin Absorption 

Studies (2004)

No 29, Guidance Document on Transformation/Dissolution of 

Metals and Metal Compounds in Aqueous Media (2001)

No 30, Detailed Review Document on Hazard Classification 

Systems for Mixtures (2001)

No 31, Detailed Review Paper on Non-Genotoxic Carcinogens 

Detection: The Performance of In-Vitro Cell Transformation 

Assays (draft)
ENV/JM/MONO(2006)22 

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No. 32, Guidance Notes for Analysis and Evaluation of RepeatDose Toxicity Studies (2000)

No. 33, Harmonised Integrated Classification System for Human 

Health and Environmental Hazards of Chemical Substances and 

Mixtures (2001)

No. 34, Guidance Document on the Development, Validation 

and Regulatory Acceptance of New and Updated Internationally 

Acceptable Test Methods in Hazard Assessment (in preparation)

No. 35, Guidance notes for analysis and evaluation of chronic 

toxicity and carcinogenicity studies (2002)

No. 36, Report of the OECD/UNEP Workshop on the use of 

Multimedia Models for estimating overall Environmental 

Persistence and long range Transport in the context of 

PBTS/POPS Assessment (2002)

No. 37, Detailed Review Document on Classification Systems 

for Substances Which Pose an Aspiration Hazard (2002)

No. 38, Detailed Background Review of the Uterotrophic Assay 

Summary of the Available Literature in Support of the Project of 

the OECD Task Force on Endocrine Disrupters Testing and 

Assessment (EDTA) to Standardise and Validate the Uterotrophic 

Assay (2003)

No. 39, Guidance Document on Acute Inhalation Toxicity 

Testing (in preparation)

No. 40, Detailed Review Document on Classification in OECD 

Member Countries of Substances and Mixtures Which Cause 

Respiratory Tract Irritation and Corrosion (2003)

No. 41, Detailed Review Document on Classification in OECD 

Member Countries of Substances and Mixtures which in Contact 

with Water Release Toxic Gases (2003)

No. 42, Guidance Document on Reporting Summary Information 

on Environmental, Occupational and Consumer Exposure (2003)

No. 43, Draft Guidance Document on Reproductive Toxicity 

Testing and Assessment (in preparation)

No. 44, Description of Selected Key Generic Terms Used in 

Chemical Hazard/Risk Assessment (2003) 

No. 45, Guidance Document on the Use of Multimedia Models 

for Estimating Overall Environmental Persistence and Long-range 

Transport (2004)
 ENV/JM/MONO(2006)22 

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No. 46, Detailed Review Paper on Amphibian Metamorphosis 

Assay for the Detection of Thyroid Active Substances (2004) 

No. 47, Detailed Review Paper on Fish Screening Assays for the 

Detection of Endocrine Active Substances (2004) 

No. 48, New Chemical Assessment Comparisons and 

Implications for Work sharing (2004) 

No. 49, Report from the Expert Group on (Q)SARs on the 

Principles for the Validation of (Q)SARs (2004) 

No. 50, Report of the OECD/IPCS Workshop on 

Toxicogenomics (2005) 

No. 51, Approaches to Exposure Assessment in OECD Member 

Countries: Report from the Policy Dialogue on Exposure 

Assessment in June 2005 (2006) 

No. 52, Comparison of emission estimation methods used in 

Pollutant Release and Transfer Registers (PRTRs) and Emission 

Scenario Documents (ESDs): Case study of pulp and paper and 

textile sectors (2006) 

No. 53, Guidance Document on Simulated Freshwater Lentic 

Field Tests (Outdoor Microcosms and Mesocosms) (2006) 

No. 54, Current Approaches in the Statistical Analysis of 

Ecotoxicity Data: A Guidance to Application (2006) 

No. 55, Detailed Review Paper on Aquatic Arthropods in Life 

Cycle Toxicity Tests with an emphasis on Developmental, 

Reproductive and Endocrine Disruptive Effects (2006) 

No. 56, Guidance Document on the Breakdown of Organic 

Matter in Litter Bags (2006) 

No. 57, Detailed Review Paper on Thyroid Hormone Disruption 

Assays (2006) 

© OECD 2006 

Applications for permission to reproduce or translate all or part of 

this material should be made to: Head of Publications Service, 

OECD, 2 rue André-Pascal, 75775 Paris Cedex 16, France 
ENV/JM/MONO(2006)22 

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ABOUT THE OECD 

The Organisation for Economic Co-operation and Development (OECD) is an intergovernmental 

organisation in which representatives of 30 industrialised countries in North America, Europe and the Asia 

and Pacific region, as well as the European Commission, meet to co-ordinate and harmonise policies, 

discuss issues of mutual concern, and work together to respond to international problems. Most of the 

OECD’s work is carried out by more than 200 specialised committees and working groups composed of 

member country delegates. Observers from several countries with special status at the OECD, and from 

interested international organisations, attend many of the OECD’s workshops and other meetings. 

Committees and working groups are served by the OECD Secretariat, located in Paris, France, which is 

organised into directorates and divisions. 

The Environment, Health and Safety Division publishes free-of-charge documents in ten different series:

Testing and Assessment; Good Laboratory Practice and Compliance Monitoring; Pesticides and 

Biocides; Risk Management; Harmonisation of Regulatory Oversight in Biotechnology; Safety of 

Novel Foods and Feeds; Chemical Accidents; Pollutant Release and Transfer Registers; Emission 

Scenario Documents; and the Safety of Manufactured Nanomaterials. More information about the 

Environment, Health and Safety Programme and EHS publications is available on the OECD’s World 

Wide Web site (http://www.oecd.org/ehs/). 

This publication was produced within the framework of the Inter-Organisation Programme for the 

Sound Management of Chemicals (IOMC). 

The Inter-Organisation Programme for the Sound Management of Chemicals (IOMC) was 

established in 1995 following recommendations made by the 1992 UN Conference on 

Environment and Development to strengthen co-operation and increase international coordination in the field of chemical safety. The participating organisations are FAO, ILO, 

OECD, UNEP, UNIDO, UNITAR and WHO. The World Bank and UNDP are observers. The 

purpose of the IOMC is to promote co-ordination of the policies and activities pursued by the 

Participating Organisations, jointly or separately, to achieve the sound management of 

chemicals in relation to human health and the environment. 
 ENV/JM/MONO(2006)22 

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This publication is available electronically, at no charge. 

For this and many other Environment, 

Health and Safety publications, consult the OECD’s 

World Wide Web site (www.oecd.org/ehs/) 

or contact: 

OECD Environment Directorate, 

Environment, Health and Safety Division

2 rue André-Pascal 

75775 Paris Cedex 16 

France 

Fax: (33-1) 44 30 61 80 

E-mail: ehscont@oecd.org 
ENV/JM/MONO(2006)22 

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FOREWORD 

This document presents a Detailed Review Paper (DRP) on Aquatic Arthropods in Life-Cycle 

Toxicity Tests with an Emphasis on Developmental, Reproductive and Endocrine Disrupting Effects. The 

initial document was prepared by the United States in 2002, to support the project on the development of a 

mysid two-generation test. Two commenting rounds were organized in 2002 and then in 2005. The scope 

of the document was extended to include more broadly aquatic arthropods species used in chronic tests 

including developmental and reproductive endpoints. Sweden agreed to take the lead for the finalization of 

the current document. 

Comments submitted by national experts and BIAC and responses to comments were presented at the 

17th Meeting of the Working Group of the National Coordinators of the Test Guidelines Programme 

(WNT) in May 2005. The document was then finalized and approved at the 18th meeting of the WNT. 

This document is published on the responsibility of the Joint Meeting of the Chemicals Committee 

and Working Party on Chemicals, Pesticides and Biotechnology.
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TABLE OF CONTENTS 

ABOUT THE OECD...................................................................................................................................... 8

FOREWORD................................................................................................................................................ 10

PREAMBLE ................................................................................................................................................. 14

ACKNOWLEDGMENTS ............................................................................................................................ 16

1.0 EXECUTIVE SUMMARY................................................................................................................ 17

2.0 INTRODUCTION ............................................................................................................................. 18

2.1 PURPOSE OF AQUATIC ARTHROPODS LIFE-CYCLE AND TWO-GENERATION 

TOXICITY TESTS ................................................................................................................................... 18

2.2 ACRONYMS AND ABBREVIATIONS....................................................................................... 19

3.0 AQUATIC INVERTEBRATE ENDOCRINOLOGY AND ENDOCRINE DISRUPTION............. 21

3.1 OVERVIEW OF AQUATIC INVERTEBRATE ENDOCRINOLOGY ....................................... 21

3.1.1 Porifera...................................................................................................................................... 24

3.1.2 Cnidaria..................................................................................................................................... 24

3.1.3 Annelida.................................................................................................................................... 25

3.1.4 Mollusca.................................................................................................................................... 25

3.1.5 Insecta ....................................................................................................................................... 25

3.1.6 Crustacea................................................................................................................................... 26

3.1.7 Echinoderms ............................................................................................................................. 28

3.1.8 Summary................................................................................................................................... 29

3.2 OVERVIEW OF POTENTIAL ENDOCRINE DISRUPTION IN AQUATIC INVERTEBRATES29

3.2.1. Endocrine Disruption in Freshwater Species........................................................................ 29

3.2.2. Endocrine Disruption in Estuarine and Marine Species ....................................................... 32

3.2.3 Summary................................................................................................................................... 36

3.3 EXTRAPOLATION ISSUES ........................................................................................................ 37

3.3.1 Extrapolation from Taxon to Taxon ......................................................................................... 37

3.3.2 Extrapolation from Individuals to Populations ......................................................................... 37

4.0 AQUATIC ARTHROPODS IN THE EVALUATION OF POSSIBLE ENDOCRINE DISRUPTION40

4.1 INSECTA ....................................................................................................................................... 40

4.1.1 Midge Larvae (Chironomus tentans [or C. dilutus] and C. riparius) ........................................ 41

4.2 CRUSTACEA ................................................................................................................................ 42

4.2.1 Amphipods................................................................................................................................ 42

4.2.2 Daphnids ................................................................................................................................... 45

4.2.3 Copepods .................................................................................................................................. 46

4.2.4 Mysids....................................................................................................................................... 46

4.2.5 Decapods (shrimp, crabs, crayfish, lobsters) ............................................................................ 52

4.3 NON-ARTHROPOD SPECIES ..................................................................................................... 53

5.0 SELECTION OF AN APPROPRIATE TEST SPECIES AND PROTOCOLS ................................ 55

5.1 CRUSTACEANS AS REPRESENTATIVE AQUATIC INVERTEBRATES ............................. 55

5.1.1 Mysids as representative crustaceans........................................................................................ 56

5.1.2 Daphnia as representative crustaceans...................................................................................... 58
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5.1.3 Harpacticoid and calanoid copepods as representative crustaceans.......................................... 59

5.2 INSECTS AS REPRESENTATIVE AQUATIC INVERTEBRATES .......................................... 61

5.2.1 Chironomids as representative insects ...................................................................................... 61

5.2.2 Protocols using other sediment-dwelling species ................................................................. 62

6.0 DESCRIPTION OF ASSAY ENDPOINTS REFLECTIVE OF REPRODUCTIVE AND 

DEVELOPMENTAL IMPAIRMENT ......................................................................................................... 75

6.1 GROWTH, MORPHOLOGICAL AND BEHAVIORAL ALTERATIONS................................. 75

6.1.1 Growth ...................................................................................................................................... 75

6.1.2 Morphology .............................................................................................................................. 77

6.1.3 Behavior.................................................................................................................................... 78

6.2 MEASURES OF REPRODUCTIVE PERFORMANCE............................................................... 79

6.2.1 Sexual Maturity......................................................................................................................... 79

6.2.2 Time to First Brood Release ..................................................................................................... 80

6.2.3 Egg Development Time ............................................................................................................ 80

6.2.4 Brood Size (Fecundity)............................................................................................................. 81

6.2.5 Intersexuality and Sex Determination....................................................................................... 81

7.0 DESCRIPTION OF POTENTIAL MODES OF ACTION REFLECTIVE OF REPRODUCTIVE 

AND DEVELOPMENTAL IMPAIRMENT................................................................................................ 83

7.1 RESPONSE TO ECDYSTEROID AGONISTS AND ANTAGONISTS...................................... 83

7.2 RESPONSE TO VERTEBRATE-TYPE ANDROGENS.............................................................. 85

7.2.1 Endpoint Sensitivity.................................................................................................................. 86

7.2.2 Gender Differences ................................................................................................................... 87

7.3 RESPONSE TO VERTEBRATE-TYPE ESTROGENS ............................................................... 88

7.4 RESPONSE TO OTHER HORMONAL DISTURBANCES ........................................................ 89

7.5 HOW TO CONFIRM MODES OF ACTION – THE USE OF BIOCHEMICAL MEASURES... 91

7.5.1 Metabolic Disruption (O:N ratios)............................................................................................ 91

7.5.2 Steroid Metabolism................................................................................................................... 92

7.5.3 Vitellogenin .............................................................................................................................. 93

7.5.4 Hormone levels......................................................................................................................... 93

7.5.5 RNA:DNA ratios ...................................................................................................................... 94

7.5.6 Protein levels............................................................................................................................. 94

7.5.7 Cytochrome P450 Enzymes...................................................................................................... 95

7.5.8 Glucose Levels.......................................................................................................................... 95

8.0 RESEARCH NEEDS AND OTHER CONSIDERATIONS ............................................................. 96

8.1 RESEARCH NEEDS ..................................................................................................................... 96

8.2 REGULATORY CONSIDERATIONS ......................................................................................... 97

9.0 REFERENCES .................................................................................................................................. 99

ANNEXES.................................................................................................................................................. 125

ANNEX 1 – Mysid life cycle draft protocol (available from the OECD upon request, please contact 

env.tgcontact@oecd.org) ........................................................................................................................ 125

ANNEX 2 – Daphnia life cycle draft protocol (available from the OECD upon request, please contact 

env.tgcontact@oecd.org) ........................................................................................................................ 125

ANNEX 3 – Copepod life cycle draft protocol (available from the OECD upon request, please contact 

env.tgcontact@oecd.org) ........................................................................................................................ 125

ANNEX 4 – Chironomid life cycle draft protocol (available from the OECD upon request, please 

contact env.tgcontact@oecd.org)............................................................................................................ 125
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Tables 

Table 1. Acronyms and abbreviations................................................................................................. 19

Table 2. Examples of hormone reported in Invertebrate Taxa............................................................ 23

Table 3. Recommended mysid life cycle toxicity test conditions....................................................... 63

Table 4. Example OECD protocoles for evaluating chronic toxicity in aquatic arthropods............... 66

Table 5. Example US EPA protocols for evaluating chronic toxicity in aquatic arthropods.............. 71
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PREAMBLE 

In 1998, a Task Force on Endocrine Disrupter Testing and Assessment (EDTA) was established at the 

request of OECD member countries. The EDTA Task Force is a Special Activity of the Test Guidelines 

Programme and its main objectives are to: 

• identify the needs and prioritize the development of new and enhanced guidelines for the 

detection and characterization of endocrine disrupting chemicals; 

• develop a harmonized testing strategy for the screening and testing of endocrine disrupters; 

• manage validation work for newly developed and enhanced Test Guidelines as appropriate; and, 

• provide practical tools for sharing testing results and assessments. 

The need for new and updated test methods to detect and characterize endocrine disrupting chemicals 

has been expressed by the Task Force for the assessment of human health effects and environmental 

effects. At early meetings of the EDTA Task Force, it appeared that existing OECD Test Guidelines would 

insufficiently cover for endocrine-related effects, especially for the environment. Member countries 

decided to list test methods which could potentially cover effects of chemicals on the reproductive system 

(estrogen agonists/antagonists and androgen agonists/antagonists) and on the development (thyroid 

system), and proposed enhancements where needed. Invertebrate life-cycle tests, encompassing 

development and reproduction, are proposed as in vivo assays including endpoints that indicate 

mechanisms of adverse effects (endocrine and other mechanisms) and potential population damage. 

This Detailed Review Paper (DRP) is intended to provide the current state-of-the-knowledge in the 

area of aquatic invertebrate endocrinology, with a particular focus on the evaluation of possible 

developmental reproductive and endocrine disruptive effects in aquatic arthropods through life cycle tests. 

The content of this review is intended to describe test methods, already existing or in development, 

that could be used in testing chemicals for their potential to interact with the endocrine system of aquatic 

arthropods and likely to create population damage. However, the methods described do not focus on 

endocrine disruption in terms of endpoints measured, but rather on reproductive and developmental 

performance of individuals exposed. It is planned that test methods under development or enhanced, will 

be validated to establish their relevance (i.e. whether the test is meaningful and useful for the intended 

purpose) and their reliability (i.e. reproducibility of test results over time within and among laboratories). 

General principles for the conduct of validation studies have been defined following the OECD 

Stockholm Conference on Validation and Regulatory Acceptance of New and Updated Test Methods in 

Hazard Assessment. The OECD Guidance Document No. 34 describes these guiding principles and 

addresses the important steps and aspects that must be considered prior to and during the validation 

process. They include: (i) the definition of the test method and related issues (e.g., purpose, decision 

criteria, endpoints, limitations); (ii) the design and conduct of pre-validation studies leading to the 

optimization of the test method; (iii) the design and conduct of the formal inter-laboratory validation work, 

based on the outcomes of the pre-validation studies and aiming at accumulation of data on the relevance 

and reliability of the test method, and; (iv) the overall data evaluation and subsequent validation study 

conclusion, keeping in mind the requirements of regulatory authorities, for submission of information 

relating to new or modified test procedures. It also discusses the need for and the extent of an independent 

evaluation, or peer review, of test methods being validated. 
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The planning and conduct of a validation study should be undertaken on a case-by-case basis since 

there may be several ways of assessing the validity of the method. As described in the Guidance Document 

No. 34, the validation process is sufficiently flexible so that it can be applied equally well to a wide variety 

of tests and procedures. The flexibility also applies regardless of whether tests are for health or 

environmental effects. Flexibility is also encouraged on issues such as the amounts of information required 

at each phase, the number of chemicals tested, when and to what extent to use blind testing, and the 

number of laboratories participating. 

A Validation Management Group for Ecotoxicity Testing (VMG-eco) has been established at the 

OECD level to supervise the planning and conduct of experimental work in fish, birds, amphibians and 

invertebrates. This VMG-eco reports back to the Task Force on Endocrine Disrupters Testing and 

Assessment (EDTA). To discuss the technical details of the test methods, an Ad hoc Expert group on 

Invertebrate Testing was created in 2002 and met for the first time in October 2003 to discuss the various 

tests proposed and to prepare initial validation work. 

The U.S. Environmental Protection Agency took the lead in preparing the initial and revised versions 

of this Detailed Review Paper for their national programme on endocrine disrupters. The draft document, 

initially called “DRP on a Mysid two-generation test”, was circulated for comments in 2002, discussed at 

the Second meeting of the Validation Management Group for Ecotoxicity tests (VMG-eco) and at the 

Seventh meeting of the Task Force on Endocrine Disrupters Testing and Assessment. These discussions 

and the comments received resulted in a recommendation to broaden the scope of the DRP to aquatic 

arthropods. The United States volunteered to revise the draft taking into account the recommendation and 

the initial comments compiled by the Secretariat in March 2003. 

In December 2004, the United States submitted a revised version of the document, called “DRP on 

Aquatic Arthropods in life cycle and two-generation toxicity tests”. This DRP was circulated in January 

2005 to members of the Invertebrate Expert Group and the VMG-eco for comments again. In the 

comments received as of 23 February 2005 experts acknowledged that aquatic arthropods are now taken 

into account in the review, but there is still a large proportion of the document dedicated to mysids. 

Additionally, some experts commented that this version of the DRP did not address all the publicly 

available literature on freshwater and marine arthropods that is relevant to endocrine disrupters. 

At the 17th Meeting of the Working Group of National Coordinators of the Test Guidelines 

Programme (WNT) it was decided that aquatic arthropods should be better represented, including a 

thorough re-evaluation of the existing literature on the subject. 

Sweden volunteered to take over the lead for the project to include contributions from experts, e.g. 

from Japan, Germany and the Netherlands, and thereby finalize the review. This was decided at the 17th

WNT meeting. It was also decided that a very pragmatic approach should be taken in order to finalize the 

review for the next WNT meeting in May 2006. The work have taken place from late 2005 to early 2006 

and responsible for the work have been experts at the Department of Applied Environmental Science 

(ITM) in Sweden, in co-operation with experts from e.g. the above mentioned countries, the US and the 

Swedish Chemicals Inspectorate (KemI). 
ENV/JM/MONO(2006)22 

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ACKNOWLEDGMENTS 

The OECD Secretariat would like to acknowledge the contributions of primary and secondary authors 

and national experts who provided extensive comments on the draft versions of this Detailed Review 

Paper: 

• Roy K. Kropp, Margaret R. Pinza, and Michael L. Blanton, as primary authors; Battelle (Unites 

States) 

• Magnus Breitholtz and Ulrika Dahl, as secondary authors responsible for revision and additional 

writing (Sweden) 

• Environmental Health Science Bureau, Health Canada (Canada) 

• Chemicals Division, Danish Environment Protection Agency (Denmark) 

• UFZ Center for Environmental Research, Chemical Ecotoxicology, Leipzig/Halle (Germany) 

• ECT Ecotoxicologie, Flörsheim (Germany), 

• Rheinisch-Westfälische Technische Hochschule, Research Group Aquatic Ecotoxicology and 

Mathematical Modelling, Aachen (Germany) 

• DIN working group on marine biotests and collaborators (Germany) 

• Federal Environment Agency, Berlin (Germany), 

• Water Research Centre and the Health and Safety Executive, United Kingdom 

• ABC Laboratories, Inc.; Chemical Development Group; Columbia (MO, United States) 

• Rohm & Haas Company; Toxicology Department; Spring House (PA, United States) 

• U.S. Environmental Protection Agency Federal Advisory Committee; Endocrine Disruptor 

Methods Validation Subcommittee; Office of Science Coordination & Policy; Washington DC 

• U.S. Environmental Protection Agency; Office of Research & Development; NHEERL Atlantic 

Ecology Division; Narragansett (RI, United States) 

• Department of Applied Environmental Science (ITM); Stockholm University, (Sweden) 

• Department of Environmental Health Sciences, Arnold School of Public Health, University of 

South Carolina, Columbia (SC, United States) 

• BASF AG, Agricultural Center, Ecotoxicology, Limburgerhof (Germany) 

• National Institute for Environmental Studies, Environmental Chemistry Division, Ecological 

Chemistry Section, Ibaraki (Japan)

• Nord-Utte (Nordic co-ordination group for the development of test methods in toxicology and 

ecotoxicology), Nordic Chemicals Group, Nordic Council of Ministers

• Swedish Chemicals Inspectorate (KemI), Stockholm (Sweden)
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1.0 EXECUTIVE SUMMARY 

i) The overall purpose of this Detailed Review Paper (DRP) is to provide state-of-the-art 

knowledge about (life-cycle) tests on aquatic arthropods. A special emphasis is put on developmental and 

reproductive toxicity as well as potential endocrine disruptive effects. To understand both the possibilities 

and the difficulties in identifying potential endocrine disrupting chemicals within future regulatory work, 

this DRP presents information on basic invertebrate endocrinology and gives an overview of endocrine 

disruption observed in these groups of species. In this context it is important to emphasize that endocrine 

disruptive- and other specific biological effects (e.g. enzymatic) related to chemical exposure will be 

covered by studying endpoints reflective of both developmental and reproductive impairment (i.e. full lifecycle studies), regardless if we know the specific mode-of-action or not. This is the strength with the lifecycle protocols presented in the following text. In addition to this, the DRP also presents a number of 

established and novel biochemical techniques and endpoints, which should be used in concert with the 

regular endpoints observed from the life-cycle testing in order to increase our understanding of modes-ofaction. 

ii) Invertebrates (especially arthropods such as insects and crustaceans) constitute the vast majority 

of animal species on earth. Although many invertebrate toxicity test protocols are routinely used in 

regulatory toxicity testing, few have been designed with endocrine-specific endpoints in mind. Although 

the growth, reproduction, development, and other aspects of invertebrate physiology and life cycle are 

known to be regulated by endocrine control, the endocrine systems and the hormones produced and used in 

the invertebrate body are not directly analogous to those of vertebrates. For example, ecdysone is a steroid 

hormone that regulates growth and molting in arthropods, and exhibits some functional and structural 

similarities to estrogen. Therefore, methods for testing effects of potential endocrine disrupters on aquatic 

arthropods are relevant to assess the adverse consequences of such chemicals on development and 

reproduction, through life-cycle or two-generation tests. 

iii) This DRP reviews the endocrinology of aquatic invertebrates in general (annelids, mollusks, 

insects, crustaceans and echinoderms, mainly) and provides examples of endocrine disruption. Issues 

associated with extrapolation are also explained in Section 3. 

iv) The remaining sections of the paper focus on aquatic arthropods. Section 4 lists those species 

that have been studied most and where test methods have sometimes been developed to assess reproductive 

and developmental impairments following exposure to chemical substances. This is followed by an 

evaluation of possible endocrine disruption in aquatic arthropods. 

v) A number of appropriate test species and standard protocols are presented in Section 5. 

vi) In Section 6 a description of endpoints reflective of developmental and reproductive impairment 

is presented. 

vii) Responses observed following exposure to ecdysteroid (ant-)agonists, androgenic and estrogenic 

compounds and other hormonal disturbances are presented in Section 7. 

viii) Research needs and implementation considerations are described in Section 8; followed by 

references in Section 9. 

ix) In Annex 1 to 4, established or proposed OECD arthropod life cycle protocols are presented. 
ENV/JM/MONO(2006)22 

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2.0 INTRODUCTION 

2.1 PURPOSE OF AQUATIC ARTHROPODS LIFE-CYCLE AND TWO-GENERATION 

TOXICITY TESTS 

1. Test methods proposed at Level 5 of the EDTA Conceptual Framework are meant i) to provide 

data on effects from endocrine and other mechanisms of action, and ii) to be used for assessment of 

adverse effects and potential population damage. To fulfill this purpose, tests are often longer-term studies 

designed to encompass critical life states and processes, a broad range of doses, and administration by 

relevant route of exposure. In addition, the effects associated with potential endocrine disrupters can be 

latent and not manifested until later in life or may not be apparent until reproductive processes occur in an 

organism’s life history. Thus, tests for endocrine disruption often encompass two generations to address 

effects on fertility and mating, embryonic development, sensitive neonatal growth and development, and 

transformation from the juvenile life state to sexual maturity. 

2. The results from such long-term tests should be conclusive in documenting a discernable causeand-effect relationship of chemical exposure to measurable manifestation in the test organisms. The tests 

are generally expected to: 

• Establish exposure/concentrations/timing and effects relationships; 

• Be sensitive; 

• Assess relevant endpoints, i.e. indicative of developmental and reproductive impairment; 

• Include a dose range for full characterization of effects; 

• Adhere to good laboratory practices; and 

• Be suitable for validation. 

3. Invertebrates (especially arthropods such as insects and crustaceans) constitute the vast majority 

of animal species on earth. Although many invertebrate toxicity test protocols are routinely used in 

regulatory toxicity testing, few have been designed with endocrine-specific endpoints in mind. Although 

the growth, reproduction, development, and other aspects of invertebrate physiology and life cycle are 

known to be regulated by endocrine control, the endocrine systems and the hormones produced and used in 

the invertebrate body are not directly analogous to those of vertebrates. For example, ecdysone is a steroid 

hormone that regulates growth and molting in arthropods, and exhibits some functional and structural 

similarities to estrogen. It has been reported that the vertebrate androgen testosterone acts as an 

ecdysteroid antagonist in a crustacean (Mu and LeBlanc 2002a). Also, the aromatase inhibitor fenarimol, 

which prevents the conversion of testosterone to the vertebrate estrogen, has been demonstrated to inhibit 

ecdysteroid synthesis and interfere with normal molting processes in a crustacean (Mu and LeBlanc 

2002b). However, other substances than vertebrate estrogens and other steroid hormones and mimics are 

likely to be more prone to interfere with the unique invertebrate hormone systems. Therefore, methods for 

testing arthropods for effects of potential endocrine disrupters are relevant to assess the adverse 

consequences of chemicals indicated to be endocrine active in screening and short-term assays and tests. 
 ENV/JM/MONO(2006)22 

 19

2.2 ACRONYMS AND ABBREVIATIONS 

4. Table 1 lists the acronyms and abbreviations used in the detailed review paper (DRP), with the 

exception of commonly used units, such as h for hour or L for liter. Each of the acronyms and 

abbreviations is also introduced at first use in the text. 

Table 1. Acronyms and abbreviations 

20-E 20-hydroxyecdysone 

4NP 4-nonylphenol 

AFDW Ash-free dry weight 

ASTM American Society for Testing and Materials 

BpA Bisphenol A 

BPDH Black pigment-dispersing hormone 

CHH Crustacean hyperglycemic hormone 

CNS Central nervous system 

CYP Cytochrome P450 enzyme 

DAH Dark adapting hormone 

DDT Dichlorodiphenyl trichloroethane 

DES Diethylstilbestrol 

DRP Detailed Review Paper 

EC Effect concentration 

EC50 Median effective concentration 

EDC Endocrine-disrupting chemical 

EDSTAC US EPA Endocrine Disrupters Screening and Testing Advisory Committee 

EDSP Endocrine Disrupters Screening Programme 

EPA Environmental Protection Agency 

GIH Gonad-inhibiting hormone 

GSH Gonad-stimulating hormone 

GSS Gonad-stimulating substance 

HPV-inerts High production volume inert compounds 

IC Inhibition concentration 

IGR Insect growth regulator 

JH Juvenile hormone 

LAH Light-adapting hormone 

LC Lethal concentration 

LC50 Median lethal concentration 

LOEC Lowest observed effects concentration 

MAR Metabolic androgenization ratio 

MF Methyl farnesoate 
ENV/JM/MONO(2006)22 

 20

MIH Molt-inhibiting hormone 

MSD Minimum significant difference 

NIEHS National Institute of Environmental Health Sciences 

NOEC No observed effect concentration 

NPPE Nonylphenol polyethoxylate 

OECD Organization for Economic Cooperation and Development 

OPPTS US EPA Office of Prevention, Pesticides and Toxic Substances 

PCB Polychlorinated biphenyl 

PoA Ponasterone A 

SAB Scientific Advisory Board 

SAP Scientific Advisory Panel 

SDWA Safe Drinking Water Act 

TBT Tributyltin 
 ENV/JM/MONO(2006)22 

 21

3.0 AQUATIC INVERTEBRATE ENDOCRINOLOGY AND ENDOCRINE DISRUPTION 

5. Invertebrates comprise 95% of the world’s animal species (Wilson 1988), and certainly a larger 

percentage of the Earth’s total animal abundance. Certainly, it is understood that hormones are important 

in controlling physiological processes in invertebrates (Lafont 2000, Oehlmann and Schulte-Oehlmann 

2003). However, until recently invertebrates received little attention regarding potential endocrine 

disruption, the main exception being studies of imposex in gastropods. Invertebrate endocrine systems, 

except perhaps those of arthropods, are generally poorly studied compared to those of vertebrates and this 

limits ability to evaluate potential EDCs (Lafont 2000, Oehlmann and Schulte-Oehlmann 2003). Because 

many invertebrate endocrine systems are not very well understood, some changes to endocrine systems 

after exposure to a chemical may not be detected or may not be measurable (Oehlmann and SchulteOehlmann 2003). However, the situation is improving with many studies being done within the last few 

years on such groups as sponges, corals, polychaetes, and echinoderms (Section 3.1), adding to what has 

been learned about arthropod endocrinology. Nonetheless, to gain a full understanding of the complex 

endocrine disruption issue, invertebrates must be included in a tiered testing approach (Vandenbergh et al. 

2003). 

3.1 OVERVIEW OF AQUATIC INVERTEBRATE ENDOCRINOLOGY 

6. There are several recent general reviews that provide good summaries of invertebrate 

endocrinology (LeBlanc et al. 1999, Lafont 2000a, Feix and Hoch 2002). Additionally, there are some 

reviews that pertain to specific groups of invertebrates including Cnidaria (Leitz 2001), Annelida (Hardege 

1999, Andries 2001, Salzet 2001), Insecta (e.g., Lafont 2000b), and Crustacea (Fingerman 1987, 1997, 

Huberman 2000, Subramoniam 2000). The intent of this section is to provide a brief overview of 

invertebrate endocrinology. More details can be found in the review articles. 

7. As animal body plans increased in complexity from simple cell-based to organ system-based 

structures, the need for a means to coordinate internal processes became more important. Neural networks 

are developed in the Cnidaria, one of the more primitive invertebrate groups (Lafont 2000a). The 

complexity of the coordination systems increased through the invertebrate evolutionary line, progressing to 

the presence of endocrine cells in the Annelida, endocrine glands in the Mollusca, and perhaps culminating 

in complexity in the Arthropoda, although Lafont (2000a) offered that this conclusion might reflect the 

general lack of study of invertebrate endocrine systems. The complexity of coordinating systems is 

reflected in the pathways by which stimulation of the central nervous system (CNS) generates a response 

in a target organ (Lafont 2000a). The stimulus can induce the CNS to produce a neurotransmitter or a 

neurohormone. Both messengers can then act directly on the target organ or on an endocrine gland. The 

endocrine gland can then act on the target organ or on another endocrine gland, which then acts on the 

target organ. 

8. There are two main types of chemical messengers (Lafont 2000a): 

• Neurotransmitters may be fast- or slow-acting depending on whether the link to ion channels is 

direct or indirect (acetylcholine, glutamate); these are low molecular weight molecules and have 

low diversity throughout invertebrates, i.e., they are evolutionarily conservative. 

• Neurohormones/hormones include peptide/protein and lipid molecules that can be grouped by 

similarity of structure (e.g., steroids, peptides, terpenoids); these messengers show important 

evolutionarily diversification and either are related to vertebrate messengers or are specific to 

invertebrates. 
ENV/JM/MONO(2006)22 

 22

9. Lafont made two key generalizations about invertebrate endocrine systems that are important to 

consider when evaluating endocrine disruption in invertebrates: (1) there are two sets of hormones in 

invertebrates—those that are similar to those found in vertebrates (i.e., they likely share a common 

ancestry) and those specific to inverts (e.g., ecdysteroids); (2) structurally related molecules may have 

different functions. This last generalization is very important and has been shown to be applicable to some 

studies of EDCs (Section 3.2) 

10. LeBlanc et al. (1999) gave an introductory overview of animal lineages as they relate to 

endocrinology. Briefly, animals diverged from the ancestral “stem” line into two primary lineages, the 

protostomes and the deuterostomes. Traditionally, embryological features have been used to separate the 

two lineages. Major stem groups that will be discussed here are the sponges (Porifera) and jellyfish, 

hydroids, and corals (all Cnidaria). Major Protostome groups include worms (Annelida), snails and clams 

(Mollusca), midges (Insecta), and amphipods, daphnids, copepods, mysids, and decapods (Crustacea). The 

primary deuterostomes mentioned here are sea stars and feather stars (Echinodermata) and fish 

(Vertebrata). This divergence into two main lineages corresponds to important endocrinological 

differences. Protostomes primarily rely on neuropeptides to regulate physiological processes, although 

more advanced groups (insects, crustaceans) have increased reliance on ecdysteroids and terpenoids. 

Invertebrate deuterostomes (echinoderms) rely more on vertebrate-like steroids (estrogens, androgens, 

progesterone) and terpenoids, but do not have ecdysteroid hormones. 

11. LeBlanc et al. (1999) also discussed the endocrinology of the major invertebrate groups. Many 

studies of individual invertebrate groups have identified single hormones or groups of hormones. 

Examples of these studies, selected primarily to build on the LeBlanc et al. (1999) review, are discussed in 

the following sections. The groups are presented in approximate phylogenetic sequence. Oehlmann and 

Schulte-Oehlmann (2003) prepared a table of the hormones that have been identified for the major 

invertebrate taxa, and the processes they control. Table 2 is a modification of their table that has been 

updated to include some additional taxa and hormones that have been identified recently. 
 ENV/JM/MONO(2006)22 

 23

Table 2. Examples of hormone reported in Invertebrate Taxa. Modified from Summary table from Oehlmann and 

Schulte-Oehlmann (2003), unless indocated otherwise 

Taxon 

Hormone 

Type Example Controlled Process 

Porifera Unknown Unknown Unknown 

Cnidaria Neuropeptides 

Thyroids 

Retinoids 

Steroids 

GLWamides 

Thyroxine 

9-cis-retinoic acid 

17β-estradiol 1

Metamorphosis 

Strobilation 

Strobilation 

Reproduction 

Nematoda Ecdysteroids 

Terpenoids 

Neuropeptides 

Unknown 

Juvenile hormone-like 

FMRFamide 

Unknown 

Growth 

Unknown 

Annelida Ecdysteroids 

Neuropeptides 

Terpenoids 

Ecdysone 

FMRFamide 

Gonadotropin 2

Eicosatrienoic acid 3 

Aracidonic acid 3 

Unknown 

Neuromodulation 

Vitellogenesis 

Metamorphosis 

Mollusca Ecdysteroids 

Steroids 

Terpenoids 

Neuropeptides 

Unknown 

Testosterone, 17β-estradiol, 

progesterone 

JH reported 

APGWamide, dorsal body 

hormone, FMRFamide, egglaying hormone, molluscan 

insulin-like peptides 

Unknown 

Sexual differentiation, 

prosobranch reproduction 

Questionable 

Sexual differentiation, gonad 

maturation, spawning, 

neuromodulation, growth, 

development, energy metabolism 

Crustacea Ecdysteroids 

Steroids 

Terpenoids 

Neuropeptides 

Ecdysone 

17β-estradiol, testosterone, 

progesterone 

Methyl farnesoate 

Androgenic hormone 

Crustacean Hyperglycemic 

Hormone 

Molt-inhibiting hormone 

Vitellogenesis-inhibiting 

hormone 

Molting, vitellogenesis 

Uncertain 

Metamorphosis 

Sexual differentiation, 

vitellogenesis inhibition 

Energy metabolism 

Ecdysteroid production 

Vitellogenesis 

Echinodermata Steroids 

Neuropeptides 

Testosterone, 17β-estradiol, 

estrone 

Gonad-stimulating substance 

Maturation-promoting factor 

Vitellogenesis, gametogenesis, 

spawning 

Spawning 

Fertilization 

Tunicata Steroids 

Neuropeptides 

Thyroids 

Testosterone. 17β-estradiol 

GRH analog 

Thyroxine 

Gametogenesis, spawning 

Gonad development 

Tunic formation (?) 1

 Tarrant et al. (1999), Pernet and Anctil (2002). 2 Gaudron and Bentley (2002). 3 Laufer and Biggers (2001). 
ENV/JM/MONO(2006)22 

 24

3.1.1 Porifera 

12. The current literature review did not identify any studies that have demonstrated the occurrence 

of hormones or endocrine functions in the sponges. However, one study (Hill et al. 2002) did examine the 

effects of three chemicals known to disrupt endocrine systems in some animals on freshwater sponges (see 

Section 3.2.1). 

3.1.2 Cnidaria 

13. The cnidarians are of particular interest because of the group’s position as one of the stem 

invertebrate phyla that existed before the divergence into the protostome and deuterostome lineages. 

Cnidarians may, therefore, show aspects of endocrinology conserved in both groups (LeBlanc et al. 1999). 

Also, some taxa within the phylum are of particular importance in tropical waters, e.g., coral reefs 

comprise a major ecosystem. Cnidarian neurosecretory cells may have been precursors to the evolution of 

the neurohormonal systems of higher animals (Feix and Hoch 2002). 

14. Leitz (2001), in a thorough review of cnidarian endocrinology, reported that cnidarians do not 

possess defined endocrine glands, but that neurones are the major source of signaling compounds in 

cnidarians, although the target cells rarely have been identified. Leitz identified several major groups of 

regulatory compounds. Non-peptide regulatory compounds include catecholamines and their precursors 

(dopamine, adrenaline, norepinephrine), serotonin (5-hydroxytryptamine), taurine, and gammaaminobutyric acid (GABA). The functions of most have not been identified clearly. Serotonin may be 

involved in metamorphosis, taurine may function in osmoregulation, and GABA may affect feeding. 

15. Most of the peptide regulatory compounds are neuropeptides and inhibit or stimulate muscle 

contraction in hydrozoans and anthozoans (Leitz 2001). Gonadotropin-releasing hormones (GnRHs) 

comprise a peptide family that is conserved in length and amino-acid sequence composition (Anctil 2000). 

There is some evidence that they may also be active in other invertebrates (e.g., mollusks and tunicates). 

Anctil (2000) found evidence for GnRHs in the sea pansy Renilla koellikeri that are physiologically active 

and function in modulation of peristalsis (GnRH inhibits peristalsis), which is strongly enhanced during 

spawning. They were also found in starlet anemone Nematostella vectensis. 

16. Retinoids occur in cnidarians. Retinoic acid X receptors (RXR) are nuclear hormone receptors 

found in vertebrates, echinoderms, arthropods, nematodes (Kostrouch et al. 1998). Kostrouch et al. found 

jellyfish RXR (jRXR), which is a close homolog of vertebrate RXR, in the jellyfish (Tripedalia 

cystophora). This hormone targets genes that encode soluble crystallins in lens of eye (as in vertebrates), 

which suggests that cnidarians are ancestral to other phyla, not an independent offshoot. 

17. Sex steroids were identified in cnidarians very recently. Pernet and Anctil (2002) studied the sea 

pansy, which is dioecious and forms separate male and female colonies. Pernet and Anctil discovered the 

vertebrate estrogen 17-β estradiol (E2) in all colonies and found that the levels varied through the 

reproductive cycle with a strong peak in March (at the onset of maturation) and June (start of the spawning 

period). These observations argue that 17-β estradiol has a role in the reproductive biology of the sea 

pansy. Tarrant et al. (1999) identified 17-β estradiol in a scleractinian coral (Montipora capitata) and 

thought that it functioned in reproduction. Corals contain a variety of steroids (Tarrant et al. 2003 among 

others), but there have been few studies on steroid metabolism. Water-born estrogens can be taken up by 

corals (Tarrant et al. 2001), metabolized (Tarrant et al. 2003), and can affect coral physiology (Tarrant et 

al. 2004). The presence of a vertebrate estrogen (17-β estradiol) in cnidarians is important for eventual 

understanding of evolution of hormonal systems. 
 ENV/JM/MONO(2006)22 

 25

3.1.3 Annelida 

18. All hormones that have been identified in annelids to date are neuropeptides secreted by 

neurosecretory cells located primarily in the head (LeBlanc et al. 1999; Salzet 2001). A cardioactive 

peptide, FMRFamide (first found in mollusks), has been found in polychaetes, but not oligochaetes; 

ecdysteroids occur in some annelids, but no function for them has been determined; juvenile hormone 

affects larval settling when studied in the laboratory. Gaudron and Bentley (2002) discovered that the 

prostomium produces a gonadotrophic hormone that controls vitellogenesis in the oocytes. The hormone 

production is under environmental control. A second hormone induces oocyte maturation. 

19. Andries (2001) reviewed the endocrine regulation of reproduction in polychaete worms. A single 

brain hormone controls reproductive development in worms that breed once and then die (semelparous 

reproduction) by inhibiting gonad maturation and promoting somatic growth. Hormone levels remain 

consistent throughout the extended period of gamete development and the population of oocytes, which 

includes various stages of development during the one- to three-year period, becomes homogeneous only 

when the hormone levels decrease. The hormone apparently also inhibits spermatogenesis. Pheromones 

are involved in the timing of broadcast spawning (Hardege 1999, Andries 2001). Reproduction in worms 

that breed annually or continuously (iteroparous reproduction) is regulated by a combination of 

environmental factors and hormones (e.g., supraesophogeal hormone stimulates ovarian-protein synthesis). 

20. Laufer and Biggers (2001) reviewed the role of methyl farnesoate and juvenile hormone-active 

fatty acids in annelid metamorphosis and reproduction. They reported the Capitella trochophores respond 

very quickly to exposure to methyl farnesoate and eicosatrienoic acid (sperm maturation factor). 

Eicosatrienoic acid is present in adults of Arenicola and was proposed as a hormone that functions in 

metamorphosis and also induces spawning. 

3.1.4 Mollusca 

21. Neurosecretory centers, which produce neuropeptides, occur in the cerebral, pleural, pedal, 

abdominal ganglia of the central nervous system and comprise the molluscan endocrine system (LeBlanc et 

al. 1999). Typical invertebrate steroids (e.g., ecdysone) have been reported only rarely to occur in 

mollusks. Vertebrate-type steroids (testosterone, progesterone) can be synthesized from precursors in the 

ovotestes, which underscores the hermaphroditic character of many mollusks. 

22. Ecdysteroids and juvenoids occur in some mollusks, but their particular functions have not been 

determined. The neuropeptide FMRFamide, a cardioaccelatory peptide that was first identified in 

mollusks, is one of the best known and most widespread neuroendocrine hormones. It regulates several 

physiological processes. Other peptides are involved in gonad maturation and egg production (egg-laying 

hormone), the development of female accessory sex organs, gonad maturation and ovulation (dorsal body 

hormones), and in growth, development, and metabolism (molluscan insulin-like peptides). 

3.1.5 Insecta 

23. Insect endocrinology, which is probably better understood than that of any other invertebrate 

group, has been thoroughly reviewed (e.g., LeBlanc et al. 1999, Lafont 2000b). Three types of structures 

comprise the endocrine system—neurosecretory cells, endocrine glands (epitracheal glands, corpora allata, 

prothoracic glands), and reproductive organs (LeBlanc et al. 1999). The neurosecretory cells produce 

neuropeptides that permit insects to respond to environmental factors that include food availability, 

temperature, and photoperiod, among others. Many hormones, most of which are unique to arthropods, 

have been identified in insects. Some insect hormones are similar to vertebrate hormones, but their 

functions have not been identified yet. 
ENV/JM/MONO(2006)22 

 26

24. Two of the key insect hormones that are very relevant to an EDC evaluation program are 

ecdysteroids and juvenile hormones. Ecdysteroids, which are compounds structurally related to ecdysone 

(Goodwin et al. 1978), comprise one of the more important groups of insect hormones because they are 

involved in growth and development (molting) and reproduction (Lafont 2000b). Ecdysteroids are 

secreted by the prothoracic gland and by endocrine cells in the gonads. Insects cannot synthesize 

cholesterol; therefore, ecdysteroids are prepared from ingested cholesterol and plant steroids (LeBlanc et 

al. 1999). The synthetic pathway leading to the production of ecdysteroids is not completely known, partly 

because unstable compounds are included in some of its early stages (Lafont 2000b). Lafont mentioned 

two of the main questions concerning ecdysteroids that have yet to be answered: (a) How many active 

ecdysteroids and ecdysteroid receptors are there? and (b) How do varying concentrations of the hormones 

influence their action? 

25. Juvenile hormones (JH) are terpenoids produced by the corpora allata that primarily regulate 

metamorphosis and reproduction. The presence or absence of JH determines the type of molt that occurs 

in the insect. The presence of JH during the initial rise of ecdysteroids in a larval stage will induce a molt 

to another larval stage (LeBlanc et al. 1999). When ecdysone levels start to rise during the final larval 

stage, the absence of JH will result in metamorphosis. In hemimetabolous species, this molt leads to the 

adult stage. In holometabolous species, JH again is present during this molt and the larva transforms into a 

pupa. The pupa molts to the adult in the absence of JH. JH biosynthetic pathways are known and are 

reviewed by Lafont (2000b). There are three main types of JH, called JHI, JHII, and JHIII, and several 

related compounds, including methyl farnesoate (MF) (Lafont 2000b). Lafont described a possible 

evolutionary scenario for JH with primitive insects (cockroaches) using MF and JHI and the most 

advanced insects (flies) using a unique form of JHIII (JHIII bisepoxide). 

26. Many of the insecticides used to control outbreaks of agricultural pests have been formulated to 

interact with either of these two hormones. Many of these compounds, especially JH analogs, are not 

species-specific in action; thus they have the potential to impact non-target animals severely. Therefore, 

these compounds are often the focus of endocrine disruption studies. 

3.1.6 Crustacea 

27. For crustaceans, biological processes are regulated by a complex endocrine system (Cuzin-Roudy 

and Saleuddin 1989). Study of this regulation began about 1921 with the publication of the results of 

experiments by R. Courrier that showed that a hormone produced by an endocrine gland, not the testes, 

determined male secondary sex characteristics (Fingerman 1997). Fingerman’s (1997) historical review 

traces the passage of crustacean endocrinology from its early development by emphasizing key discoveries 

such as the nature and role of the sinus gland and other glands, and the roles of specific hormones in 

crustacean biology. Many of the descriptions are enhanced by the author’s personal experiences gained 

through 40 years of active research on crustacean endocrine systems. Crustacean endocrinology has been 

reviewed about every decade since the early studies of the 1920s and 1930s. Among the relatively recent 

reviews, Quackenbush (1986) reviewed studies of the four types of compounds that help regulate 

crustacean biology—peptides, steroids, terpenoids, and biogenic amines. Later reviews have approached 

crustacean endocrinology by examining physiological processes, particularly growth and reproduction 

(Charmantier et al. 1997, Chang 1997, Subramoniam 2000). Chang (1993) compared endocrine control of 

molting and reproduction in crustaceans to that in insects. 

28. Basically, inputs from the environment are integrated by a central nervous system. 

Neurotransmitters and neuromodulators govern the release of neuropeptides, which govern the production 

of hormones by the endocrine glands (Cuzin-Roudy and Saleuddin 1989). Molting, for example, is 

controlled by the release of molting hormones, which are ecdysteroids, and by neurosecretions for the 

central nervous system, which are accumulated and released by the sinus gland. The main endocrine 
 ENV/JM/MONO(2006)22 

 27

centers for crustaceans described to date include the Y-organ, mandibular organ, androgenic gland, Xorgan, and sinus gland. 

29. Peptide hormones in crustaceans include compounds, such as red pigment concentrating hormone 

(RPCH) and pigment concentrating hormone, that affect the chromatophores and retinal pigments that 

comprise the complex color control systems found in crustaceans (Quackenbush 1986). Peptides also 

affect crustacean molting and reproduction. Molt-inhibiting hormone (MIH) is one factor that regulates 

molting. Peptides function in the control of vitellogenesis in crustaceans. Crustacean hyperglycemic 

hormone (CHH) is a peptide that regulates blood sugar, particularly glucose, in crustaceans (Quackenbush 

1986). The hormone is unique among crustaceans in that it is taxon-specific. For example, CHH produced 

by crayfish does not affect crabs. 

30. Steroid hormones include the ecdysteroids, which are the molting hormones in crustaceans 

(Charmantier et al. 1997, Chang 1997). These growth-regulating hormones also function in the control of 

reproduction and embryogenesis (Subramoniam 2000). Ecdysteroids are synthesized by the ecdysial 

glands or Y-organs. The Y-organ secretes ecdysone, which is converted to 20-hydroxyecdysone, the active 

ecdysteroid in most crustaceans. Several studies have shown that the Y-organ in some brachyuran crabs 

also secretes 3-dehyroxyecydsone and 25-deoxyecdysone (summarized in Subramoniam 2000). 25-

deoxyecdysone is the precursor to ponasterone A (PoA), the primary circulating ecdysteroid in the premolt 

stage of the crabs (Subramoniam 2000). Other sources for ecdysteroids are the ovary, epidermis, and the 

oenocytes (Delbecque et al. 1990). In many crustaceans, molting, and hence somatic growth, continues 

after maturity, with the result that the Y-organ is active in adults. For most crustaceans, growth and 

reproduction can be grouped into three functional categories. In the first category, represented by crab and 

lobster, reproduction occurs after a long intermolt period. The second category includes isopods and 

amphipods, the growth and reproduction of which are concurrent. The last category relates to the rapidly 

molting cirripedes, for which reproduction requires several molt cycles. Molting and limb regeneration are 

intertwined (Fingerman et al. 1998). When limb regeneration occurs, first a limb bud develops within a 

layer of cuticle, and then becomes free and unfolds when ecdysis occurs as part of the molting process. 

Synthesis and secretion of ecdysteroid by the Y-organs is inhibited by molt-inhibiting hormone, a peptide 

that is released from the sinus gland. 

31. It has been documented that there is synchronization between reproduction and molting. At least 

in mysids, accumulation of ovary ecdysteroid takes place during the premolt stage, when the hemolymph 

ecdysteroid levels rise sharply. It is presumed that the hemolymph ecdysteroids are transported to the 

ovary along with the yolk precursor material. This trend is seen in other species as well, and shows that 

the Y-organ is active during premolt and that it produces ecdysteroids that are transported to the ovaries. 

This observation was confirmed by Subramoniam (2000). A Y-organ ablation was performed on the 

shrimp, Lysmata seticaudata, which caused a subsequent depression of vitellogenin synthesis and 

retardation in ovarian growth. Further findings on the same shrimp revealed a failure of folliculogenesis, 

which is a necessary prerequisite for vitellogenin-uptake by oocytes during secondary vitellogenesis. 

32. Terpenoids, which are unique to arthropods (Quackenbush 1986), include methyl farnesoate 

(MF). MF is secreted by a mandibular organ and there is evidence that this compound is involved with the 

control of ecdysteroid synthesis. When a mandibular organ was experimentally implanted into the shrimp, 

Penaeus setiferus, there was a subsequent shortening of the molt cycle (Subramoniam 2000). Secretion by 

the Y-organ is controlled by methyl farnesoate, whereas inhibition is exercised by MIH from the X-organ 

sinus gland. The mandibular organ has also been implicated in the control of reproduction in crustaceans. 

Mandibular organ implants stimulated ovarian growth in the juvenile spider crab females and methyl 

farnesoate levels increased in the hemolymph and the mandibular organ during vitellogenesis in the crab, 

suggesting that this compound has a gonadotropic role similar to that of JH in insects. However, other 
ENV/JM/MONO(2006)22 

 28

studies showed no methyl farnesoate- level effects within the vitellogenic period in the lobster, for example 

(Subramoniam 2000). 

33. Serotonin (5-hydroxytrptamine, 5-HT), a biogenic amine, is one of the most important 

biologically active substances in animal kingdom as it regulates many physiological and behavioral 

functions (Moreau et al. 2002). 5-HT has been found in many invertebrate groups, including crustaceans. 

Moreau et al. (2002) documented its presence in mysids, although they did not study its specific function. 

Studies reviewed by Quackenbush (1986) suggested that 5-HT induces the release of molt-inhibiting 

hormone and crustacean hyperglycemic hormone from the eyestalks of decapods. 

34. Rather than providing a detailed review of crustacean endocrinology, the few paragraphs 

presented here summarize the four main types of compounds involved in the regulation of crustacean 

biology and show the interplay among them in regulating major physiological processes. In summarizing 

about 75 years of crustacean endocrinological studies, Fingerman (1997) concluded that despite the many 

significant advances, work in the filed “has really just begun.” This is especially true considering the tasks 

ahead in examining the potential for the disruption of crustacean endocrine systems by anthropogenic 

compounds. 

3.1.7 Echinoderms 

35. Echinoderms (e.g., sea stars, feather stars, sea cucumbers) are deuterostomes and are relatively 

closely related to vertebrates. Therefore, their endocrine systems may have some similarities to those of 

vertebrates and may share similar targets susceptibilities to chemicals known to have endocrine-disruptive 

effects on vertebrates (LeBlanc et al. 1999). Vertebrate sex steroids may play a role in echinoderm 

reproduction (Oberdörster and Cheek 2001). 

36. No ecdysteroids or juvenoids are known in echinoderms (echinoderms don’t molt). Processes 

under hormonal, neurohormonal, or local growth factor control include gametogenesis, spawning, growth, 

and regeneration (reviewed in LeBlanc et al. 1999). Reproduction influenced by steroids and 

neuropeptides (summarized in LeBlanc et al.). 

37. Sea stars synthesize two steroids, progesterone and testosterone, and steroidogenic pathway 

enzymes (3β-hydroxysteroid hydrogenase, cytochrome P450) occur in sea stars. Estrogen and estradiol 

synthesis has not been demonstrated, but a receptor for estradiol has been identified. High levels of 

progesterone in males have been demonstrated at the beginning of spermatogenesis. Therefore, it is very 

likely that steroids function in sea star reproduction. There is very little information about endocrine 

functions in other echinoderm groups. 

38. Gonad-stimulating substance (GSS), which has been found in radial nerve extracts and stimulates 

spawning, may be a neuropeptide. GSS indirectly stimulates release of a maturation-promoting factor that 

readies oocytes for fertilization. Other neuropeptides have been discovered in echinoderms and may 

regulate feeding. 

39. Regeneration is a form of asexual reproduction in some taxa, but also serves to replace lost body 

parts. Loss of body parts, which may be a defense from predation, most frequently involves arms (e.g., sea 

stars), but also often may involve sections of epidermis (e.g., sea cucumbers; Kropp 1982). Arm 

regeneration in feather stars (Crinoidea) occurs by the proliferation of migratory undifferentiated cells and 

is under nervous system control, which provides the primary regulatory factors (Candia Carnevali et al. 

2001b). 
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3.1.8 Summary 

40. The endocrine system of an invertebrate differs from that of a vertebrate organism in the type of 

endocrine glands present and in the chemical structure (and consequently in the function) of specific 

hormones that are produced. Invertebrates produce some hormones that vertebrates do not. For example, 

crustaceans and most other invertebrates produce di- and tri-iodothyronine, but have no thyroid gland, and 

the function of the thyronines is unknown. Invertebrates use hormones that are not found in vertebrates. 

Crustecdysone is in some ways analogous to a vertebrate’s estrogen hormone, but it is structurally, 

functionally, and metabolically different from the vertebrate hormone (J.M. Neff, personal communication, 

January 15, 2002). 

3.2 OVERVIEW OF POTENTIAL ENDOCRINE DISRUPTION IN AQUATIC 

INVERTEBRATES 

41. There are two basic types of compounds with the potential to disrupt endocrine systems: 

synthetic chemicals (xenobiotics) and natural plant chemicals (phytoestrogens) (Crisp et al. 1998). Some 

examples of xenobiotics are compounds used in plastics (nonylphenol, bisphenol-A), PCBs, and some 

pesticides. Phytoestrogens include hormone-mimicking substances contained in some agricultural plants 

and in paper mill effluent. These can be estrogenic or anti-estrogenic. 

42. Lafont (2000) described and gives examples of four levels at which EDCs can disturb endocrine 

systems. At the first level, EDCs block the availability of the precursors required for the synthesis of 

hormones. The interruption of hormone biosynthesis occurs at the second level. For example, some 

chemicals inhibit cytochrome P450s, which catalyzed biosynthesis, thus breaking the synthetic pathway 

leading to hormone production (Lafont 2000). The third level at which an EDC can act is on hormone 

catabolic processes. In this case, an EDC could act to increase the rate at which a hormone is catabolized, 

resulting in lower levels in an animal. At the fourth level, an EDC directly interferes with the actions of 

hormones. EDCs can act as agonists by binding to a hormone receptor and acting as that hormone would 

to regulate gene transcription (LeBlanc et al. 1999). EDC also may act antagonistically by binding to a 

hormone receptor without inducing its activity (LeBlanc et al. 1999), thereby obstructing normal endocrine 

function. 

43. The following sections present a brief overview of endocrine disruption in invertebrates that 

builds on recent reviews by LeBlanc et al. (1999), Oehlmann and Schulte-Oehlmann (2003), and Segner et 

al. (2003) among others. The overview is not meant to be exhaustive and is organized by major taxon 

within habitat type. Some information about endocrine disruption in aquatic arthropods, especially 

crustaceans, is included here, but a more detailed discussion, structured in the context of hormonal effects, 

is presented in Sections 7, 8 and 9. 

3.2.1. Endocrine Disruption in Freshwater Species 

44. Porifera: Hill et al. (2002) tested the effects of three chemicals known to disrupt endocrine 

systems in some animals (ethylbenzene, nonylphenol, bisphenol-A) on two species of freshwater sponges 

(Heteromyenia sp., Eunapius fragilis). The study was initiated with gemmules and tested the effects of 

these chemicals on growth (morphology and rate). Hill et al. found that higher doses caused a 

morphological abnormality that was similar across chemicals for both species. Germination rates for 

gemmules in the control treatments were relatively low (70% and <50% for each species) and germination 

could be inhibited at highest concentrations. Hill et al. could not determine if the observed effects were 

attributable to a disruption of an endocrine pathway or a more general toxic effect, yet argued that the 

former was likely (despite the general lack of information about sponge endocrinology). 
ENV/JM/MONO(2006)22 

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45. Hydrozoa: Pascoe et al. (2002) examined the effects of 17α-ethinylestradiol and bisphenol-A on 

the hydrozoan Hydra vulgaris and found no physical or physiological damage to the polyps at 

environmentally relevant concentrations (ng/L) of either chemical. However, they determined that toxicity 

occurred at relatively high chemical concentrations and concluded that signaling processes necessary for 

normal development, regeneration, and sexual reproduction were not affected by these estrogenic 

pollutants at low, environmentally relevant concentrations. 

46. Cladocera: Baldwin et al. (2001) documented the effects of several EDCs on the Daphnia 

magna. One experiment focused on exposure to 20-hydroxyecdysone (20-E), the crustacean molting 

hormone, and to ponasterone-A (PoA), an endogenous compound that has 20 times higher affinity for the 

ecdysone receptor. The 21-day exposure had little effect on reproduction for either compound, except at 

the highest concentrations. However, adults suffered high mortality rates and either did not produce 

broods or produced smaller broods. Second-generation effects were not observed as a result of 20-E 

exposure, but there was a slightly significant effect on reproduction from PoA exposure. The effect on 

reproduction could be attributed to the structure of PoA: it has fewer hydroxyl groups, and could be less 

easily metabolized as is 20-E. The resulting longer exposure could allow second-generation effects. It is 

also possible that the higher affinity for PoA to ecdysone receptor sites caused a limited effect on 

secondary vitellogenesis in developing daphnids, which manifested itself as reduced reproduction because 

of smaller brood size (Baldwin et al. 2001). 

47. Recently, many studies conducted by LeBlanc and coworkers have elucidated many aspects of 

daphnid endocrinology and have subsequently documented the impacts of EDCs. For example, Olmstead 

and LeBlanc (2002) determined that methyl farnesoate was the likely endocrine factor regulating the 

development of males in D. magna. They later predicted that a juvenile hormone analog, pyriproxyfen, 

would stimulate the production of males and found that this indeed occurred (Olmstead and LeBlanc 

(2003). Tatarazako et al. (2003) also found that methyl farnesoate and pyriproxyfen stimulated male 

production in D. magna, even under uncrowded, high-food (i.e., low stress) conditions. They also found 

that three additional juvenoids—fenoxycarb, methoprene, and JHIII—stimulated male production. The 

five juvenoids showed similar dose-dependant responses (reduction in fecundity, male production), 

although fenoxycarb and pyriproxyfen caused the effects at substantially lower concentrations than the 

other three compounds. Companion studies by Mu and LeBlanc (2002a, 2002b) showed that ecdysteroids 

are important in regulating daphnid development and that ecdysteroid antagonists, testosterone and 

fenarimol, interfered with normal development. Testosterone affects early and late developmental stages 

(acting as a hormone receptor antagonist), whereas fenarimol affected late development (acting as a 

hormone synthesis inhibitor). Mu and LeBlanc (2004), in a study with more significance to EDC 

evaluations, examined the potential synergistic effects of a fenarimol-testosterone mixture on daphnid 

development. Mu and LeBlanc focused on the different actions of each compound, predicting that 

fenarimol would lower ecdysone levels in daphnids, which would increase the testosterone binding to the 

ecdysone receptor with the net overall effect of increasing toxicity beyond that of either individual 

chemical. Their study confirmed this prediction and underscored the importance of considering the 

potential synergistic of mixtures of chemicals in the environment. 

48. Kashian and Dodson (2004) studied the effects of several vertebrate hormones on sex 

determination and development in D. magna. Among the hormones they tested, only long-term (26 days) 

exposure to progesterone affected sex determination. The second clutch of young produced by D. magna

exposed to 100 μg/L progesterone contained significantly more males than controls. The effect 

disappeared with succeeding clutches. Daphnids exposed to 100 μg/L testosterone had significantly 

reduced fecundity compared to control animals. Two hormones affected daphnid growth. Exposure to 

diethylstilbestrol (100 μg/L) significantly reduced growth, whereas exposure to gonadotropin (30 μg/L) 

significantly increased growth compared to controls. Resting egg production and molting were not 

affected by any of the vertebrate hormones tested. 
 ENV/JM/MONO(2006)22 

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49. Other studies of endocrine disruption in daphnids showed that styrene dimers and trimers reduced 

fertility in Ceriodaphnia dubia (Tatarazako et al. 2002) and the 17α-ethinylestradiol decreased the number 

of offspring in Daphnia magna (Goto and Hiromi 2003). Goto and Hiromi also found that another 

contraceptive ingredient, norethindrone, did not affect offspring production or sex ratio in D. magna. 

50. Amphipoda: Vandenbergh et al. (2003) used a multigenerational assay to examine the effects of 

sublethal doses of 17α-ethinylestradiol, a synthetic estrogen, on sexual development in Hyalella azteca. 

They compared the second gnathopod size and reproductive tract histology among treated and untreated 

males and found that F1 males exposed to the lowest 17α-ethinylestradiol doses had significantly smaller 

second gnathopods than controls, whereas the second gnathopods of those exposed to higher doses did not. 

This suggested a U-shaped response where effects were observed at a low concentration, but were masked 

at high concentrations. Vandenbergh et al. also observed cellular abnormalities in males exposed to all 

concentrations. Among these were larger and more spherical germ cells, a hollow cell structure, and less 

dense cytoplasm, features that are analogous to female gonad morphology. Oocyte-like structures were 

detected at some concentrations and smaller spermatids, fewer vas deferens, and irregular spermatogonia 

were found at highest concentration. The observed histological abnormalities indicated some degree of 

hermaphroditism in the exposed amphipods. The authors concluded that sublethal exposures of 17αethinylestradiol may affect sexual development in H. azteca, but they also noted that test concentrations 

were higher than those reported to affect vertebrates and higher than those observed in natural 

environments. 

51. Watts et al. (2001b) looked at the toxicity of 17α-ethinylestradiol and bisphenol-A and effects of 

the chemicals on the precopulatory guarding behavior of Gammarus pulex. They found that 17αethinylestradiol was more toxic than bisphenol-A, but that there was no disruption of precopula or the 

propensity to reestablish it except at the highest concentrations, which were close to doses that were 

determined to be acutely toxic. Therefore, these two xenoestrogens had no effect on any endocrine 

systems that facilitate the behavior. In a follow-on study, Watts et al. (2002) designed a longer study 

specifically to determine if the lack of response to 17α-ethinylestradiol in the acute test predicted a similar 

lack of response at the population level in G. pulex. The results showed that all of the 17α-ethinylestradiol 

concentrations that were tested showed population increases greater than those in control treatments (the 

control treatments did increase over the starting population values). The difference was attributed to the 

recruitment of neonates and juveniles. The dosed treatments also had more females than the control 

treatments and had a M:F ratio of ~0.4–0.5:1 versus the ratio of 1:1 attained in the control treatments. The 

authors concluded that exposure to 17α-ethinylestradiol resulted in significant increase in population size, 

despite prediction of no effect on reproductive behavior that was derived from the acute assay. The 

increase in population size was attributed to an accelerated female maturation rate, such that more young 

were produced earlier, and an increase in number of females. These two studies underscore the need for a 

chronic assay to develop a more complete understanding of potential EDC effects on populations. 

52. Insecta (Chironomus): Most of the studies of endocrine disruption in midge larvae has been done 

on Chironomus riparius. In the sole study found during this review that directly concerned endocrine 

disruption in C. tentans (also known as C.dilutus), Kahl et al. (1997) conducted an assay to examine the 

effects on 4-nonylphenol on the life cycle of the midge. The only significant result was on 20-d survival at 

highest concentration. There were no effects on egg production or viability, emergence time, and sex ratio. 

Kahl et al. (1997) noted that some egg cases were deformed, particularly at the high nonylphenol 

concentration. 

53. Studies on potential endocrine disruption in C. riparius have considered two basic types of 

endpoints, mouthpart deformities and life cycle reproductive parameters. Deformities of the mouthparts 

(mentum, mandibles, epipharyngeal pecten) have been observed in association with some contaminated 

sediments. Physiological disturbance during molting are thought to cause the deformities. Meregalli et al. 
ENV/JM/MONO(2006)22 

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(2001) tested the hypothesis that hormonal disruptions during molting contributed to the development of 

the deformities by subjecting the midge larvae to the known endocrine disrupter 4-nonylphenol. Mentum 

deformities were significantly more common in larvae exposed to sublethal concentrations of nonylphenol 

than in the control larvae. The mandibles and epipharyngeal pecten showed few deformities. In a more 

broadly scoped study, Watts et al. (2003) studied the effects of sublethal doses of the estrogenic endocrine 

disruptors 17α-ethinylestradiol and bisphenol A on molting and mouthpart structures in C. riparius. The 

highest concentration (1 mg/L) of both chemicals significantly reduced larval weight and delayed molting. 

Watts et al. noted that the high concentration was environmentally unrealistic and much greater than the 

concentration known to elicit a response in fish. The remaining doses of either chemical did not affect 

either parameter. The lowest concentration (10 ng/L) of each chemical caused deformity of the mentum 

but not the other two primary mouthparts. 17α-ethinylestradiol had a slightly stronger effect than 

bisphenol A on the mentum. No mouthpart deformities occurred at the highest concentration of each 

chemical. The mouthpart studies also highlights that classical dose-response curves often do not apply to 

EDCs, but that inverted U-shaped curves (no/low effects at the concentration extremes, high effects at 

middle concentrations) more accurately depict the response (Watts et al. 2003). 

54. Watts et al. (2001a) used a two-generational study of C. riparius to evaluate the effects of 17αethinylestradiol and bisphenol-A on the midge life cycle. They found that the main effect of the two 

chemicals was delay of emergence at higher test concentrations, especially for the second generation. The 

delay did not affect the typical protandrous emergence pattern in chironomids (males emerge before 

females). There was no effect on egg production or egg viability. One interesting effect was the alteration 

of the second generation adult sex ratio where males outnumbered females 2:1 at all but the highest 17αethinylestradiol concentration. Watts et al. thought that this effect was aligned with that expected for an 

estrogenic compound and suggested that this could mean the chemical did not act as an estrogen in this 

particular case. Bisphenol A did not affect sex ratio. 

55. Hahn et al. (2001, 2002) tested the effects of tebufenozide, a pesticide that acts as a molting 

hormone agonist, on C. riparius development and vitellogenin/vitellin (Vg/Vn) levels. The molting 

hormone agonist was anticipated to accelerate development. Tebufenozide affected development, but that 

effect did not occur until the final molt from pupa to adult. Hahn et al. pointed out that this effect was 

counter to that desired for the pesticide’s intended butterfly targets, stimulation of an early molt that leads 

to death. The effects levels determined during this chronic study (e.g., LC50 = 21.14 μg/L) were much 

lower than those determined for acute exposures. The 2002 study also examined the effects of bisphenol-A 

and 4-nonylphenol on Vg/Vn immunoreactivity. Males exposed to most contaminant concentrations 

showed reduced Vg/Vn immunoreactivity. Females were not affected by the contaminants, except the 

highest bisphenol-A concentration reduced Vg/Vn levels. These results were not expected because 

bisphenol A and nonylphenol are known to stimulate vitellogenesis in vertebrates. The observed responses 

were not dose-dependent as all concentrations reduced Vg/Vn somewhat equally. 

56. Hahn and Schulz (2002) found that relatively short-term exposure to tributyltin reduced 

ecdysteroid synthesis in female C. riparius larvae at test concentrations as low as 50 ng/L, whereas males 

showed increased biosynthesis at a concentration of 500 ng/L. Imaginal discs developed more slowly in 

females, but faster in males at all test concentrations than in the respective control treatments. The authors 

thought that these gender-specific differences showed that either the ecdysteroid pathways were impacted 

differently, or that different reactions occurred to the same impact. 

3.2.2. Endocrine Disruption in Estuarine and Marine Species 

57. Cnidaria: Tarrant et al. (2004) evaluated the effects of exogenous estrogens (17β-estradiol, 

estrone) on scleractinian coral reproduction. They determined that 29% fewer egg bundles were released 

by Montipora capitata colonies in estradiol-treated colonies than in control colonies. Thus, estrogens seem 
 ENV/JM/MONO(2006)22 

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to be natural bioregulators in corals, which suggests that exogenous estrogens can reduce coral fecundity. 

However, the number of eggs per bundle did not differ in treated corals than from corals in control 

treatments. Growth in Porites compressa was significantly reduced in treated corals at estrone levels only 

slightly greater than ambient levels. However, the mode of action of estrogen in corals is unknown. 

Tarrant et al. suggested that these results (and others for invertebrates) imply that the potential disruptive 

effects of estrogens are not limited to interruption of mammalian reproduction. 

58. Annelida: The review prepared by LeBlanc et al. (1999) mentioned that there was only one 

known case of endocrine disruption in annelid worms; juvenile hormone and JH analogs stimulate larval 

settling and metamorphosis in Capitella. In a 78-d study, Hansen et al. (1999) found that the lowest 

treatment of sediment-bound 4-n-nonylphenol stimulated asymptotic body volume growth and increased 

mean brood size, but that these effects did not translate to changes in population growth rates. The highest 

concentration of nonylphenol significantly reduced several reproductive metrics, including brood size, 

volume-specific fecundity, time to first reproduction, and the growth rate of individuals. The authors did 

not link the observed effects to possible endocrine disruption, although they noted that the stimulation at 

the low dose was another case of hormesis, which has been documented for other organisms. 

59. Mollusca: The most frequently cited examples of endocrine disruption in mollusks concern 

imposex, which is an irreversible condition in gastropod snails in which females develop secondary male 

sex organs (Matthiessen and Gibbs 1998). Imposex, which is most frequently a response to exposure to 

tributyltin contained in anti-fouling paints, may result from the inhibition of aromatase or other steroid 

precursors, or from direct impacts to neurohormones (Rotchell and Ostrander 2003). Since the topic has 

been the subject of very many studies and has been considerably reviewed, it will not be discussed in detail 

here. Reviews and other significant discussions of imposex are included in Matthiessen and Gibbs (1998), 

Evans and Nicholson (2000), the series of studies by Oehlmann and others (Oehlmann et al. 2000, SchulteOehlmann et al. 2000, Tillman et al. 2001), and Axiak et al. (2003). 

60. Nice et al. (2003) studied the effects of single short-term (48 h) nonylphenol exposure of 7-d old 

larvae of oyster Crassostrea gigas on long-term physiological processes. Exposure to environmentally 

relevant levels of nonylphenol resulted in a sex ratio biased towards females and an increased incidence of 

hermaphroditism. The exposure also affected gamete viability, severely impacting embryonic and larval 

development in second generation. 

61. Jobling et al. (2003) tested effects of 17α-Ethinylestradiol, 4 tert-octylphenol, bisphenol-A, and 

sewage effluent on egg and embryo production in a common freshwater European snail Potamopyrgus 

antipodarum. All estrogen and xenoestrogen treatments stimulated embryo production, except at the 

highest concentration, which had an inhibitory effect. These reproductive effects were generally similar to 

those the authors observed for three fish species (fathead minnow, rainbow trout, carp) in a companion 

study. Jobling et al. suggested that this snail is a sensitive species whose testing against estrogens may be 

relevant to estrogenic activity in vertebrates. 

62. Crustacea: Marine crustaceans are important organisms to include in the evaluation of the 

adverse consequences of EDCs, and the selection of suitable species is the one focus of the present review. 

Early studies of the effects of EDCs on estuarine crustaceans, most of which have occurred within the last 

few years, have focused on three primary groups, barnacles, copepods, and decapods (Ingersoll et al. 1999, 

Hutchinson 2002). The results of these studies showed that some crustacean groups were affected by 

exposure to the tesed EDCs, but others were not and, therefore, extrapolation of the results from testing 

one crustacean group to another is problematic. 

63. Barnacles.—Billinghurst et al. (1998, 2000) examined the effects of two estrogens, 4-nnonylphenol and 17β-estradiol on larval settlement and the production of a larval storage protein (cypris 
ENV/JM/MONO(2006)22 

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major protein, CMP) in Balanus amphitrite. Cyprids use CMP during settlement and the early postsettlement development. Because CMP is structurally related to vitellin, which is analogous to 

vitellogenin, it can be used as a biomarker of estrogen exposure in lower vertebrates. The expectation in 

these studies was that cyprid settlement might be affected by the stimulation of CMP synthesis after larval 

exposure to environmental estrogens. The results of the 1998 study, however, showed reduced settlement 

after exposure to both estrogens, but that the cause was not related to endocrine disruption. The second 

study (Billinghurst et al. 2000) measured levels of CMP and found that they were elevated after exposure 

of nauplii to low levels of the estrogens. 

64. Copepods.—Hutchinson et al. (1999a, 1999b) found that exposure to several steroids had no 

effect on the survival and development of harpacticoid copepod (Tisbe battagliai) nauplii and cautioned 

against extending the reported effects of steroid exposure in some species of crustaceans to the group as a 

whole. At about the same time, Bechmann (1999) showed that high levels (>62 μg/L) of nonylphenol 

were acutely toxic to T. battagliai, but that exposure to a low level (31 μg/L) did not affect any of the 

measured life-table parameters (survival, sex ratio, fecundity). Breitholtz and Bengtsson (2001) did not 

find evidence of endocrine disruption in the harpacticoid copepod Nitocra spinipes after exposure to the 

estrogens 17β-estradiol, 17α-ethinylestradiol, and diethylstilbestrol. Chandler and coworkers studied 

potential endocrine disruption in a third harpacticoid species, Amphiascus tenuiremis. Bejarano and 

Chandler (2003) used chronic multi-generation exposures to evaluate the effects of the herbicide atrazine 

on reproduction and development in A. tenuiremis. While atrazine did not have significant effects on 

several parameters (e.g., time to maturity, time to egg extrusion, time to hatching), some concentrationrelated effects occurred. Reproductive failure (mating pairs unable to produce living offspring or females 

unable to extrude more than one brood) increased with atrazine dose. Nauplii production by F0 females 

was reduced at the highest concentration and that by F1 females was reduced at all concentrations. Both 

factors combined to reduce total population growth of the F1 generation at doses lower than those 

considered “safe” for chronic exposures (26 μg/L). Chandler et al. (2004a) found that sublethal, 

environmentally relevant concentrations of the insecticide fipronil (and also its degradation products; 

Chandler et al. 2004b) delayed the maturation of A. tenuiremis copepodites to adults, and reduced or 

virtually eliminated the production of young. These effects were modeled and predicted a 62% decline in 

the population size of the copepod at the lowest parent concentration tested (0.16 μg/L). Fipronil is a 

GABA-inhibitor and it is likely that the effects were mediated via disruption of the neural system. Still, a 

follow-up to Chandler 2004a showed that it was actually only males that were affected at these low levels 

of fipronil, as indicated by lowered ecdysteroid levels, elevated male vitellogenin, and malfunction in male 

reproductive ability (Cary et al. 2004). This may consequently be regarded as an example of both neural 

and possibly endocrine disruption. 

65. Copepods probably have been tested against more potential EDCs than any other marine group. 

For example, Anderson et al. (2001) tested the effects of 14 compounds, including natural vertebrate 

hormones (17β-estradiol, estrone, testosterone, progesterone), natural invertebrate hormones (20-

hydroxyecdysone, juvenile hormone-III), hormone antagonists (flutamide, tamoxifen, hydroxyflutamide), 

xenoestrogens (17α-ethinylestradiol, 4-octylphenol, bisphenol-A), and environmentally relevant 

compounds (nonylphenol ethoxylate, diethyl phthalate) on larval metamorphosis (nauplius to copepodite) 

in the calanoid copepod Acartia tonsa. The important points of this study were that chemicals differed in 

their relative effects on survival and larval development. Some of the chemicals tested affected 

development at concentrations well below those determined to be toxic. Others delayed development only 

at concentrations close to those that were toxic, thus the main effect of the chemical was likely toxicity. 

Chemicals having very similar effects on toxicity, and having similar octanol/water partition coefficients 

(Kow), could have very different effects on larval development. For example, flutamide and testosterone 

have the same Kow and had similar effects on copepod survival, but flutamide was a much stronger 

inhibitor of larval development than testosterone. 
 ENV/JM/MONO(2006)22 

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66. Amphipods.—There have been few studies of endocrine disruption in marine amphipods. Brown 

et al. (1999) found that exposure to 4-nonylphenol reduced growth in the gammaridean amphipod 

Corophium volutator, but that this was probably a general response to the exposure rather than an 

interaction with molting hormones. They also reported that males exposed to 4-nonylphenol had larger 

second antennae than those in control treatments and suggested that the compound may have acted on the 

androgenic gland. Field studies suggested that intersexuality observed in an estuarine amphipod, 

Echinogammarus marinus, might be indicative of endocrine disruption, although a causative link was not 

established (Ford et al. 2004a, 2004b). 

67. Mysids.—Mysid crustaceans have been used in regulatory (and other) toxicity testing for more 

than 20 years. Standard testing protocols have been developed for some species. Despite that little is 

known about general endocrine functions in mysids (Ingersoll et al. 1999) and there have been few direct 

links between potential EDCs, beyond certain IGRs, and endocrine disruption in mysids, they have been 

suggested as providing a useful model of the hormonal control of crustacean molting (Cuzin-Roudy and 

Saleuddin 1989). McKenney and Celestial (1996) studied the effects of methoprene, which is a JH analog 

used to control mosquitoes, on Americamysis bahia and found that mysids grown at sublethal 

concentrations were smaller, had a longer time to the production of the first brood, and produced fewer 

young per female than control animals. They suggested that the effects shown were likely from the 

interruption of endocrine function by the methoprene. 

68. Advances in biological control agents to control insect pests have inspired the synthesis of insect 

growth regulators (IGR), which find their way into the estuarine environment by either direct or indirect 

application. Crustaceans, which along with insects are in the phylum Arthropoda, could also be sensitive 

to these compounds. The mysid group has been shown to be among the most sensitive members of the 

estuarine community (McKenney 1982, 1985, 1986, 1996; Nimmo and Hamaker 1982; Nimmo et al. 

1981). One study focused on exposure of Americamysis bahia to methoprene, a JH analog (McKenney 

and Celestial 1996). The goal was to determine whether typical application rates shown to control 

mosquito larvae also cause problems for non-target organisms. The results showed a significant effect 

during the mysid life cycle test. Total lethality occurred at 125 μg/L in a 4-day test. Similar 

concentrations caused significant mortality in the larvae of an estuarine crab and shrimp. Other sublethal 

endpoints, such as reduced growth (weight), longer time to first brood, and a significant reduction in brood 

size were also observed. These results suggest that methoprene could interfere with the endogenous 

endocrine system, which uses hormones that act like JH. Retarded growth rates were also accompanied by 

bioenergetic disruption, resulting in lower net growth efficiency values. This suggested that increased 

metabolic demands reduced the amount of assimilated energy available for new tissue production 

(McKenney 1982, 1985). The delays in mysid first brood production could be the result of slowing sexual 

maturity and/or embryogenesis. Diminished reproductive success could be the result of inhibited 

vitellogenesis, modifications in ovarian development, or disruption of successful embryogenesis. In either 

case, further work is required with the mysid to determine a more conclusive cause-and-effect relationship 

between potential EDCs and their effects, as observed by test measurement endpoints. 

69. Verslycke et al. (2002, 2003a, 2003b, 2004b) studied testosterone metabolism by mysids 

(Neomysis integer) and examined the changes in energy and testosterone metabolism after exposure to 

potential EDCs. Because these studies focused on potential endpoints, they are presented in more detail in 

Section 6.4.3. 

70. Decapods.—Several studies have investigated the effects of EDCs on crustacean life cycles using 

decapod larvae as the test organisms (e.g., Lee and Oshima 1998; McKenney et al. 1998; Nates and 

McKenney 2000). Several studies (reviewed by McKenney 1999) have reported effects of the exposure of 

decapod larvae to JH analogs suggestive of the interruption of endocrine processes, but direct links were 

not established. More recently, Nates and McKenney (2000) found that exposure to the pesticide 
ENV/JM/MONO(2006)22 

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fenoxycarb disrupted lipid metabolism in mud crab larvae and suggested that the compound could be 

interfering with the endocrine regulation of lipid metabolism. Exposure to fenoxycarb delayed maturation 

of xanthid crab (Rhithropanopeus harrisii) larvae by about 25% compared to controls (Cripe et al. 2003). 

Some evidence of endocrine disruption in decapods was provided by Snyder and Mulder (2001) who found 

that exposure of lobster (Homarus americanus) larvae to the pesticide heptachlor altered ecdysteroid 

hormone levels that were linked to delays in molting. Metals such as mercury, cadmium, and zinc have 

been reported to affect molting and limb regeneration in crabs. Organic compounds such as Aroclor 1242 

and sodium pentachlorophenate reportedly had a similar effect–inhibition of limb regeneration–in the grass 

shrimp, Palaemonetes pugio, but had no effect on the molting cycle. This suggests that these chemicals act 

directly on limb development, but not on the hormonally controlled molting cycle (Fingerman et al. 1998). 

71. Echinodermata: Candia Carnevali et al. (2001a, 2001b) studied the feather star Antedon 

mediterranea (Crinoidea) arm regeneration in response to PCB exposure as endocrine disruptor effect. 

They found that the early phases of regeneration were not different in feather stars exposed to PCBs than in 

those from control treatments, but noticeable effects occurred later (first at ~7 d, more extensive effects at 

~14 d). The primary effect observed in exposed treatments was increased growth rate. Effects, including 

rearrangement and/or dedifferentiation of some tissues in the regenerating stump, also occurred at cellular 

level. No abnormal histological effects on feather star arms were observed. The authors concluded that 

the observed growth and tissue/cellular effects from the exposure PCBs were consistent with 

pseudoendocrine activities and steroid dysfunction. 

3.2.3 Summary 

72. Rather than summarize the various impacts, or lack of impacts, of the EDCs that have been 

studied to date, this section will highlight some of the general findings that are important to consider. 

• Chronic assays often reveal impacts by a chemical at doses that are much lower than those 

eliciting effects during acute exposures. This emphasizes the importance of chronic assays in an 

EDC evaluation program (e.g., Watts et al. 2001b, 2002). 

• Potential EDCs may have effects in invertebrates other than those anticipated by knowledge of 

their actions in vertebrates (e.g., Hahn et al. 2002, Watts et al. 2001b). 

• Pesticides that target certain aspects of insect physiological process may affect non-target 

organisms in ways not predicted by the pesticide’s desired action (e.g., Hahn et al. 2001, 2002). 

• Males and females may react differently to chemicals that are not thought to cause genderspecific responses (e.g., Hahn and Schulz 2002). 

• Chemicals that have very similar effects on toxicity, and have similar octanol/water partition 

coefficients (Kow), can have very different effects on endocrine-controlled processes (e.g., 

Anderson et al. 2001). 

• Some of the endpoint responses (e.g., larval development, ecdysone balance) measured in a study 

of one or more potential EDCs may also be caused by non-EDC stressors (e.g., food 

quality/quantity, hypoxia, thermal stress). 

73. Finally, and perhaps most importantly, chemical mixtures may show synergistic impacts that are 

not evident in the actions of the individual components (e.g., Mu and LeBlanc, 2004). In nature, most 

animals will almost certainly be exposed to chemical mixtures rather than to single compounds, thus 

considering the potential effects of mixtures is important to an EDC evaluation program. However, to be 
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able to predict the possible outcomes of mixtures, at a minimum the individual impacts and mechanisms of 

action of the component chemicals need to be understood. 

3.3 EXTRAPOLATION ISSUES 

74. The traditional practice in toxicology of testing the effects of a stressor on individuals from one 

or a few species has raised two major issues that are of concern to an EDC evaluation program. Simply 

put, can data collected from tests involving individuals of a species be used to predict potential impacts to 

other species, or to populations? These two extrapolation issues are discussed in this section. 

3.3.1 Extrapolation from Taxon to Taxon 

75. It is not practical or even possible, to conduct toxicity tests on every species in a particular 

ecosystem (e.g., Hutchinson 2002), and therefore, the reasonableness of extrapolating the effects of a 

stressor (e.g., an EDC) observed for one species to another is an important concern (Segner et al. 2003). 

There are several approaches to establishing the transferability of the results from testing one species to 

another. One method is to compare the structural and functional similarity of endocrine systems and/or 

hormones among taxa. For example, Oberdörster and Cheek (2001) reviewed studies on the ecdysteroid 

systems of various arthropods, mentioning that the structure of ecdysone in crustaceans is identical to that 

in insects and that the primary function of the hormone in both groups is to regulate molting. They 

concluded that information on this system in one arthropod species should be applicable to another. More 

support for this type of argument is that there is some evidence that peptides are conserved through various 

evolutionary lineages from invertebrates to vertebrates (Anctil 2000). This could be interpreted to mean 

that effects on particular ecdysteroid or peptide systems in one species are likely to be the same as in 

another species that has the same or very similar system. There are several issues that render this 

reasoning inadvisable. Lafont (2000) stressed that one of the key general observations concerning 

invertebrate endocrine systems was that molecules may be structurally related and yet have very different 

functions in different taxa. Chang et al. (2001) reviewed many studies and showed that the same hormone 

may have functions that differ among the different life stages of an individual. Finally, even if a hormone 

can be shown to have the same function in a taxon, the sensitivity of that hormone to disruption by a 

chemical could vary greatly among taxa. For example, Watts and Pascoe (2000) who clearly demonstrated 

that Chironomus tentans (also known as C.dilutus) and C. riparius differed significantly in their responses 

to the same chemical stressor. These observations argue for caution in the use of data from a few species 

to craft generalizations about many (Lafont 2000). 

76. Despite this difficulty in extrapolating from one taxon to another, a program to evaluate potential 

EDCs can be successful if it selects a sensitive taxon that is ecologically relevant (Nimmo and Hamaker 

1982). This approach may not offer protection to all species, but should at least offer some degree of 

protection to critical parts of the ecosystem that have similar exposure probabilities. The program should 

also develop robust test protocols (Hutchinson 2002) that can be used efficiently by many laboratories and 

eventually should include representatives from several taxonomic groups (Hutchinson 2002, Oehlmann 

and Schulte-Oehlmann 2003). 

3.3.2 Extrapolation from Individuals to Populations 

77. The toxicological effects of chemicals, including potential EDCs, as measured during tests of 

exposure to individuals may not have the same potential to predict possible impacts to populations because 

the population response may not occur at the same exposure levels as the individual response or may be 

masked by another stress response. Therefore, the challenge posed to programs charged with assessing the 

potential risks of EDCs is to establish likelihood that a chemical, or suite of chemicals, will have adverse 

effects on populations (Crisp et al. 1998, Gleason and Nacci 2001). Thus, the critical need is to be able to 
ENV/JM/MONO(2006)22 

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detect specific cause-and-effect relationships between a chemical and the responses observed in the field 

when several factors may cause similar responses (Crisp et al. 1998). 

78. Two general approaches to extrapolating from EDC effects on individuals to potential impacts to 

populations have been identified. One approach uses data gathered during bioassays to attempt direct 

extrapolation to populations. The other approach uses population modeling techniques to connect effects 

observed on individuals to potential population impacts. 

79. As an example of the first approach, Watts et al. (2002) used survival and sublethal endpoints 

(growth, development, reproduction) to attempt to predict effects at higher levels of organization. They 

used a life table or demographic study where they recorded age-specific mortality and fecundity of 

individuals during life span of a single cohort and used those data to estimate the intrinsic rate of 

population increase. Watts et al. (2001b) had done an earlier acute study that showed no effect of 17αethinylestradiol on the freshwater amphipod Gammarus pulex. A subsequent chronic exposure resulted in 

changes in maturation rate and sex ratio that translated to measurable effects at the population level (at 

least in the laboratory) (Watts et al. 2002). Watts et al. concluded that chronic tests provide a more 

integrated approach to predicting population responses. 

80. McTavish et al. (1998) presented a general modeling approach to the evaluation of potential 

EDCs. McTavish et al. mentioned that many EDCs may persist in natural environments for some time 

because they are lipophilic and have fairly slow decay rates. These persistent EDCs may change mortality, 

reproductive, and life-stage transition rates, might become noticeable only after some delay, could affect 

different life stages present, and may affect offspring of exposed individuals. This introduces a level of 

complexity into EDC evaluations that may interfere with efforts to use typical toxicity test data to predict 

population-level effects. Mathematical models are useful tools to investigate these complex dose-response 

relationships (McTavish et al. 1998, Gleason and Nacci 2001). McTavish et al. built a general model that 

enabled mortality, reproductive, and life-stage transition rates to be evaluated alone or in combination. 

The model also allowed delayed responses and transgenerational impacts to be analyzed. The model 

evaluated the effects of two dosing scenarios; a single pulse of a chemical into a system and its subsequent 

decay, and continuous dose of a chemical. McTavish et al. argued that the effective dose response should 

be based on chemical concentrations in the test animals rather than those in the water because traditional 

dose-response assessments based on water concentrations may not detect all potential risk to natural 

systems. This approach probably is not very practical for many invertebrates, especially small arthropods 

because the amount of tissue required for the analyses would significantly increase the number of animals 

required for the test. The authors concluded that models are very useful for examining the changes from 

chemical exposure to the effective dose, and from that dose to populations. Models can synthesize 

laboratory toxicity test data and use them to extrapolate to a composite picture under a variety of dosing 

regimes and also can be used to assess the relative importance of various bioassay endpoints. 

Combinations of stressors can act synergistically to cause significant effects on individuals or populations 

even though the individual compounds did not when tested separately (Arnold et al. 1996, Mu and LeBlanc 

2004). These synergistic effects can be evaluated by the model that McTavish et al. used. 

81. Kuhn et al. (2000, 2001) applied the general principals proposed by McTavish et al. in a general 

evaluation of the ecological relevance of mysid toxicity tests using Americamysis bahia. Kuhn et al. 

(2000) used a model to predict the concentration of a contaminant that would result in no population 

growth and compared that concentration to standard toxicity test data. They also evaluated which test 

endpoint, survival or reproduction, was better at predicting population responses. Later, Kuhn et al. (2001) 

evaluated the ability of their age-classified projection matrix model to predict the population response of A. 

bahia over more than three generations maintained in the laboratory. Data incorporated into the model 

were gathered from daily records of survival and reproduction, which begins at about Day 17, during 

toxicity testing with several types of chemicals (e.g., metals, organic compounds). The analyses showed 
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that the population growth rates were dependent on concentration for most of the chemicals tested. Kuhn 

et al.’s comparison of endpoints showed that the LC50 estimated from a 96-h exposure and the chronic lifecycle test endpoints were highly correlated to population changes (r = 0.96 and 0.93, respectively). 

However, linear regression analysis showed that the life-cycle test was better than the LC50 at predicting 

population-level effects. Within the chronic assay, reproduction was more strongly correlated than 

survival to population growth rate (r = 0.98 and 0.79, respectively). The model was verified (Kuhn et al. 

2001) by using the results of a 28-d life-cycle test conducted with para-nonylphenol as the toxicant, to 

generate data that allowed the model predictions of mysid abundances at the end of the multigenerational 

(55-d) assay. The predicted abundances then were compared to actual abundances from the assay. The 

results showed that the model could predict population effects “reasonably well” (Kuhn et al. 2001) in the 

laboratory. The model-predicted concentration (16 μg/L) at which no population growth would occur did 

not differ significantly from that derived from the multigenerational test data (19 μg/L). The data from the 

28-d test also were used to calculate a chronic exposure value (12 μg/L), which was compared to the 55-d 

“no-growth” concentration. The lower value was selected as the conservative concentration at which the 

population should be protected from exposure to para-nonylphenol. 

82. While either or both approaches may be appropriate, the important consideration is that if 

population-level impacts are ignored, the probability that the direct effects of EDCs on ecosystems will be 

underestimated increases (McTavish et al. 1998). 
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4.0 AQUATIC ARTHROPODS IN THE EVALUATION OF POSSIBLE ENDOCRINE 

DISRUPTION 

83. Many anthropogenic pollutants eventually end up in the world’s oceans, carried there through 

riverine and estuarine pipelines (Nimmo and Hamaker 1982). Since the mid-1990s, there is an increased 

awareness that many sewage constituents or chemicals associated with industrial production that enter the 

environment can disrupt endocrine systems, and that these compounds will likely affect marine organisms 

(Depledge and Billinghurst 1999; Oberdörster and Cheek 2001). Although early concern over EDCs 

focused on vertebrates, attention recently has broadened to include invertebrates because they are 

ecologically important and have distinctive endocrine systems that differ from those of vertebrates. 

Estuaries, which are intrinsically and commercially important ecosystems, are among the earliest recipients 

of EDCs. Among the many estuarine organisms that could be adversely affected by these compounds, 

crustaceans are good candidates for study of potential impacts. Some of these species are discussed in this 

section. Attention here is focused on aquatic arthropods (insects and crustaceans) because they are often 

among the most abundant organisms in freshwater and estuarine systems, form vital links in estuarine food 

webs, and are known to be susceptible to the effects of EDCs. 

84. Hutchinson et al. (2000) offered a counter argument to the inclusion by the EDSTAC and EDSP 

of an arthropod, specifically a daphnid or mysid, in the endocrine screening program at this time. The 

principal objection applicable to both taxa was the general data gap in the basic endocrinology of mysids 

and daphnids. They argued that the assertion by EDSTAC that vertebrate estrogen or androgen disruptors 

could interfere with ecdysteroid activity was only speculation and that any effects on reproduction or 

development by known vertebrate EDCs have not been directly linked to ecdysteroid or juvenile hormone 

activity. This is indeed an important issue to consider in this context, but at the same time, the 

understanding of invertebrate endocrine systems is rapidly improving (Section 3.0) and the links between 

hormones and processes in crustaceans are constantly being established. For example, in companion 

studies, Mu and LeBlanc (2002a, 2002b) showed that ecdysteroids regulate the embryonic development of 

daphniids and that this development could be interrupted by testosterone and fenarimol, a fungicide used in 

agriculture. Testosterone did not display ecdysteroidal activity, but interfered with ecdysteroids by 

outcompeting them to occupy the ecdysteroid receptor (Mu and LeBlanc 2002a). Fenarimol acted by 

reducing body levels of ecdysone in individual daphnids (Mu and LeBlanc 2002b). Vandenbergh et al. 

(2003) found that the synthetic estrogen 17α-ethinylestradiol altered sexual development in individuals of 

the amphipod Hyalella azteca. Males did not develop secondary sex features, reproductive system 

morphology was altered, and sex ratios in exposed groups favored females. Also, recent studies of 

crustaceans have identified that methyl farnesoate has a role in regulating larval development in lobsters 

(summarized by Chang et al. 2001) and stimulates gonad development in a freshwater prawn (Nagaraju et 

al. 2003). The industrial chemical bisphenol-A has also been shown to to delay development of young 

stages in calanoids, harpacticoids and cladocerans (Andersen et al. 2001; Marcial et al. 2003; Mu et al. 

2005) As a disrupting mode of action, bisphenol-A was supposed to exert its activity via an enhancement 

of the activity of methyl farnesoate (Mu et al. 2005), thereby decreasing larval and juvenile development. 

As Verslycke et al. (2003a) stated, the discussion over whether or not an observed effect is directly related 

to endocrine disruption or not may be appropriate, but neither that discussion, nor the cause of the effect, is 

of concern to population maintenance of the affected organism. 

4.1 INSECTA 

 85. Although many insect species might be chosen for use in EDC evaluations the focus in this 

section is on the chironomids, which are ecologically important, have had well-developed testing protocols 
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prepared and evaluated, and have been widely used in general toxicological testing in addition to endocrine 

disruptor screenings. 

4.1.1 Midge Larvae (Chironomus tentans [or C. dilutus] and C. riparius) 

86. Natural History: Chironomus spp. are true midges (Order Diptera) that have a widespread 

worldwide distribution (Pennak 1989). Adult midges are small, nonbiting, and swarm near bodies of water 

at night. Larval stages are aquatic, inhabiting fine tubes in the upper layers of the sediment. Midge larvae 

feed on algae, higher plants, and organic detritus (Pennak 1989). Chironomid larvae are ecologically 

relevant for toxicity testing because of their widespread distribution, numerical abundance, and importance 

as prey for juvenile and adult fish. Two species of Chironomus are typically used in toxicity testing; C. 

tentans in North America and C. riparius in Europe. Chironomus tentans eggs hatch about 2 d to 6 d after 

laying and begin the first of four aquatic instar larval stages (summarized in Benoit et al. 1997). The larval 

development lasts about 23 d and is followed by a short (~1–2 d) pupal stage. Males emerge about five 

days before females (protandry). Adults are short-lived, about 7 d. The aquatic development period for C. 

riparius is shorter as adults emerge within about 15–17 d after beginning the first instar stage (Watts and 

Pascoe 1996, Watts et al. 2001a) and live about 4 d (e.g., Ristola et al. 2001). This was confirmed in a 

direct comparison in natural and artificial control sediments between the two species (Watts and Pasco 

2000). 

87. Availability, Culture, and Handling: Both species of Chironomus are readily available from 

commercial sources and do not need to be collected from the field. They are easily maintained in 

laboratory culture. Detailed culture requirements for C. tentans have been described by EPA (2000). 

Larval are held at about 23°C in glass aquaria (e.g., ~19 L volume containing ~8 L water) with finely 

shredded paper toweling or silica sand used as substrate. About 600 eggs per 8 L should be stocked into 

each aquarium. High densities of larvae increase development time, delaying adult emergence. Larval are 

fed a slurry prepared of commercial fish food flakes at a concentration of ~0.04 mg dry food/mL culture 

water. Adult emergence commences in about three weeks after egg hatching (23°C). Chronic tests must 

be initiated with <24-h old larvae, therefore, laboratories must maintain cultures to ensure larvae of the 

correct age are available for testing. 

88. McCahon and Pascoe (1988) presented information on culturing C. riparius. Conditions for 

culture are somewhat similar to those for C. tentans. The substrate is based on homogenized cellulose 

paper (filter paper). Cultures are maintained at 18°C to 20°C and are fed a ration of commercial fish food 

flakes daily. 

89. Strengths and Weaknesses: The principal strengths of using Chironomus in toxicological testing 

are the high ecological relevance, ready availability, and ease of culturing of the two primary species. 

Formal test procedures for both species are well-developed and the species have been used in a variety of 

laboratory (e.g., Benoit et al. 1997, Environment Canada 1997a, US EPA 2000) and in situ (Castro et al. 

2003) testing situations. At least one other species, C. prasinus, has been used in some EDC testing 

(Sánchez and Tarazona 2002). The life cycle of chironomids is relatively short and amenable to 

multigenerational testing. The life cycle of C. tentans is about 33–38 days (Benoit et al. 1997); that of C. 

riparius is likely shorter (Watts & Pascoe 1996, Watts et al. 2001a). Recent studies have included several 

evaluations of potential endocrine disrupting chemicals on C. riparius (Hahn et al. 2001, Watts et al. 

2001a, Hahn et al. 2002, Hahn and Schulz 2002) and C. tentans (Kahl et al. 1997). One possible weakness 

is the apparent lack of response, or conflicting responses, of life cycle endpoints in C. riparius to exposure 

to two known xenoestrogens, bisphenol-A and 17α-ethinylestradiol, or in C. tentans exposed to the 

surfactant 4-nonylphenol (Kahl et al. 1997). Chironomid larvae are sediment dwellers; therefore the 

testing protocol must provide the test animals with a suitable substrate. Watts and Pascoe (2000) showed 

the type of sediment (artificial or natural) was an important contributor to the differences in response that 
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they observed. Concentrations of contaminants were greater in the porewaters of the artificial sediment 

than in those of the natural sediment. 

90. Another important issue that must be considered when using either species of Chironomus is that 

the two were shown to have different sensitivities to the same toxicant. Watts and Pascoe (2000) found 

that C. tentans was more sensitive than C. riparius to the same toxicants evaluated under the same test 

conditions. They also determined the C. tentans was less physically robust than C. riparius, which led to 

increased variability in the data for C. tentans, especially emergence data (there was poor emergence in 

control animals). The response criteria were affected by type of sediment, the choice of species, the 

particular toxicant, and the duration of the experiment. 

4.2 CRUSTACEA 

91. Although insects, crustaceans, and pycnogonids (sea spiders) are the primary arthropod taxa that 

occur in estuarine and marine communities, crustaceans are certainly the more abundant and, therefore, the 

more ecologically important group. Many insect species occur in these waters but they are relatively 

poorly studied. Little is e.g. known about pycnogonid biology and endocrinology. Thus, the focus in this 

section is directed towards the freshwater, estuarine and marine crustacean species that so far have been 

identified as likely to be amenable to studies on development and reproduction as well as specific EDC 

evaluations. 

4.2.1 Amphipods 

Hyalella azteca Saussure 1857

92. Natural History: Historically, Hyalella azteca has been considered a very common inhabitant of 

permanent fresh water systems throughout North, Central, and northern South America (Bousfield 1973, 

Pennak 1989, Gonzalez and Watling 2002). As discussed in the “Strengths and Weaknesses” section 

below, the species’ actual distribution is probably much more restricted. Hyalella inhabits a variety of 

unpolluted springs, streams, brooks, pools, ponds, and lakes (Pennak 1989), but also may occur in large 

coastal rivers and into tidal fresh waters (Bousfield 1973). These epibenthic-burrowing amphipods may 

occur in large numbers; they are detritivorous, feeding primarily on algae and bacteria. Hyalella produces 

about 18 eggs per brood although the number can vary with female age and size (Pennak 1989). At 

warmer temperatures (24°C–28°C), eggs hatch in about 5–10 d and females can produce about 15 broods 

every 150 days (Pennak 1989, EPA 2000). 

93. Availability, Culture, and Handling: Hyalella azteca has been used in toxicity testing for many 

years and is widely available from many commercial suppliers. US EPA (2000) provides a detailed 

description of culture requirements. The amphipods can be cultured in containers of various sizes (e.g., 

2L–80L) at a water temperature of about 23°C and a light:dark photoperiod of 16:8 h. The water should be 

renewed periodically and monitored for various parameters such as hardness, alkalinity, pH, and ammonia. 

Amphipods can be fed commercial fish food flakes, or a combination of yeast-Cerophyl®-trout chow 

(YCT, prepared according to the recipe in US EPA 2000 or purchased commercially) and the alga 

Selenastrum capricornutum. 

94. Strengths and Weaknesses: Hyalella azteca is an ecologically relevant testing organism. It has 

been used in many types of toxicity testing and several protocols, including life cycle testing, to guide 

testing have been written for it (e.g., Environment Canada 1997b, US EPA 2000). This amphipod is 

readily available and relatively easy to culture. However, endocrine disruption evaluations, and any other 

toxicity evaluations, with this species are hampered by recent revelations that the widespread “species” 

Hyalella azteca is actually a mixture of many species. Several studies, beginning in the late 1990s, have 
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investigated various populations of H. azteca from around North America. Duan et al. (1997) studied six 

lab “populations” of H. azteca; one from Burlington, Ontario, one from Conesus Lake NY; and four 

originally derived from the Corvallis, Oregon population [the Nebeker strain that was originally collected 

in 1982 (Duan et al. 2000b)], and another species in the genus, H. montezuma. They found that the four 

populations derived from Oregon were relatively closely related and distinct from the other two and from 

H. montezuma. Duan et al. concluded that the populations represented four species (including H. 

montezuma). McPeek and Wellborn (1998) found allelic and genotypic differences between large-bodied 

and small bodied ecotypes of H. azteca in southeastern Michigan. They determined that the two ecotypes 

did not interbreed and that it was likely that they represented different species. Duan et al. 2000a studied 

12 populations, all presumably H. azteca, from geographically isolated parts of N. America. They used 

electrophoresis to identify nine groups. The large genetic differences that were identified led them to 

conclude that all nine probably represent different species. Witt and Hebert (2000) studied populations 

from wide geographical area in NA, most of which were from around the Great Lakes (24 lakes and ponds 

in Ontario and Wisconsin), but also included the Yukon and New Brunswick. They found low levels of 

gene flow, reduction in genetic variability, low heterozygosity, unique alleles, and strong genetic 

differentiation and divergence among populations, concluding that there were at least seven mitochondrial 

lineages. Gonzalez and Watling (2002) conducted a morphological study of type specimen of H. azteca

and compared it to specimens from Maine, Texas, Mississippi, Michigan, Oklahoma, and Hawaii 

concluding that H. azteca is a species complex. Thus, the taxon currently known as “Hyalella azteca” is 

probably comprised of at least a dozen species. Gonzalez and Watling (2002) recommend that any further 

toxicological studies confirm identification of species. If animals are obtained from a commercial supplier, 

the original source of that supplier’s culture should be identified. 

95. The potential ramifications of using different genetic species on toxicity testing was underscored 

by Duan et al. (2000b, 2000c) who tested the effects of various metals, pH, and fluoranthene on different 

genotypes within a population of H. azteca derived from the Nebeker strain. They discovered that some 

genotypes were more susceptible to the stressors than others, and that the susceptibilities differed across 

genotypes. This differential susceptibility implies that a population that has a higher frequency of a 

particular genotype that happens to be resistant to a particular stressor will show better survival than a 

population that has a high frequency of a genotype that is susceptible to the stressor. That is, different 

genetic populations may provide different responses to the same stressor. 

Gammarus pulex

96. Natural History: Gammarus pulex, an amphipod that is also known as scud, is widespread 

throughout Europe, where it is often the most abundant invertebrate in streams. G. pulex is a primary prey 

item of fish and used in biotic indices to describe pollution in rivers (summarized by McCahon and Pascoe 

1988). The reproductive biology of G. pulex involves precopulatory guarding of the female by a male 

(e.g., Watts et al. 2001b). The male clasps onto a female and swims with her until she molts to assuring 

sexual contact during the time at which mating can occur. This particular behavior has been developed 

into a bioassay procedure (Poulton and Pascoe 1990, Pascoe et al. 1994). It is likely that a chemical signal 

facilitates this behavior, although evidence for pheromonal control of mating in this species is 

contradictory (summarized by Watts et al. 2001b). A similar assay was developed for the estuarine 

gammarid, Gammarus duebeni (Lawrence and Poulter 1996). 

97. Availability, Culture, and Handling: G. pulex is available from at least one commercial supplier, 

but it is not clear if those animals are cultured by the supplier or field collected. Most studies report using 

field-collected animals and hold them for a short time before testing. McCahon and Pascoe (1988) 

developed culture methods for G. pulex. Gravid females and pre-copula pairs are removed from the 

collection and placed in breeding chambers (1-L plastic jars with an open, mesh-covered bottom to allow 

juveniles to escape); these chambers are placed into larger (2–8 L) tanks that are provided with flow-
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through dechlorinated water (the chambers can be maintained static, if water is renewed periodically). 

Adults are fed a variety of conditioned tree leaves (e.g., horse chestnut, elm, sycamore, or oak) that are 

collected during fall, air dried, and stored until just before required for feeding. The leaves are prepared by 

being placed in organically enriched dechlorinated water to promote the bacterial and fungal growth that 

provides an important food for G. pulex. About 200 gravid females in a chamber will provide enough 

juveniles (500–1000) to stock the rearing tank. Juveniles are fed conditioned leaves, which are 

supplemented with adult feces from the adult containers. Timing can be arranged to provide different age 

ranges for testing. Culture temperatures were not provided, but at 13oC about 70% of juveniles reach 

maturity in 130 days, after completing 10 molts. 

98. Strengths and Weaknesses: Gammarus pulex provides a test organism that has considerable 

ecological relevance in European freshwaters. Because organisms are usually field collected, the potential 

for population-related differences in pollutant sensitivities is high and should be assessed before comparing 

the results of tests performed on different populations. However, collected animals are relatively easily 

maintained in culture systems once established and can provide a suitable number of test organisms for 

some time. G. pulex has been used in many types of bioassay procedures, including behavioral (Poulton 

and Pascoe 1990, Pascoe et al. 1994) and in situ (Matthiessen et al. 1995) testing. However, formal 

chronic test protocols have not been developed (Segner et al. 2003). 

Leptocheirus plumulosus:

99. Several species of benthic estuarine and marine amphipods are used regularly in regulatory 

toxicity testing, including Ampelisca abdita, Eohaustorius estuaries, Rhepoxynius abronius, and 

Leptocheirus plumulosus. Test guidelines have been established for these species (e.g., US EPA/USACE 

1998, US EPA 2001). Several other species have been used in studies of the toxic effects of various 

contaminants. Among these are Corophium volutator, Gammarus spp., Microdeutopus gryllotalpa, 

Echinogammarus marinus, and Grandidierella japonica. Among these amphipods, L. plumulosus 

probably is the best suited for adaptation to an EDC testing program and, therefore, is discussed in more 

detail below. 

100. Natural History: Leptocheirus plumulosus is a common infaunal amphipod occurring in estuaries 

along the east coast of the U.S. from northern Florida to Massachusetts (Bousfield 1973). L. plumulosus

lives in simple burrows it digs into the upper layers of the sediment. It is a detrital feeder, extracting 

organic material that is suspended in the water and from ingested sediment. L. plumulosus tolerates a wide 

range of salinities, from 1‰ to 35‰ (US EPA 2001). Its generation time is short, requiring about 24 days 

at 23°C (US EPA 2001). Another species in the genus, L. pinguis, is known to use parental care, sheltering 

juveniles in its burrows (Thiel 1997). L. plumulosus has become increasingly selected as the test species in 

contaminant sediment evaluations. A chronic survival, growth, and reproduction testing protocol for the 

species was recently developed (US EPA 2001). 

101. Availability, Culture, and Handling: L. plumulosus is readily available from several commercial 

suppliers who maintain large cultures of the species. Laboratories can obtain amphipods having a wide 

variety of ages. Laboratories conducting the chronic testing protocol need to culture the amphipods inhouse because the protocol requires that the test be initiated with neonates (<48-h old). Cultures are easy 

to start and are maintained at temperatures of 20°C to 25°C and at the salinity that will be used in testing, 

which is usually 20‰ (US EPA 2001). Cultures are fed finely milled fish food flakes and the culture water 

is renewed three times per week. Cultures may have to be thinned because these amphipods are very 

prolific. Cultures require about 6 weeks maturing sufficiently to provide enough neonates for testing. 

102. Strengths/Weaknesses: The principal strengths of L. plumulosus for EDC evaluations are its 

ready availability, ease of culture, and sensitivity to contaminants. The species is becoming widely used in 
 ENV/JM/MONO(2006)22 

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toxicity testing programs. A 28-day chronic testing protocol, which could be adapted to EDC testing 

programs, has been developed for the species (Emery et al. 1997, US EPA 2001). Some aspects of the 

species’ reproductive endocrinology have been studied (Volz et al. 2002, Block et al. 2003) and the 

impacts of some chemicals known to affect endocrine systems have been tested on the species (Lussier et 

al. 2000, Zulkosky et al. 2002). Spencer and McGee (2001) used data collected from field populations of 

L. plumulosus to develop a stage-structured population model that allows extrapolation from laboratory 

toxicity tests to population impacts. McGee and Spencer (2001) tested the model by using laboratory 

testing data to project the effects of sediment toxicity on the population growth rate. They found that the 

population projections were very similar to amphipod abundances observed at the site from which the 

sediment was collected. L. plumulosus abundance in the area from which the sediment was collected. L. 

plumulosus occurs only along the east coast of the U.S. and, therefore, serves only as a surrogate species 

for estimating the impacts of EDCs on faunas from other regions. 

4.2.2 Daphnids 

103. Natural History: These three species of daphnids are widespread, occurring in freshwater ponds 

or lakes throughout much of the world (Weber 1993). C. dubia is a small species commonly found among 

plants in the littoral zone of lakes and ponds. D. magna lives primarily in lakes where the hardness exceeds 

150 mg/L CaCO3, whereas D. pulex lives in clean, well-oxygenated ponds (Weber 1993). Much of the 

year, when environmental conditions are favorable, populations of these species consist primarily of 

females; males are produced when the environment begins to become unfavorable. Male C. dubia usually 

appear in autumn and those of D. magna in spring or autumn. C. dubia and D. magna reproduce via cyclic 

parthenogenesis with the male contributing genetic material only during certain times of the year. D. pulex

may reproduce by either a cyclic or obligatory parthenogenesis. Daphnids produce ephippia, encased 

embryos that are resistant to harsh environmental conditions, during periods of sexual reproduction. 

104. The life span of these species is highly variable, depending principally on environmental 

conditions. At 25°C, C. dubia and D. magna live about 30 days and 40 days, respectively. D. magna lives 

about 56 days at 20°C, whereas C. dubia and D. pulex live about 50 days at that temperature (Weber 1993). 

These daphnids produce about 10 eggs per clutch, which hatch into juveniles after about 38 h; the 

maximum egg production occurs at temperatures from 18°C to 25°C. 

105. Availability, Culture, and Handling: All three species are readily available from commercial 

sources and are easily maintained in culture. Only 20–30 individuals are required to start a culture within a 

testing laboratory. Cultures should be maintained for at least two generations under the same conditions 

and feeding regimes that will be used during testing. Cultures can be maintained in natural waters, but 

Weber (1993) recommends “synthetic” water because it produces reliable results. Culture temperature is 

optimal at about 20°C. D. magna requires relatively hard water (160–180 mg/L CaCO3), whereas D. pulex

requires softer water (80–90 mg/L CaCO3). Dissolved oxygen concentrations should be held above 5 

mg/L. Cultures should be maintained carefully. The culture medium should be replaced three times per 

week and the cultures should be thinned when the populations exceed 200 individuals per 3 mL of culture 

water. Daphnids should be fed a YCT mixture and the alga Selenastrum capricornutum. 

106. Strengths and Weaknesses: Daphnids offer many advantages to an EDC evaluation testing 

program. They are ecologically relevant, widely distributed, and are readily available via commercial 

suppliers. Cultures are easy to start and maintain. Daphnids are usually tested at intermediate temperatures 

(20°C). Acute and chronic protocols for testing daphnids are available and daphnids are actively used in 

many testing scenarios. Daphnids have been actively used in the evaluations of potential endocrine 

disrupting chemicals (Section 3.2). The primary weakness with using daphnids in endocrine disruption 

evaluations is that the asexually reproducing stage is typically the one used in testing because the presence 

of sexual reproduction is taken as an indication that culture conditions are less than ideal or that testing 
ENV/JM/MONO(2006)22 

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methods are deficient (Olmstead and LeBlanc 2000). Parthenogenic reproduction results in lower genetic 

variability within the test population than would occur with a sexually reproducing taxon (Lagadic and 

Caquet 1998). This reduced variability would reduce the potential variation in response to stressors. 

However, the emphasis on only testing the parthenogenic phase for endocrine disruptive effects ignores the 

sexual reproduction phase, which is vital to daphnid population viability (Lagadic and Caquet 1998, 

Olmstead and LeBlanc 2000). Therefore, EDC impacts to daphnid populations may be severely 

underestimated. The species discussed above are temperate species and may not be feasible for use in 

tropical areas. Buratini et al. (2004) evaluated a tropical species, Daphnia similis, and found it to be 

similar in sensitivity to the two temperate Daphnia species. D. similis is easily cultured and may be more 

appropriate for use in evaluating potential EDCs in tropical waters that have low hardness. 

4.2.3 Copepods 

107. Natural History: Copepods comprise one of the most important groups of marine invertebrates in 

estuarine and marine systems whether planktonic or benthic. They form a vital link in water column food 

webs, feeding on smaller plankton and being consumed by larger predators, such as fish. In benthic 

systems they are often important grazers on microalgae, yet provide forage for young life stages of many 

fish. The life history of copepods involves morphologically distinct stages (nauplius and copepodite 

stages). Therefore, it is possible to test the effects of chemicals, including those that are potential 

endocrine disruptors, on the metamorphosis from naupliar to copepodite stages that might detect 

interference with processes regulated by ecdysteroids (Anderson et al. 2001). 

108. Availability, Culture, and Handling: Some species of copepods, Acartia tonsa for example, may 

be available commercially. However, the species typically used in toxicity testing are easy to maintain in 

cultures so that sufficient animals for testing are always available. Some Acartia tonsa cultures have been 

maintained for at least 20 years (Anderson et al. 2001). Culture temperatures depend on the species 

involved. A. tonsa and Tisbe battagliai are cultured at 20°C (Anderson et al 2001, Hutchinson et al. 1999a, 

1999b), Nitocra spinipes at 22°C (Breitholtz et al. 2003), Amphiascus tenuiremis at 22°C, and Tigriopus 

japonicus at 25°C (Marcial et al. 2003). Copepods are typically fed unicellular algae, such as 

Rhodomonas, Nanochloropsis, and Isochrysis. 

109. Strengths and Weaknesses: Copepods are ecologically relevant animals for use in the evaluation 

of EDCs. They are readily available and easy to maintain and handle. The test conducted by Anderson et 

al. was a 5-6 day subchronic test with a very simple and easy to measure endpoint, the proportion of larvae 

that metamorphose from nauplius to copepodite. A. tonsa was more sensitive than Daphnia magna to 

interruption of growth and molting. Chronic (full life-cycle) tests may be completed within a relatively 

short time, from 21 days to 25 days. Although only one formal chronic test protocol has been developed 

(ASTM E2317-04 with A. tenuiremis), many copepod species have been used in toxicity testing, including 

the evaluation of potential EDCs (Section 3.2). The procedures used for those experiments could be 

adapted for standardized testing worldwide. There is ongoing work to develop an OECD copepod test 

guideline; a draft is written and pre-validation activities have been completed. 

4.2.4 Mysids 

110. Several mysid species are considered below for their potential utility in EDC testing. For each, a 

discussion of natural history, availability, and culture and handling is offered, along with a summary 

paragraph on strengths and weaknesses of the species as test organism. Ingersoll et al. (1999) included 

several additional characteristics that are important considerations in the selection of test species. Some of 

the features, for example mode of reproduction and knowledge of endocrinology, do not allow for 

discrimination among candidate mysid species. [For several of the species in the section below, the name 
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of the researcher who described them is cited. If the species was originally described in a genus other than 

the one listed, the researcher’s name is enclosed within parentheses. 

Americamysis bahia (Molenock, 1969)

111. Americamysis bahia is a small mysid crustacean that occurs in coastal estuaries and embayments 

ranging from the Gulf of Mexico to Narragansett, Rhode Island (Price et al. 1994). 

112. Natural History: A. bahia typically occurs in areas where the salinity is >15‰, but is more 

abundant in higher-salinity habitats. Molenock (1969) originally described the species as Mysidopsis 

bahia, but Price et al. (1994) transferred it to a new genus, Americamysis, during their taxonomic revision. 

The natural history of the species is well known; the following description is based on Weber (1993). 

Adults of A. bahia may reach almost 10 mm in total length, and females attain a larger size than males. 

Individuals become sexually mature in about 12 to 20 days, and the genders can be distinguished when the 

animals reach about 4 mm in total length, at which time the brood pouch typically has started development. 

At about 12 days, the female’s ovaries begin to contain eggs, and the brood pouch is fully formed at about 

15 days. The developing young are carried in the brood pouch for an additional 2 to 5 days, resulting in a 

life cycle of about 17 to 20 days. New broods may be produced about 4 to 7 days. Females produce an 

average of about 11 eggs per brood, and the number is directly related to female body length. 

113. Availability, Culture, and Handling: A. bahia is cultured commercially by many laboratories 

located throughout the United States and therefore is readily available to testing laboratories. With only a 

few days’ notice, commercial suppliers can ship <24-h-old mysids via overnight service, allowing testing 

laboratories time to acclimate the animals to test conditions. However, testing laboratories can also easily 

culture the species. The mysids can be raised in 80-L to 200-L aquaria provided with continuous flowthrough or recirculation systems. The water temperature and salinity within the culture tanks are typically 

24oC to 26oC and 20‰ to 30‰, respectively. The cultures are fed newly hatched brine shrimp (Artemia

sp.). Several sources provide information on culturing this species (e.g., Lussier et al. 1988). Although a 

small species, individuals of A. bahia are relatively easily handled. 

114. Strengths and Weaknesses: As a candidate test species, A. bahia has many strengths and few 

weaknesses. Its primary advantages include its widespread availability and ease of culture. Because 

animals can be obtained from commercial cultures, the likelihood of misidentifying the species is very low. 

Its relatively short generation time makes it desirable in life cycle testing. The species is widely used in 

toxicological testing, and appropriate test conditions are well known. Standardized life cycle test protocols 

have been developed (US EPA 1996; ASTM 1997) and applied (e.g., McKenney 1982, 1985, 1986, 1994; 

McKenney et al. 1991; McKenney and Celestial 1996) or evaluated (Lussier et al. 1999). Included among 

the standardized test protocols are many of those necessary to measure EDC-related endpoints; others 

appear in peer-reviewed publications (Ingersoll et al. 1999). McKenney (1998) synthesized the results of 

several of his earlier studies on the effects of chronic exposures to various pesticides on A. bahia and 

concluded that life-cycle endpoints (slowed juvenile growth and reduced production of young) were more 

sensitive than survival. More recently, studies showed that the results from toxicity tests, which included 

standard and multigenerational tests using A. bahia, could be used to extrapolate from laboratory to 

population effects (Kuhn et al. 2000; 2001). Verslycke et al. (2004a), in their review of the use of mysids 

for endocrine disruptor evaluations, favored A. bahia as a reasonable test species. One criticism of the 

widespread use of A. bahia in toxicological testing is that because it is a warm-temperate or subtropical 

species, it may not be ecologically relevant to colder-water materials testing. 
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Holmesimysis costata (Holmes, 1900)

115. Holmesimysis costata, previously referred to as Acanthomysis sculpta, is an ecologically 

important species that ranges from southern California to British Columbia (Hunt et al. 1997). 

116. Natural History: H. costata is a dominant member of the plankton community living within the 

surface canopy of the giant kelp, Macrocystis pyrifera (Chapman et al. 1995) and is an important pelagic 

prey of the California gray whale, Eschrichtius robustus (Dunham and Duffus 2002). Adults may reach 

lengths of about 7 mm to 13 mm (Daly and Holmquist 1986), and females attain a larger size than males 

(Turpen et al. 1994). Sexual maturity occurs at about 42 days, at which time males, which are recognized 

by an extended fourth pleopod, can be distinguished from females, recognized by the developing brood 

pouch (Turpen et al. 1994). Young are released when the females are about 65 to 73 days old. Brood size 

among laboratory-cultured females averaged about 16 released juveniles per female; in contrast, fieldcollected females released substantially larger broods that averaged 27 young (Turpen et al. 1994). Brood 

size was directly related to female size. US EPA (2002a) reported that females may produce multiple 

broods during their 120-day life span. However, Turpen et al. (1994) also reported that all laboratoryreared females died before releasing a second brood of young. 

117. Availability, Culture, and Handling: Field-collected animals are available from a few suppliers 

and are likely to be available year round, but the species is not cultured commercially (Turpen et al. 1994). 

Brood stock can be collected by sweeping a small-mesh net through the canopy of the giant kelp (US EPA 

2002a). Field-collected mysids can be cultured in the laboratory, and guidelines for doing so have been 

established (Chapman et al. 1995, US EPA 2002a). However, brood stocks should be rejuvenated 

periodically by the addition of field-collected animals (Turpen et al. 1994). Culture tanks can range in 

volume from 4 L to 1000 L and should be provided with aeration and fronds from the giant kelp, M. 

pyrifera (Chapman et al. 1995, US EPA 2002a). H. costata is typically maintained and tested at 

temperatures that range from about 13oC to 15oC for animals collected north or south of Point Conception, 

CA, respectively (US EPA 2002a). Adults are fed newly hatched Artemia, whereas juveniles are fed 

Artemia supplemented with a small amount of ground fish-food flakes (e.g., Tetramin). The animals are 

easily handled by using a combination of small-mesh dip nets and pipettes to transfer them from culture 

tanks to test chambers. 

118. Strengths and Weaknesses: The primary strengths of this species are its ecological relevance to 

northeast Pacific regional testing conditions, its relatively large brood sizes, and its ease of handling and 

maintenance. H. costata occurs in relatively cold waters and may serve as a coldwater alternative to A. 

bahia. The species has been used in several toxicological tests (e.g., Singer et al. 1998) and standardized 

test protocols for the species have been developed (Chapman et al. 1995, US EPA 2002a) and evaluated 

(Martin et al. 1989, Hunt et al. 1997). Martin et al. (1989) determined that H. costata had sufficient 

sensitivity to evaluate effluent toxicity and that the results were repeatable by different testing laboratories. 

The principal disadvantages inherent in using H. costata are its long generation time (~70 days) and the 

difficulty in raising multiple broods in the laboratory. The tests required to measure many EDC-related 

endpoints must be developed (Ingersoll et al. 1999) or are impractical because of the species long 

generation time and the difficulty in raising multiple broods. Because the original animals are fieldcollected, they must be identified carefully prior to their use in testing. Also, potential population-related 

differences in pollutant sensitivities should be assessed before comparing the results of tests performed 

with mysids from different populations. 

Mysidopsis intii holmquist, 1957

119. Mysidopsis intii is an epibenthic species that occurs in the eastern Pacific from South America to 

the southern California coast of the United States (Price et al. 1994; Langdon et al. 1996). 
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120. Natural History: M. intii has only recently been reported from the United States (off Los 

Angeles), but it could be more widespread (Langdon et al. 1996). M. intii is a relatively small species that 

attains body length of about 6 mm to 7 mm. The genders can be distinguished at about 9 to10 days after 

hatching. Eggs enter the brood pouch at about 13 days, and juveniles are released at about Day 20 

(Langdon et al. 1996). 

121. Availability, Culture, and Handling: Animals to be used in testing must be obtained from field 

collections, because there is no commercial culture of the species. Individuals of M. intii are collected by 

using an epibenthic sled, and the wild-caught animals are then used to establish breeding stocks in the 

laboratory. The species is easily cultured in the laboratory in 40-L to 90-L tanks continuously supplied 

with flowing, filtered seawater. Langdon et al. (1996) determined that the optimal temperature for high 

juvenile production is 20oC. To ensure high reproductive output, adults should be fed recently hatched 

Artemia and adult copepods, Tigriopus californicus (Kreeger et al. 1991; Langdon et al. 1996). The 

Artemia diet can be enriched with fatty acid supplements, and the mysid cultures provided with fatty acid 

boosters (UCSC 1998). Separation of juvenile mysids from the adult cultures is easily accomplished by 

using light to attract them through a 1-mm-mesh divider into an isolation chamber (Langdon et al. 1996). 

122. Strengths and Weaknesses: M. intii represents an indigenous, ecologically-relevant species for 

testing contaminants that could negatively affect northeast Pacific coast ecosystems. The life cycle of M. 

intii is much shorter (~20 days) than that of H. costata (~70 days), the other Pacific coast species 

commonly used in toxicity testing. The US EPA sponsored the development of a 7-day toxicity test 

protocol (Langdon et al. 1996) that has been applied (UCSC 1998) and evaluated (Harmon and Langdon 

1996). M. intii was recently used in a series of tests examining acute and chronic effects of nickel on three 

species of marine organisms (Hunt et al. 2002). The chronic test was a 28-d full life-cycle test that 

included survival, growth (as the change in length and weight), and two reproductive endpoints (the 

percentage of females carrying eggs or juveniles in the brood sac, and the number of live juveniles 

produced). However, during this test no juveniles were released and the number of gravid females was low 

in the test controls. Hunt et al. did not discuss possible explanations for the reproductive failures. The 

primary disadvantages associated with using M. intii is the lack of available commercial culture and that 

testing protocols to measure EDC-related endpoints need to be developed. Regardless of source of test 

animals, M. intii individuals used in testing must be identified carefully prior to their use in tests. Further, 

the requirement to supplement an Artemia diet with copepods (T. californicus), which initially must be 

field collected, could be an impediment to the use of M. intii by some laboratories. Finally, potential 

population-related differences in pollutant sensitivities should be assessed before comparing the results of 

tests performed on different populations. 

Neomysis integer (Leach, 1814)

123. Neomysis integer is found throughout northern Europe (Mees et al. 1994) and has been suggested 

as an appropriate species for use in European toxicity testing programs (Roast et al. 1998, 2000b). 

124. Natural History: N. integer is a relatively large, hyperbenthic species that occurs in relatively 

low-salinity portions of estuaries (Roast et al. 2001). Females may attain a standard length of about 18 mm 

(measured from the base of the eyestalk to the end of the last abdominal segment); males are smaller (Mees 

et al. 1994). Brood size is strongly correlated with the size of the female: the number of larvae per brood 

extends to about 80 individuals for females of 16 mm or more in length (Mees et al. 1994). Winkler & 

Greve (2002) determined that individuals matured in about 45 d after hatching at 15°C, but at about 3.5 

months at 10°C; the total generation time was about 69 d at 15°C. 

125. Availability, Culture, and Handling: Because it is the dominant mysid inhabiting northern 

European estuaries (Mees et al. 1994), it is readily available, but animals to be used in testing must be field 
ENV/JM/MONO(2006)22 

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collected. It is not cultured commercially, but wild-collected animals are easily maintained in the 

laboratory. 

126. Strengths and Weaknesses: The principal strength of this species for use in toxicity testing is that 

it is a common and ecologically important component of European estuaries. It is also a very well-studied 

species, ecologically, physiologically, and toxicologically. However, it is not commercially cultured and 

all animals to be used in tests must be collected from estuaries and raised in the testing laboratories. N. 

integer can be tested at relatively cool temperatures of about 15°C, although it is probably not a useful 

species for testing at colder temperatures (~10°C) because of the much longer generation time at cold 

temperatures. Testing to measure potential EDC-related endpoints and studies addressing some aspects of 

the endocrinology and of this species, and its potential utility in EDC testing, has recently been completed. 

Verslycke and Janssen (2002) developed an indicator, the CEA, that could be used to detect changes to 

energy metabolism of N. integer in response to environmental stressors and tested the effects of tributyltin 

chloride on the CEA (Verslycke et al. 2002, 2003a). Verslycke et al. (2002) found that N. integer

produced 11 monohydroxy testosterone metabolites and two nonpolar metabolites (androstenedione and 

dihydrotestosterone). They also found the anabolic steroid β-boldenone, which had not been previously 

reported in invertebrates. The function of the steroids in mysids is not clear. Verslycke et al. (2002) did 

not detect the vertebrate estrogen 17β-estradiol in N. integer. Biotransformation experiments conducted by 

Verslycke et al. (2002) revealed that mysids have a complex steroid hydroxylase system comprised of 

several P450 isozymes. Alteration of P450 activity in mysids could be used as an endpoint in EDC testing. 

Verslycke et al. (2003b) tested the effects of TBTCl on testosterone metabolism in N. integer and found 

non-polar and polar metabolite induction at the lowest TBTCl concentration (10 ng/L) tested, but there 

were no significant differences from controls at higher concentrations. TBTCl had no effect on the 

elimination of testosterone by glucose conjugation, in part because of high within-treatment variability. 

Testosterone elimination by sulfate conjugation was significantly lower at high TBTCl concentrations. 

[Verslycke et al. calculated a metabolic androgenation ratio (the ratio of the oxido-reduced products to the 

hydroxylated plus conjugated products), which has been used to interpret the total effect of a chemical on 

testosterone metabolism, and found no significant treatment-related differences from the control.] Despite 

this, they contended that the ratio could be used to summarize the effect of TBTCl, which was an increase 

in the ratio at the two lowest concentrations. What was also clear in the data, but not discussed, was the 

very high variability in the ratio within treatments, especially within the controls. Control variability, as 

indicated by coefficient of variation values of almost 100%, was very high, which calls into question the 

general utility of the metric in evaluating endocrine disruption.] 

127. Gorokhova (2002) stated that to understand growth of crustaceans (which is a potential EDC 

endpoint) it is imperative to understand the molt staging of the organism. Gorokhova determined that the 

duration of the molt cycle of N. integer was about 9 days and that the effects of salinity and temperature on 

molt staging differed according to the food supply. 

128. These animals must be identified carefully prior to testing. Also, potential population-related 

differences in pollutant sensitivities should be assessed before comparing the results of tests performed on 

different populations. 

Other mysid species

129. A few other species of Americamysis, including A. almyra (Bowman 1964) and A. bigelowi (W. 

Tattersall 1926), have been used in toxicity testing or related studies. A. almyra is closely related to A. 

bahia, and the two species have similar geographic distributions; however, A. almyra inhabits less saline 

waters (Price et al. 1994). The reproductive biology including brood size and generation time of the 

species is very similar to that of A. bahia (reviewed in Reitsema and Neff 1980). The species is amenable 

to laboratory culture. Recirculation (Reitsema and Neff 1980) and static (Domingues et al. 1998, 1999) 
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culture systems have been developed. A. almyra can be maintained on a diet of Artemia nauplii 

(Domingues et al. 2001a, 2001b). A. bigelowi is also biologically similar to A. bahia and to A. almyra. It 

occurs along the east coast of the United States from Massachusetts to Florida (Price et al. 1994). 

Although Gentile et al. (1982) found it to be suitable for use in toxicology testing; A. bigelowi has not 

received widespread attention as a test species. 

130. Neomysis mercedis Holmes 1897, an ecologically important Pacific coast species, was advocated 

as an acute toxicity test organism appropriate for estuarine waters having low salinities, ranging from 1‰ 

to 3 ‰ (Brandt et al. 1993). The suggested temperature range for testing with this species is 16°C to 19°C 

(Brandt et al. 1993); thus it could serve as a cool water alternative to A. bahia. It occurs in freshwater and 

brackish waters from California to southern Alaska (Daly and Holmquist 1986) and recently has been used 

in toxicity testing (Farrell et al. 1998a, 1998b; Hunt et al. 1999, 2002). N. mercedis is an important 

predator on Daphnia in a freshwater lake (Lake Washington, Washington), where its consumption is 

enough to control Daphnia populations (Chigbu 2004). Brandt et al. (1993) found that N. mercedis was 

similar to A. bahia in sensitivity to several environmental contaminants. A standardized acute toxicity 

testing protocol has been developed for N. mercedis (ASTM 1997). N. mercedis is not cultured 

commercially, but can be reared in the laboratory (Brandt et al. 1993). N. mercedis has a relatively long 

generation time of about three to four months (Brandt et al. 1993). The tests required to measure many 

EDC-related endpoints must be developed (Ingersoll et al. 1999) or are impractical because of the species’ 

long generation time. 

131. Neomysis americana (S.I. Smith, 1874) is a western Atlantic species that occurs on sandy 

bottoms at depths of 0 to –240 m from Florida to Newfoundland, and also South America (Anderson et al. 

2004). N. americana plays a dual role in marine food webs, being an important item in the diets of several 

species of demersal fish (Steimle et al. 2000) and a significant predator on zooplankton (Winkler et al. 

2003). The species was used in some toxicity tests in the late 1970s, early 1980s, but has been used very 

little. The US EPA lists N. americana as an alternative species for an acute effluent testing protocol (US 

EPA 2002a), but not for chronic effluent testing (US EPA 2002b). Typical test temperatures are 20°C and 

25°C and the typical salinity range is from 10‰ to 32‰ (US EPA 2002a). 

132. Praunus flexuosus (Müller 1776) is one of the two predominant mysid species in northern 

European waters, particularly in shallow water in the outer Elbe Estuary, where it is predominant in 

summer (N. integer is other abundant species) (Winkler and Greve 2002). P. flexuosus is tolerant of a wide 

range of salinities and temperatures. It is not commercially cultured and must be collected from coastal 

waters before testing. Two studies by Garnacho and coworkers have highlighted one of the important 

caveats that must be recognized when using wild-collected animals for EDC evaluations. Garnacho et al. 

(2000, 2001) examined the effects of copper on survival and the metabolism of P. flexuosus. The mysids 

were fed <48-h old Artemia nauplii during the holding period and daily during testing. Testing was 

conducted with water at a salinity of 33‰ and temperatures of 10°C and 20°C. Water and test solutions 

were renewed every 48 h. Copper was found to be toxic to the mysids collected and tested during the 

summer, whereas it was not toxic to mysids collected and tested during the winter (Garnacho et al. 2000). 

Garnacho et al. found that the O:N ratio varied seasonally, being 2.5 times greater in winter than in spring. 

Metabolism of species is protein based all year. Garnacho et al. also found that the O:N ratio decreased 

quickly (24 h) at the highest copper dose and decreased in all doses by 10 days. The response in summer 

was faster than that in winter. The shift in O:N ratio reflects a greater metabolic reliance on protein. The 

faster response in summer occurs when metabolism is already more dependent on protein. The observed 

response to copper differs from other mysid responses to stress. Garnacho et al. concluded that O:N ratio 

changes appears to be a sensitive early indicator of stress in mysids. However, one significant finding 

from the two studies was that the effects of a particular toxicant or stressor on animals that are wildcollected might vary importantly depending on the season during which the animals were collected. 

Winkler and Greve (2002) determined that P. flexuosus did not mature at a test temperature of 10°C, but 
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did grow to large size (20 mm). At 15°C, individuals matured at about 3.5 months after hatching and 

reached a mean size of 16 to 18 mm. The mean incubation time within the marsupium was about 23 d at 

15°C, and the total generation time (from egg to mature individual) was about 133 d. Despite its 

ecological importance, P. flexuosus would not be a useful colder water test animal, and its long generation 

time makes it undesirable for EDC evaluations. 

133. Tenagomysis novaezealandiae Thomson 1900 is found on New Zealand’s North and South 

Islands (Nipper and Williams 1997). The species is abundant and widely occurring. Although it is not 

available through commercial culture, field-collected animals can be maintained in the laboratory. The 

species’ life cycle lasts about four weeks and laboratory populations can be reproductively active yearround (Nipper and Williams 1997). Mysids can be collected by using hand dip nets and are cultured in the 

laboratory at a temperature of 20°C, a salinity of 34 ‰, and a light:dark photoperiod of 16:8 hours. 

Mysids are held at a density of 10–20 individuals/L and are fed newly hatched brine shrimp nauplii daily. 

Nipper and Williams (1997) conducted several experiments designed to determine the appropriate physical 

conditions and food regimen for holding and testing the species. One of the more interesting findings from 

these experiments was that survival was higher when tests were conducted in complete darkness compared 

to the 16:8-h light:dark cycle. Nipper and Williams attributed this to the typical vertical migration habits 

of mysids, which involves migration up in the water column at night and resting in dark areas near the 

bottom during the day. Nipper and Williams (1997) concluded that T. novaezealandiae is suitable for use 

in toxicity testing and can be cultured for year-round testing. 

4.2.5 Decapods (shrimp, crabs, crayfish, lobsters) 

134. Decapods probably have been the subjects of more endocrine function and possible EDC effectsrelated testing than any other crustacean group (Fingerman et al. 1998, Ingersoll et al. 1999). Decapods are 

particularly appealing to EDC studies because much is known about their endocrine systems and their freeswimming larvae are likely to be susceptible to JH analogs and other Insect Growth Regulators (IGRs) 

produced to control insects (McKenney 1999). 

135. Two species, the mud crab (Rhithropanopeus harrisii) and the grass shrimp (Palaemonetes 

pugio), have been the primary subjects for many of the endocrine studies and have been advocated as being 

potentially useful in studies of EDCs (McKenney 1999). Several studies have investigated the effects of 

EDCs on crustacean life cycles using decapod larvae as the test organisms (Section 3.2). Most of the 

recent work on the impacts of various EDCs on decapods has centered on the grass shrimp Palaemonetes 

pugio and the mud crab Rhithropanopeus harrisii, which, therefore, are the focus of this section. 

136. Natural History: The grass shrimp, Palaemonetes pugio, is widespread in estuaries from Nova 

Scotia to Texas (Williams 1984) and has been used considerably in bioassays. Grass shrimp feed on 

bacteria-laden benthic detritus and are an extremely important component of energetic flux pathways in 

estuaries (studies summarized in Williams 1984). Gravid females are available from late spring to late 

summer. Grass shrimp usually mature about six months after hatching, although maturation may occur 

quicker where water temperatures are warm (Williams 1984, Volz et al. 2002). The mud crab, 

Rhithropanopeus harrisii, occurs in estuaries from the southwestern Gulf of St. Lawrence, Canada to 

Veracruz, Mexico and has been introduced into parts of Europe and the U.S. west coast (Williams 1984). 

Mud crabs tolerate a wide range of salinities and temperatures. Muds crabs are omnivorous, feeding on a 

variety of aquatic plants, detritus, and other estuarine animals. Gravid females typically can be collected 

during the summer months. Mud crab larvae, rather than adults, are typically used in toxicity studies. 

137. Availability, Culture, and Handling: Grass shrimp are available from commercial suppliers and 

can be maintained in culture by testing laboratories. Mud crabs must be collected from the field before 

testing. Grass shrimp can be held at a wide range of conditions with temperatures ranging from 18°C to 
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25°C and salinities ranging from 20‰ to 28‰. Ovigerous females are held in the laboratory until the eggs 

hatch. Larvae, which are used in toxicity testing, can be cultured at the above conditions in the laboratory. 

Many studies involving the mud crab use the larval stages rather than adults. Ovigerous females can be 

collected and held in cultures until the larvae, which are relatively easy to raise, are released. Larvae can 

be cultured at a water temperature of about 25°C and a salinity of about 20‰. Larvae of both species are 

fed brine shrimp (Artemia) nauplii daily. 

138. Strengths/Weaknesses: The principal strength that both species offer an EDC evaluation program 

is that the knowledge of decapod endocrine systems is relatively well-developed. Both are common 

animals in estuaries of the U.S. Gulf and East Coasts. One species, the grass shrimp, is commercially 

available. Both are amenable to individual, focused studies of the specific impacts of EDCs on various 

endocrine system components and, thus, are important species to study. However, neither is advantageous 

for use in an overall EDC screening program, primarily because the long generation time precludes 

relatively rapid multigenerational studies. Additionally, the mud crab must be collected from the field and 

would be difficult for many testing laboratories to obtain. 

4.3 NON-ARTHROPOD SPECIES 

139. Estuarine and marine communities are comprised of taxa from many phyla other than the 

Arthropoda. Some taxa, especially polychaete worms, are of very high ecological importance and are 

probably very susceptible the EDCs. Many frequently are used for various types of regulatory toxicity 

testing. However, the endocrine systems of most have not been studied sufficiently, nor have had adequate 

protocols been developed, for these taxa to be included in an EDC evaluation program. Nonetheless, three 

of these groups of taxa are briefly reviewed in this section. 

140. Polychaetes: Polychaete worms comprise an important component of most estuarine and marine 

ecosystems. Primarily infaunal sediment dwellers, polychaetes contribute significantly to sediment 

bioirrigation and occupy a key part of marine food webs. They represent an ecologically significant group 

on which to evaluate the effects of potential EDCs. Several species of polychaetes are widely used in 

sediment toxicity testing. Polychaete (e.g., Nereis) bioaccumulation assays are one of the primary tools 

used in evaluating the suitability of dredged material for disposal into U.S. coastal waters (US 

EPA/USACE 1998 ITM). A general toxicity testing protocol for has been developed for several species of 

nereid polychaete worms, including a chronic survival and growth assay for Neanthes arenaceodentata

(ASTM 2000) and a fertilization/embryo-larval development assay for Platynereis dumerilii (Hutchinson 

et al. 1995). P. dumerilii has been widely used in Europe, especially in genotoxicity studies (e.g., Haggar 

et al. 2002) and some aspects of its reproductive endocrinology have been studied (review in Andries 

2001). Several age categories of N. arenaceodentata are available from a commercial supplier. Several 

breeding cultures of P. dumerilii have been established at European laboratories (e.g., the European 

Molecular Biology Laboratory in Heidelberg, Germany). The species can be bred continuously in the 

laboratory (Fischer and Dorresteijn 2004) and may eventually be a good candidate for EDC evaluations. 

The prime weakness in using polychaetes for EDC evaluations probably is the lack of understanding of 

functional roles for hormones in polychaetes (Andries 2001). 

141. Mollusks: Mollusks may seem an obvious choice for use in EC evaluations because imposex is 

frequently offered as the best example of endocrine disruption in invertebrates (Section 3.2). Bivalve (e.g., 

Macoma) bioaccumulation studies are important in contaminated-sediment evaluations (US EPA/USACE 

1998), general chronic assays have not been developed. Endocrine disruption in marine gastropods has 

been studied considerably, almost exclusively with respect to the induction and incidence of imposex. 

However, the effects, including reproductive endpoints, of some endocrine disrupting chemicals on the 

hermaphroditic freshwater snail Lymnaea stagnalis have been studied (Czech et al. 2001, Coeurdassier et 

al. 2004). Lymnaea peregra was included in a multi-species multi-generation, whole life-cycle testing 
ENV/JM/MONO(2006)22 

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protocol developed by Sánchez and Tarazona (2002). However, at this point suitable chronic protocols for 

estuarine and marine mollusks have not been developed that could be adapted for a chronic reproductive 

assay. 

142. Echinoderms: Echinoderms are often included in some regulatory testing protocols, particularly 

those involving effluent discharges or the effects of the suspended particulate phase of sediments, but these 

focus on acute effects on embryonic and larval development (e.g., US EPA 2002b, 2002c). Some 

echinoderms have been used in studies of specific EDCs (Section 3.2), but these do not involve chronic 

procedures or reproductive endpoints. Echinoderms would be of interest to an EDC evaluation program 

because they are known to have steroids that also occur in mammals (LeBlanc et al. 1999). Currently, 

echinoderms do not seem well-suitable to multi-generational EDC evaluations. 
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5.0 SELECTION OF AN APPROPRIATE TEST SPECIES AND PROTOCOLS 

143. Several characteristics should be considered in determining the most appropriate species for use 

in EDC-related testing. Nimmo and Hamaker (1982) and Roast et al. (1998) suggested that the test 

organism should be ecologically relevant, and sensitive to contaminants. Because it must be available as 

needed for testing, it must be abundant and easily collected, or amenable to laboratory culture. Its diet 

should be well understood and easy to provide in the laboratory, therefore allowing it to be well-adapted to 

laboratory conditions and reducing the need for extensive acclimation periods. The ideal test organism 

should have a short, relatively simple life cycle that allows for the testing of successive generations. 

144. Ingersoll et al. (1999) included several additional characteristics that are important considerations 

in the selection of test species. These characteristics include the species’ mode of reproduction and 

knowledge of their endocrinology. Species that reproduce parthenogenetically may produce a test 

population within which genetic variability is relatively low compared to species that reproduce sexually. 

High genetic variability may yield highly variable responses to a stressor within a test population (Lagadic 

and Caquet 1998), which may make it difficult to detect impacts other than those that are substantial. 

However, using parthenogenetically reproducing animals means that important processes related to sexual 

reproduction (e.g., gametogenesis) are not being evaluated (Lagadic and Caquet 1998, Olmstead and 

LeBlanc 2000). Knowledge of the endocrinology of a candidate species is important to the selection of 

appropriate test endpoints and the interpretation of the test results (Segner et al. 2003). Also important to 

selecting a candidate species for use in the multigenerational EDC testing program are the ability to culture 

the species in the laboratory (with a strong probability that transgenerational testing is possible); relatively 

short generation time, allowing for full life-cycle testing; size (larger animals provide more tissue for 

measuring hormone titers, but usually have longer generation times); and the availability of standard 

(consensus-based) testing methods, including whether or not new methods must be developed to measure 

EDC-relevant endpoints. 

145. Recent studies of the genetic variability within a common and widespread species emphasizes the 

need to document the test species identity and the source from which the animals originated. Several 

studies have shown that the freshwater amphipod Hyalella azteca is a species complex comprised of at 

least a dozen species and that different genotypes may vary in their response to the same stressors 

(discussed in detail in section 4.2.1). The identity of test organisms should be verified and documented 

through a voucher collection, as described by Huber (1998) that is maintained by the testing laboratory. 

5.1 CRUSTACEANS AS REPRESENTATIVE AQUATIC INVERTEBRATES 

146. Many insecticides are considered to be EDCs, because they are specifically formulated to attack 

insect endocrine systems, affecting in particular the systems that are involved in molting and larval 

metamorphosis (Oberdörster and Cheek 2001). Most of these insecticides are JH analogs (Oberdörster and 

Cheek 2001) or ecdysteroids. Although insects and crustaceans represent two classes within Phylum 

Arthropoda and consequently exhibit many similarities as well as differences, several recent studies 

showed that insecticides formulated as JH analogs adversely affect crustacean larvae by disrupting molting 

and metamorphosis (e.g., see McKenney and Celestial 1996, which focuses on mysids). Crustaceans 

probably do not synthesize JH (LeBlanc et al. 1999); however, they do produce methyl farnesoate in the 

mandibular gland, and it is likely a natural JH analog. It is known to be involved in crustacean 

reproduction (LeBlanc et al. 1999) and development (Borst et al. 1987), but its specific role is uncertain. 

147. Chronic testing protocols for freshwater and estuarine or marine arthropod species have been 

developed by several regulatory agencies. Examples of protocols developed by national regulatory 
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agencies and by OECD on daphnids, on chironomids, on other sediment dwelling invertebrates, on mysids 

and on copepods are highlighted below. Also, current OECD or draft OECD guidelines for each group of 

species are presented in the annex section. 

5.1.1 Mysids as representative crustaceans 

148. Mysid crustaceans are distributed from 80 oN to 80 oS and occur in most aquatic environments, 

including brackish, freshwater, and marine (e.g., Mauchline 1980). In some habitats, particularly coastal 

temperate waters, mysids are very abundant and are very important in freshwater, estuarine, and marine 

food webs. Mysids diets include consumption of detritus, phytoplankton, and zooplankton. They may be 

abundant enough to control population densities of some prey (Chigbu 2004). Stomach-content analysis 

revealed that mysids are a staple food for striped bass, the tidewater silverside, and several species of 

flounder (Gentile et al. 1983). Many estuarine mysids are also hyperbenthic (Roast et al. 1998) and make 

diurnal migrations into the water column (Dauvin et al. 1994). For these reasons, mysids could serve as 

sensitive indicator species to monitor the effects of EDCs through exposure to either the water column or 

sediment or through dietary uptake.

149. Mysids, despite their superficial resemblance to decapod shrimp, have been considered more 

closely related to amphipods and isopods (Brusca and Brusca 1990). However, recent studies of 28S 

rDNA sequences among several malacostracan orders indicated that mysids (Suborder Mysida) are more 

closely related to krill (Euphausiacea) than to amphipods or isopods (Jarman et al. 2000). This proposed 

phylogeny has not been accepted universally (Richter and Scholtz 2001). Regardless of phylogenetic 

relationships, mysids, amphipods, and isopods are characterized in part by the retention of developing 

young in a marsupial brood pouch. All three taxa would be good candidates for toxicological testing, and 

amphipods and mysids are routinely used. However, for EDC testing, especially for life cycle tests, 

mysids offer clear advantages over amphipods. Most marine amphipods used in toxicological testing must 

be collected from their natural habitats just prior to use in tests. Although they can be held for a few weeks 

prior to testing, they generally are not cultured for tests. Currently, only one marine amphipod, 

Leptocheirus plumulosus, has been cultured successfully and used for growth and reproduction tests (US 

EPA 2001). Conversely, several mysid species have been cultured in the laboratory and used in such life 

cycle tests. 

150. Nimmo and Hamaker (1982) advocated the use of mysids in toxicity testing, including acute and 

chronic life-cycle testing. Nimmo and Hamaker compared acute and chronic toxicity data for a variety of 

taxa and concluded that mysids were more sensitive than most of the taxa, including grass shrimp, an 

estuarine fish, and 17 freshwater species. Nimmo and Hamaker foresaw a role for mysids in the evaluation 

of the effects of pesticides and metals on successive generations, in part because the life cycle of mysids is 

short. This role is being fulfilled by the inclusion of mysids in an EDC evaluation program. Mysids are 

currently also used in routine toxicity tests to examine the potential toxicity of marine sediment (Carr et al. 

1998; Cripe et al. 2000). It has been documented (Cripe et al. 2000) that the mysids were observed to 

collect sediment, manipulate it at the mouth region, and drop it. This suggests that mysids could be used to 

test sediment suspected of containing EDCs. Further, Verslycke et al. (2004a) evaluated the potential for 

using mysids in an endocrine disruption evaluation program. They reviewed aspects of the biology and 

ecology of mysids in the context of this potential use and summarized the literature on toxicity studies 

done using mysids. Their review also included a discussion of many endpoints that may be useful when 

using mysids to test for endocrine disruption. Verslycke et al. (2004a) concluded that mysids have the 

ecological relevance and sensitivity to stressors required of a taxon that would be suitable for evaluation of 

endocrine disruption in marine and estuarine invertebrates and could serve as a surrogate for other 

crustacean species. 
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5.1.1.1 Protocols using mysid species 

151. Specific guidance for conducting short-term toxicity tests with species other than A. bahia has 

been published. It is possible that these protocols can be modified to allow longer life cycle testing. 

152. Holmesimysis costata.—Chapman et al. (1995) described a 7-day test protocol designed to 

measure growth and survival in tests using the west coast mysid species, H. costata. In addition to 

recommended test conditions, guidance in culturing the animals and analyzing the data are presented. The 

protocol describes the ecological and culture requirements for H. costata; this information could be used to 

modify the ASTM and OPPTS protocols described above to allow longer life cycle testing. The test 

protocol was evaluated by means of a series of intra- and interlaboratory comparisons (Hunt et al. 1997), 

which concluded that this test had sufficient sensitivity and precision to make it useful in testing possible 

contaminant impacts. 

153. Mysidopsis intii.—A short-term toxicity test protocol for a west coast species, M. intii, was 

developed with support from the US EPA (Langdon et al. 1996). The protocol concisely describes the test 

conditions required to conduct a 7-day toxicity test to measure survival and growth of this species. Initial 

test development was performed using zinc sulfate as the toxicant. The test protocol was evaluated by 

means of an interlaboratory comparison that employed sodium dodecyl sulfate as the toxicant (Harmon and 

Langdon 1996). Harmon and Langdon (1996) also compared M. intii test with those using A. bahia and H. 

costata, and reported that its sensitivity was equal to that of the A. bahia test, but that it was lower than that 

of the H. costata test. 

154. A proposed guideline submitted to OECD for consideration for further development, validation, 

and acceptance has been developed (OECD 2004a, see also Annex 1). The objective of a mysid twogeneration reproductive and developmental toxicity test is to provide the most precise and accurate 

estimate of toxicity associated with endocrine disruption and reproductive fitness for an identified potential 

EDC. The results of the Tier 2 testing should be conclusive in documenting a discernible cause-and-effect 

relationship of chemical exposure to measurable manifestation in the test organisms. The test protocol will 

be designed to: 

• Determine whether effects are a primary or secondary disturbance of endocrine function 

• Establish exposure/concentrations/timing and effects relationships 

• Be sensitive and specific 

• Assess relevant endpoints 

• Include a dose range for full characterization of effects (EDSTAC 1998). 

155. The test is an extension of previous mysid life-cycle test procedures produced by ASTM (ASTM 

1997; see Table 3) and US EPA (Nimmo et al. 1977, 1978, McKenney 1986, 1998; see Table 3). Basic 

principles of experimental design from these two protocols must however be followed. The protocol 

provides general guidance for conducting a two-generation toxicity test in which the parent generation (F0) 

is exposed to the toxicant and the parent and first offspring (F1) generations are monitored for 7 days after 

the mean date of release of the second brood by the control treatment parent generation (F0). The 

guideline also allows for the exposure of the F1 generation to the test substance. The guideline is 

applicable to Americamysis bahia, but could be modified to suit other mysid species. The two-generation 

test is initiated when healthy < 24-h-old mysid juveniles are placed randomly into replicate test chambers. 

These original juveniles comprise the parent (F0) mysid stock for the test. Development, sexual 

maturation, reproduction, and growth are observed in the F0 and F1 generations. The assay is conducted 
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with at least five toxicant concentration treatments and appropriate control treatments (typically a 0% 

concentration and a solvent control, if one was used to deliver the toxicant to the test treatments). The 

highest exposure concentration should be equal to the lowest concentration that caused adverse effects in 

the acute test or 1/10 the LC50 (EC50). Toxicant and control treatments are delivered to the mysids in 

water, which is continuously or intermittently delivered to the test chambers via a proportional diluter 

system or infusion pumps. The specific exposure duration will vary, but is at least 7 days longer than the 

median second-brood-release date by the original parent mysids in the control treatments. During the test, 

chambers are examined for mysid mortality, the presences of molted exuviae, the presence of ovigerous 

females, and the release of young, all of which are recorded. Young from the first brood (= F1') release by 

the parent stock are held for four days, after which they are counted and measured. The parent stock is 

allowed to produce a second brood (= F1"), after which the parent mysids are counted and measured. The 

F1" mysids are exposed to the test materials and maintained until their first brood is released (= F2), after 

which the F1" mysids are counted and measured. The test is terminated with the release of the F2 mysids, 

which are counted. Surviving organisms may be analyzed biochemically, as appropriate. 

5.1.2 Daphnia as representative crustaceans 

156. Daphnia is a genus of small cladoceran crustaceans, most species of which inhabit lakes and 

ponds as a major constituent of zooplankton communities. Using pairs of legs with setae (hairs), they feed 

on protozoa, algae, bacteria and organic detritus. They are often positioned as key species in aquatic 

ecosystems, both as grazers, which feeds on primary producers (algae), and as a major part of the diet of 

fish and invertebrate predators. Most cladocerans exhibit cyclical parthenogenesis with both asexual and 

sexual reproduction existing in the life cycle. They produce females by parthenogenesis as long as 

environmental conditions are favorable. Males appear and sexual reproduction occurs, triggered by the 

change in environmental conditions, such as day length, food concentration and population density 

(Hobaek and Larsson 1990, Kleiven et al. 1992). Males may also be produced parthenogenetically and are 

genetically identical to females as a consequence of the environmental sex determination (Hebert 1987). 

Sexual reproduction occurs by the mating between males and females having haploid eggs. Fertilized eggs 

develop into resting eggs, which are enclosed with protective membranes and are resistant to desiccation 

and freezing. The production of resting eggs enables daphnids to be dispersed to other lakes or ponds by 

their attachment to waterfowl and aquatic insects. The eggs can survive for years preserved in bottom mud 

as “egg banks” analogous to the seed banks of many terrestrial plants (Marcus et al. 1994). 

157. With its short generation time and ease of handling in the laboratory, daphnids have become one 

of the most popular organisms in ecotoxicological research and testing. There are several methods for the 

regulatory reproductive toxicity testing of daphnids (e.g., Environment Canada 1992, OECD 1998a, US 

EPA 2002b, National Institute of Technology and Evaluation, Japan 2004, see also below for details). It 

might be claimed that the particular nature of the reproductive system of cladocerans and the lack of 

pertinent endocrinological information may render them unsuitable organisms for testing the effects of 

potential endocrine disrupting chemicals on crustaceans in general. Recent studies, however, revealed that 

exposure to juvenile hormones in insects and crustaceans, or to pesticides formulated as juvenile hormone 

mimics, induce D. magna to produce male neonates, and that juvenile hormones and their analogs are 

involved in sex determination (Olmstead and LeBlanc 2002, 2003, Tatarazako et al. 2003, Oda et al. 

2005b). Ten chemicals are known to induce the production of male neonates in D. magna (Tatarazako et 

al. 2003, Oda et al. 2005b). The same phenomenon has been reported for other cladoceran taxonomic 

groups (Oda et al. 2005a, Kim et al. 2006). Although the mechanism is not yet clear and the evidence is 

only circumstantial, it is suggested that juvenile hormones are involved in the determination of offspring 

sex by mothers or in the shift of reproductive mode from parthenogenesis to sexual reproduction in 

cladocerans that exhibit cyclical parthenogenesis. Ecdysteroids are also known to inhibit ecdysis or to 

cause an abnormal ecdysis in D. magna (Baldwin et al. 2001). This empirical evidence supports the 
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suitability of using daphnia methods for testing chemicals with possible endocrine disrupting effects on 

crustaceans. 

5.1.2.1. Protocols using daphnid species 

158. US EPA (2002a) developed a chronic protocol for use in the National Pollutant Discharge 

Elimination System (NPDES) Permits Program to identify effluents and receiving waters containing toxic 

materials in chronically toxic concentrations. The method includes chronic testing procedures for a 

daphnid (Ceriodaphnia dubia) that, although not designed specifically for EDC testing, could be modified 

use in such a program. The method also may be adapted for other daphnid species, such as Daphnia major

and D. pulex. Test conditions and endpoints from the US EPA protocol are presented in Table 5. The test 

is continued until 60% of the control animals produce three broods of young. If this does not occur by 8 

days, the test must be repeated. Environment Canada (1992) also has developed a similar chronic protocol 

for C. dubia that includes reproductive endpoints. 

159. OECD revised its original daphnid test procedure (TG 201), separating the acute and chronic 

components into two guideline documents. The new chronic guideline (TG 211; see Table 4), which was 

written specifically to test Daphnia magna, was adopted in September 1998 (OECD 1998). OECD 

recommends that the data from the acute test be available for use in determining the appropriate test 

substance concentrations for the chronic evaluation. Newly hatched daphnids (<24-h old) are exposed to 

the test substance for 21 days. The primary endpoint measured is the total number of living young 

produced per adult daphnid that is still alive at the termination of the test. Any young produced by adults 

that die during the test are not included in the endpoint calculation. Secondary endpoints may include 

adult survival, growth, and the time to first brood. Japan (OECD 2003b) has proposed a draft enhancement 

to TG 211 that includes endpoints specifically intended to detect endocrine disruption in Daphnia. These 

are the offspring sex ratio, which is used as an endpoint for juvenile hormone-like chemicals; and molt 

inhibition, which is used as an endpoint for molting hormone-like chemicals. These endpoints are based 

on data reported by Baldwin et al. (2001), Olmstead and LeBlanc (2002, 2003), and Tatarazako et al. 

(2003). See Annex 2 for the OECD technical guidance Document No 211. 

160. ASTM reapproved its standard guide for conducting life-cycle assays using Daphnia magna

(ASTM 2004a). Newly hatched daphnids (<24-h old) are exposed to the test substance for 21 days. 

Biological data collected include mortality (recorded daily), the number of young produced (determined 

three times per week), size (dry weight) of the first-generation individuals still living at the end of the test, 

time to first reproduction, and behavioral abnormalities. Second-generation daphnid responses, such as 

survival, development, and behavior, may be obtained by observing these daphnids for an additional four 

days or more. 

5.1.3 Harpacticoid and calanoid copepods as representative crustaceans 

161. Copepods occur widely in almost every marine, brackish and freshwater ecosystem. They 

represent important prey items for the larvae and juveniles of many fish and larger invertebrates, and are 

increasingly used as a live food source in aquaculture. A number of copepod species (predominantly 

marine and brackish water species) have been used for many years to evaluate the acute and chronic 

toxicity of single chemicals and complex mixtures in water and sediments (Chandler and Green 2001). 

Copepod development is characterized by the accomplishment of eleven molt cycles. All together the 

juvenile copepods pass through six naupliar and five copepodite stages before turning into the adult (12th) 

stage. Upon hatching from the egg, the unsegmented first nauplius stage already possesses three pairs of 

appendages. At each successive molt these appendages develop by the addition of further setae (hair-like 

processes on limbs and mouth parts) and/or segments. Rudimentary forms of other appendages and 

additional body segments (somites) also develop during the nauplius (larval) stages. During the copepodite 
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(juvenile) stages a new somite is added at each successive molt. The adult copepod bears six pairs of 

appendages and consists of ten somites (Huys et al. 1996). The time to develop from hatch to adult as well 

as the generation time (hatch to hatch) varies with temperature, food-availability, population density, etc.

In general, however, time from hatch to adulthood takes 10-20 days and the generation time is 15-40 days 

at 15-25 oC. The shape of the egg, presence and number of egg sacs, the number of eggs within each egg 

sac as well as the number of broods (from a single fertilization) differs widely between species (Hicks and 

Coull 1983). 

162. Harpacticoid copepods comprise more than 3.000 species, are mostly free-living benthic 

organisms and are usually the second most abundant group of animals (after nematodes) in marine benthic 

communities (Huys et al. 1996). They almost always dominate the gut contents of bottom or phytal feeding 

larval and juvenile fish (Hicks and Coull 1983). Several authors have described toxic effects of single 

substances, oils, complex effluents etc. on harpacticoid copepods, such as Amphiascus tenuiremis

(Chandler et al. 2004a, b; Bejarano and Chandler 2003), Tisbe battagliai (Hutchinson et al., 1999a, b),

Nitocra spinipes (Bengtsson 1978; Renberg et al. 1980; Tarkpea et al. 1999; Breitholtz and Bengtsson 

2001; Breitholtz and Wollenberger 2003), Nitocra affinis (Ustach 1979) and Tigriopus japonicus

(D’Agostino and Finney 1974). 

163. Zooplankton communities in the Baltic Sea and other waters are often dominated by calanoid 

copepods, such as Eurytemora spp. and Acartia spp. (Voipio, 1981), which provide an important link 

between phytoplankton and higher communities (e.g. fish) in the pelagic food chain (e.g. Størrup et al., 

1986). For instance, Acartia tonsa has been widely used for ecotoxicological testing owing to its ecological 

importance and since it is easy to cultivate in the laboratory (Bushong et al., 1990; Kusk and Petersen, 

1997). A sensitive larval inhibition test has been developed and used to study effects of e.g., tributyltin 

(TBT) (Kusk and Petersen, 1997) and estrogens (Andersen et al., 2001). In short, this test covers the period 

of development from egg until approximately 50 % of the control organisms reach a copepodite stage 

(after 5-8 days depending on temperature). 

5.1.3.1 Protocols using copepod species 

164. Acute toxicity tests with copepods have been established as Swedish (SIS 1991) and International 

Standards (ISO 1997). One chronic test protocol has been established by the American Society for Testing 

and Materials (ASTM E2317-04; ASTM 2004b). 

165. OECD is preparing two test guidelines to standardize procedures for evaluating the effects of 

chemicals on calanoid (OECD 2004b) and harpacticoid (OECD 2004c) copepods (Annex 3 and Table 4). 

The calanoid guideline is designed specifically for Acartia tonsa and involves exposure of the animals 

from the egg to adult egg producing stage (F0 generation). Endpoints for this portion of the test include 

early life stage development, the onset of egg production and stable egg production, survival, sex ratio, and 

copepod length. The early life stages of the F1 generation are exposed and the effects of the test substance 

on development are evaluated. The primary endpoint associated with this portion of the test is the Larval 

Development Ratio, which compares the total number of copepodites produced to the total number of 

early-stage individuals (nauplii + copepodites). The first portion of the test lasts 14–17 days, whereas the 

second runs 5–7 days. The harpacticoid guideline (Table 6) suggests that Amphiascus tenuiremis, Nitocra 

spinipes, and Tisbe battagliai are appropriate species to use for this guideline. Newly hatched larvae (<24-

h old) are placed individually into microwell chambers (or other suitable containers) and exposed to the 

test substance. When the larvae mature (about 10–15 days) they are examined to determine gender, and 

then mating pairs are created and placed into individual microwell chambers. The pairs are exposed to the 

test substance for 7–14 days, during which time mating occurs. The test continues until the females have 

produced at least one brood. The numbers of young produced during exposure to the test material are 

compared to those produced in the control treatments. Additional endpoints that should be recorded are 
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adult survival and the time to the first brood. Brood size, the proportion of infertile eggs, and time between 

broods may also be used in the evaluation of a test substance. Basically the same life cycle protocol already 

exists as an ASTM standard (ASTM 2004b). 

5.2 INSECTS AS REPRESENTATIVE AQUATIC INVERTEBRATES 

5.2.1 Chironomids as representative insects 

166. The family Chironomidae (true or non-biting midges) belongs to the order of Diptera and 

consists of several thousands of species with a worldwide distribution (Pennak 1989). As insects they 

represent the species richest and ecologically one of the most important classes of invertebrates. 

Chironomid larvae (but also other life stages) are an important food source for fish and other predators, 

which is partly linked to their numerical abundance. The larvae inhabit nearly all freshwaters and often 

represent more than half of the total number of macroinvertebrate species. The larvae of some species are 

found in saltwater and a number of species develops in terrestrial habitats as well. Chironomid larvae have 

a length of 1 to 30 mm (depending on the species and larval stage) and feed on algae, higher plants, and 

organic detritus (Pennak 1989; Armitage et al. 1995). They undergo a complete metamorphosis (i.e. are 

holometabolic) and the life stages comprise an egg stage, four larval stages (L1 to L4), pupa and imago. 

Chironomids are sexually reproducing organisms and the sexes are easily separated since the males have 

conspicuous plumose antennae. The larvae are sediment dwellers and inhabit different types of sediment 

depending on physical and nutritional criteria. The four larval stages take up most of the life span, while 

the imago lives only for a few days (at maximum 15% of the life span). Considering that the short-lived 

adults have strongly reduced mouthparts, hence only a limited ability to take up food, chironomids may be 

regarded as true aquatic insects. Adult midges swarm near water bodies and in the swarm mating takes 

place. Shortly after mating the female produces an egg-rope containing a few hundred eggs, which is 

usually attached to some substrate at or just below the water surface. After hatching, larvae sink to the 

sediment or alternately may be carried by the water flow and thus are able to colonize downstream 

habitats. 

5.2.1.1 Protocols using chironimid species 

167. Chironomid species such as Chironomus riparius and C. tentans have been used extensively in 

the toxicity testing of chemicals. Hence, many data are available for comparison and numerous laboratories 

and authorities are familiar with this test organism. A great variety of toxicity tests with chironomids have 

been published. Experimental designs include: a 48-h larval acute test; a 10-d larval survival and growth 

test; a 28-d emergence and development test up to a 65–d life-cycle study. All test systems include water 

and sediment, apart from the 48-h test, which is a water-only test. OECD toxicity testing guidelines are 

available for examining spiked water (OECD 2004e; i.e. OECD TG No 219) and spiked sediment 

scenarios (OECD 2004d; i.e. OECD TG No 218) (Table 4). For C. tentans, standard toxicity testing 

methods have been published by the US EPA (US EPA 1996b, 1996c). Despite the fact that these tests are 

chronic, they do not include reproductive parameters. Reproduction is typically covered in life-cycle tests, 

and effects of different chemicals on reproductive parameters have been published in life-cycle studies 

with different designs (Taylor et al. 1993; Kahl et al. 1997; Watts et al. 2001; Ristola et al. 2001; Hooper et 

al. 2003, 2005; Hruska and Dube, 2005). The US EPA (2000) published a method for conducting a lifecycle test with C. tentans (background information provided in Benoit et al., 1997; Sibley et al. 1997, 

2001) and Taenzler et al. (2006) and Weltje et al. (2006) developed a life-cycle testing method for 

C. riparius, based upon the OECD TG 219 (Table 5). A step further is the multi-generation test, which was 

conducted by Postma and Davids (1995) who studied adaptation to cadmium in nine consecutive 

C. riparius generations. Finally, it is possible to use deformities in the mouthparts of the larval head 

capsule as a sublethal parameter for toxicant exposure. These deformities are mainly linked to the presence 

of metals in sediments, as was demonstrated for field-collected larvae, but it is also observed in laboratory 
ENV/JM/MONO(2006)22 

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assays. Interestingly, the vertebrate estrogen 4-n-nonylphenol induced mentum deformities in a 

concentration-dependent manner (Meregalli et al. 2001), while 17α-ethynylestradiol did not (Meregalli and 

Ollevier 2001). 

168. The endocrine system of insects is the best described within the invertebrates. This is related to 

the silk-worm industry and the development of 3rd generation insecticides that specifically target the insect 

endocrine system. These insecticides act as juvenile hormone (ant)agonists or ecdysone (ant)agonists, and 

interfere with various processes, including molting, metamorphosis, vitellogenesis and reproduction. Third 

generation insecticides can be used as reference chemicals for inducing effects on endocrine processes in 

insects. In addition, a co-exposure of the test compound with an antagonistic reference chemical can help 

elucidating a working mechanism. The available knowledge on the insect endocrine system and the 

availability of specific reference chemicals is a major advantage of working with aquatic insects on 

endocrine disruption in invertebrates. Furthermore, the size of chironomid larvae allows for making 

biochemical measurements, such as vitellogenesis, to generate evidence on the endocrine working 

mechanism of the chemical under investigation (Hahn and Schulz 2002; Hahn et al. 2002). Moreover, a 

cell line of C. tentans is available to test the interactions of chemicals with the ecdysteroid receptor (Quack 

et al. 1995; Spindler-Barth et al. 1998). This cell line could be used to identify potential endocrine 

disruptors for arthropods. Finally, population growth models are available (e.g. Pery et al. 2002, 2005; 

Ducrot et al. 2004; Forbes and Cold 2005) that can be used to analyze the data from life cycle tests and 

assess the importance of the observed effects on a population level. In the end this is more important than 

clarifying the exact working mechanism. 

5.2.2 Protocols using other sediment-dwelling species 

169. The amphipod Hyalella azteca has been used in chronic toxicity testing to study potential 

sublethal effects of contaminated sediments. As e.g. C. tentans this species is an ecologically important 

sediment dweller in freshwater systems. The protocol includes survival, growth, and reproductive 

endpoints (Table 5). The species has been used successfully in EDC testing, although not necessarily 

following procedures similar to this protocol. The EDC testing already performed on this species 

highlights the utility of including sediment-dwelling species in EDC evaluations. Similar protocols have 

been developed by Environment Canada (1997a, 1997b). 

170. Leptocheirus plumulosus: US EPA (2001) developed a chronic protocol to evaluate the potential 

sublethal toxicity of contaminated sediments to an estuarine amphipod, Leptocheirus plumulosus (Table 5). 

The procedure runs for 28 days and includes survival, growth rate, and reproductive endpoints. The test 

requires sediment as a substrate for the amphipods, but can be used with sediments having a wide variety 

of pore water salinities (1‰–35‰). This protocol could be adapted to the EDC program to test the 

potential effects of sediment-bound EDCs (created by spiking sediments) on a sediment-dwelling estuarine 

organism (Table 5). 
 ENV/JM/MONO(2006)22 

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Table 3. Recommended Mysid Life Cycle Toxicity Test Conditions 

ASTM E1191 (ASTM 1997) OPPTS 850.1350 (US EPA 1996)

Test Species: Americamysis bahia 

A. bigelowi 

A. almyra 

Americamysis bahia

Holding Conditions: Hold at conditions similar to test or acclimate 

gradually to test conditions (Temperature at 

3oC/12 h salinity at <3‰/24 h) 

76-L aquaria 

Flow through or recirculating system 

14 h light:10 h dark, or 16 h light: 8 h dark, 

with 15–30 min transition period 

Gentle aeration 

Feed excess ≤24 h old Artemia; 150/mysid/day; 

may supplement with algae or other food 

Hold at conditions similar to test or acclimate 

gradually to test conditions (Temperature at 

1oC/24 h; salinity at <5‰/24 h) 

Flow through or recirculating system 

14 h light:10 h dark, with 15–30 min transition 

period 

Aeration if needed 

Test Setup: 

Test organism age: ≤24 h ≤24 h 

Duration: ≥7 days after median first brood release in 

controls 

28 days 

Test Material: Reagent grade or better NS a

Endpoint(s): Survival, growth, reproduction Survival, growth, young produced 

Number of Treatments: ≥5 plus control (add solvent control if 

necessary) 

≥5 plus control (add solvent control if 

necessary) 

Concentration Series: Test concentrations should bracket the highest 

concentration at which there is not an 

unacceptable effect; each concentration should 

be at least 50% of the next highest 

concentration 

5 or more concentrations chosen in geometric 

ratio between 1.5 and 2.0. 

Dilution Water: 

Solvent: 

Natural or reconstituted seawater acceptable to 

saltwater mysids; uniform quality during test; 

should not affect test outcome 

Must allow satisfactory survival, growth, and 

reproduction 

If solvent used, ≤0.1 mL/L concentration. 

Natural (>20-μm-filtered) or artificial 

seawater 

If solvent used, ≤0.1 mL/L concentration 

Flow Conditions: Flow through Flow through 

Delivery System: 

Flow Rate: 

Calibration limit: 

Calibration/ Check: 

Proportional diluter 

>5 volume additions/24 h (must be capable of 

10 additions/24 h ) 

<10%/chamber/time 

Prior to test; visual check twice daily 

Proportional diluter 

5 × chamber volume/24 h 

<10%/chamber/time 

Prior to test; twice daily 

Number of Replicates: Variable, estimated according to expected 

variation, desired detection limit, and selected 

power. 

5+ (minimum 40 mysids/treatment) 

Test Chamber: e.g., 300 mm × 450 mm × 150 mm deep with 

adequate compartments (to provide 30 

cm2

/mysid). 

Volume NS; materials must minimize sorption 

of test chemicals; loosely covered 

Test Volume: Solution depth ≥100 mm (in above specified 

chamber) 

NS 

Number of organisms/rep: NS (recommends 1 male-female pair/ 

compartment, but can’t determine gender 

for~12 d) 

8 (maximum) 

Other Setup Notes: NS NS 

Initiation Notes: NS NS 

Test Conditions: 

Light: NS NS 

Photoperiod: 14 h light:10 h dark, or 16 h light:8 h dark, with 

15–30 min transition period 

14 h light:10 h dark with 15–30 min transition 

period 
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ASTM E1191 (ASTM 1997) OPPTS 850.1350 (US EPA 1996)

Temperature: 27oC (for A. bahia);  3oC individual 

measurements: ±1oC time-weighted average; 

<2oC difference between any two jars measured 

concurrently. 

25°C ± 2oC 

pH: 6.6–8.2 NS 

Dissolved Oxygen: 

Aeration: 

A concentration between 60–100% of saturation 

is best. 

yes 

60–105% saturation 

Salinity: 15–30‰; variation among treatments should be 

<5, must be <10 ‰ 

20‰ ± 3‰ 

Monitoring: 

Test Concentration Twice prior to test, at 24 h apart; Measured 

concentration ≤30% of nominal concentration. 

During test frequently enough to establish 

average and variability, at least weekly. 

At Day 0, 7, 14, 21, 28; should vary <20% 

among replicates/concentration 

WQ Frequency: Salinity daily; temperature in one chamber 

hourly or min/max measured daily; pH at start 

and end of test and weekly in control, include 

highest concentration; dissolved oxygen in at 

least one test chamber at start and end and 

weekly 

Weekly (includes pH) 

Observation Frequency: Daily 

Count, determine gender, remove dead G1 

mysids; count live mysids; record number live 

females 

Record day of brood release; count and remove 

young daily 

Record abnormal development and aberrant 

behavior 

Periodically; record number dead on Day 7, 

14, 21, 28 

Feeding: Live brine shrimp nauplii at least once daily; 

may supplement 

Dead brine shrimp should be removed daily 

before feeding occurs. 

Recommend 48-h-old Artemia. Frequency and 

amount not specified. 

Other Monitoring Notes: Weekly determinations of particulate matter, 

total organic carbon, and total dissolved gasses 

desirable 

NS 

Termination Notes: Count live G1 mysids and determine gender 

Desirable to measure total body length (anterior 

tip of carapace to tip of uropod) 

Obtain dry weight of surviving G1 (males and 

females separate); remove any brine shrimp 

present; rinse mysids in deionized water, dry at 

60oC for 72–96 h 

Morphological observations at end of test may 

be desirable 

May be desirable to hold G2 mysids for 4+ day 

longer to observe possible effects 

Record number of dead on Day 7, 14, 21, 28 

Record number of males & females and 

measure body length (anterior tip of carapace 

to tip of uropod) when distinguishable and on 

Day 28. 

Count and separate G1 offspring as produced, 

hold at test concentrations. 

If possible (i.e., by Day 28), count, determine 

gender and measure G2 mysids. 

Record abnormal behavior or morphology. 

Test Validity Criteria: A test is valid if 

General test requirements are met 

≥70% G1 control survival 

≥75% G1 control females produce young 

≥3 average number of young/female 

A test is valid if 

≥75% G1 control females produce young 

≥3 average number of young/female/day 

Range-Finding Test 

Concentration Series 

number of samples 

test volume 

test containers 

number of animals/rep 

duration 

NS Widely spaced; e.g., 1, 10, 100mg/L 

1 

NS 

NS 

Minimum 10/concentration 

NS, allow estimate of test concentrations 
 ENV/JM/MONO(2006)22 

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ASTM E1191 (ASTM 1997) OPPTS 850.1350 (US EPA 1996)

Termination Notes: NS NS 

Test Validity Criteria: NS NS 

Reference Toxicant: NS NS 

Concentration Series 

 number of reps 

test volume 

test containers 

number of animals/rep 

NS NS 

Termination Notes: NS NS 

Test Validity Criteria: NS NS 

 a) NS Not specified. 
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Table 4. Example OECD protocoles for evaluating chronic toxicity in aquatic arthropods. 

OECD TG 211 

September 1998 

OECD TG218/219 

February 2001 

OECD Draft Calanoid TG 

Test Species: Daphnia magna Chironomus riparius, C. tentans, C. 

yoshimatsui

Acartia tonsa 

Holding/ Culturing Conditions: 

Hold at conditions similar to test Hold at conditions similar to test Hold at conditions similar to test 

Test Setup: Test organism age: <24 h; must not be first brood 

progeny 

1st instar larvae 1st stage nauplii 

Duration: 21 d 20–28 d; 28–65 d depending on 

species. If midges emerge earlier, test can be terminated five days after emergence of the last adult in the control. 

19–25 d 

Test Material: Chemical in water Spiked Sediment (218) or spiked 

water (219) 

Seawater (natural or artificial) 

Endpoint(s): Primary: Number of offspring per 

adult alive at end of test Other: parent survival; time to first brood Draft enhanced endpoints for endocrine disruption evaluations: offspring sex ratio; molting inhibition 

Total adult emergence; development time 

F0: (14–17 d) early life-stage development, onset of egg production and stable egg production, survival, sex ratio, bodlength F1: (5–8 d) development (larval development ratio, LDR) 

Number of Treatments: 

≥5 and a laboratory control ≥5 and a laboratory control ≥5 and a laboratory control 
 67

OECD TG 211 

September 1998 

OECD TG218/219 

February 2001 

OECD Draft Calanoid TG 

Concentration Series: Geometric series with separation 

factor 

≤3.2; based on acute test or 

range-finding test 

Factor between concentrations should be 

≤2 for ECx studies; 

≤3 for 

LOEC/NOEC studies 

Geometric series with separation factor 

≤3.3 

Test concentrations should not include any that have a statistically significant effect on survival of the F0 generation since the main 

objective is to measure sublethal effects 

Dilution Water: Solvents: Dispersants: 

May be used to make stock for dosing test (e.g., acetone, ethanol, methanol) May assist in accurate dosing and dispersion (e.g., Cremophor RH40, HCO-40) 

Any suitable natural or synthetic water may be used. The water has to be aerated before use. May be used to make stock for dosing test (e.g., acetone, ethanol, methanol) 

DO saturation value >70%; pH = 8± 0.3 before use 

Flow Conditions: Semi-static; flow-through 

allowable 

Static Static-renewal 

Number of Replicates: 

10 (semi-static); 4 (flow through) 

≥3 for ECx study; 

≥4 for 

LOEC/NOEC study 

Control: minimum 6, 12 recommended for LDR test Test substance: depends on design(linear regression), 6 (ANOVA) 

Test Chamber: Glass beakers; volume not 

specified 

600-mL, 8-cm diameter glass beaker 100-mL beaker 

Test Volume: 50–100mL 1.5–3.0 cm sediment depth 80 mL Number of organisms/rep: 

1 (semi-static); 10 (flow through) 20 for C. riparius and C. 

yoshimatsui; 12 for C. tentans

10 

Other Setup Notes: Renew test solution at least 3 

times/week; more frequently if not stabile (<80% initial concentration over 3 d) Prepare new test vessels and transfer adults to them 

Formulated sediment (also called reconstituted, artificial or synthetic sediment) should be used. Spike sediment for TG218; spike water for TG219 

LDR test: renew 50–80% on Day 3 or increase volume Basic exposure: renew 50–80% every 2nd day or increase volume Prepare new test vessels and transfadults to them 
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OECD TG 211 

September 1998 

OECD TG218/219 

February 2001 

OECD Draft Calanoid TG 

Test Conditions: Light: Ambient laboratory illumination 

<15-20 μE/m2/s 

About 500 to 1000 lux Wide-spectrum fluorescent lights; 

low intensity; ~5-10 μmol/s/m2

Photoperiod: 16L:8D 16L:8D 12L:12D Temperature: 18–22°C; should not vary >2°C 

during any one test 

20°C ± 2°C; 23°C ± 2°C for C. tentans

15–20°C; vary < 2°C during test 

pH: 6–9; should not vary >1.5 units 6–9 at start of test 8.0 ± 0.3 Dissolved Oxygen: Aeration: 

3 mg/L None allowed 

>60% saturation value Gentle aeration; 1 bubble/second 

>70% saturation value if DO level falls below 70% saturation: 1 bubble/second 

Hardness/ Salinity: >140 mg/L CaCO3 <200 mg/L CaCO3 ~20 ‰ ± 2 ‰ Monitoring: WQ Frequency: DO, temperature, hardness, pH: 

once per week in fresh and old media, controls, and highest test concentration Determine concentration of test substance regularly 

DO measured daily in all test vessels Hardness and ammonia at the beginning and at the end of the test in controls and one vessel of highest concentration. Temperature and pH at start and end of test. 

Measure dissolved oxygen, pH, salinity, and temperature in controand all test concentrations each timtest medium is renewed. As a minimum, measurements shall be made in the control(s) and highest test concentration. Temperature preferably monitored continuously

Observation Frequency: 

Not specified At least three times per week; visual 

assessment of abnormal behaviour (e.g. leaving sediment, unusual swimming), compared with the control. 

Daily for egg production test; not specified for LDR test 
 69

OECD TG 211 

September 1998 

OECD TG218/219 

February 2001 

OECD Draft Calanoid TG 

 Feeding: At least 3 times per week, 

preferably daily (semi-static); concentrated algal suspension 

At least 3 times per week, preferably daily; 0.25-0.5 mg fish-food (a suspension in water or finely ground food) per larvae per day for the first 10 days. Slightly more food for older larvae; 0.5–1.0 mg per larvae per day Reduce food ration in all treatments and control if fungal growth occurs or if mortality is observed in controls. If fungal development cannot be stopped the test is to be repeated. 

LDR test: 5 × 104 cells/mL of medium added on days 0 and 3 Basic exposure: 5 × 104 cells/mL omedium added on day 0 and subsequently to every renewal of thmedium. Egg production: 5 × 104 cells/mL medium added daily. 

Other Monitoring Notes: 

Record parent animal mortality daily Remove and count offspring daily Record presence of aborted eggs or dead offspring daily 

During the period of expected emergence, a daily count of emerged midges is necessary. 

Termination Notes: Record number of fully emerged 

male and female midges 

Fix animals in Lugol’s solution; score biological parameters (countmeasure, determine gender, etc.) 
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OECD TG 211 

September 1998 

OECD TG218/219 

February 2001 

OECD Draft Calanoid TG 

Test Validity Criteria: 

Parent animals (female Daphnia) mortality ≤ 20% at the end of the 

test Mean number of live offspring produced per parent animal surviving at the end of the test > 60 

Control mortality ≤30% at end of 

test C. riparius and C. yoshimatsui emergence to adults occurs 12–23 days after initiation; C. tentans emergence to adults occurs 20–65 days after initiation Mean emergence in controls 50−70 

per cent DO concentration ≥60% air saturation value at the end of the test pH of overlying water 6–9 range in all test vessels Water temperature vary ≤ 1.0 °C between vessels during test Temperature held within the ranges specified for test species 

Mortality in mass culture control(s≤20% at end of exposure DO concentration ≥70% air saturation value throughout the exposure period Temperature vary ≤ 2°C during the test pH vary ≤ 0.3 units from average control pH Salinity vary ≤ 10% from the contrstart value Sex ratio in controls = 40:60 to 60:40) Hatching success in hatching contr≥ 80% Average mortality in the control(s) on day of LDR observation ≤ 20% Average number of eggs/female /d controls ≤ 30 at Day 12 and later Control mortality of the isolated females in the egg production test (Day 7–12) ≤ 20%. 
 71

Table 5. Example US EPA protocols for evaluating chronic toxicity in Aquatic Arthropods 

US EPA-821-R-02-013 

October 2002 

US EPA/600/R-99/064 

March 2000 

US EPA/600/R-99/064 

March 2000 

Test Species: Ceriodaphnia dubia Chironomus tentans Hyalella azteca Holding Conditions: Hold at conditions similar to test Hold at conditions similar to test or 

acclimate gradually to test conditions 

(Temperature at 1 oC/1-2 h) 

Hold at conditions similar to test oacclimate gradually to test conditio(Temperature at 1 oC/1-2 h) 

Test Setup: Test organism age: <24 h and all within 8 h of the same 

age 

1 d (<24 h) 7 to 8 d 

Duration: Maximum of 8 d 

Until 60% or more of surviving control females have three broods 

50–65 d depending on emergence. Each treatment may need to be terminated separately 

42 d 

Test Material: Effluents and receiving waters Sediment Sediment Endpoint(s): Survival and reproduction 20-d survival and weight; female and 

male emergence, adult mortality, the number of egg cases laid, the number of eggs produced, and the number of hatched eggs 

28-d survival and growth; 35-d survival and reproduction; and 42-survival , growth, reproduction, annumber of adult males and females on Day 42 

Number of Treatments: 

Effluents: 5 and a control Receiving Water: 100% receiving water (or a minimum of 5) and a control 

Concentration Series: 

Effluents: 

≥ 0.5 

Receiving Waters: None or 

≥ 0.5 

Dilution Water: Solvent: 

Uncontaminated source of receiving or other natural water, synthetic water prepared using Millipore milli-Q® or equivalent deionized water and reagent grade chemicals or DMW 
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US EPA-821-R-02-013 

October 2002 

US EPA/600/R-99/064 

March 2000 

US EPA/600/R-99/064 

March 2000 

Flow Conditions: N/A 2 volume additions/d; continuous or 

intermittent (e.g., one volume addition every 12 h) 

2 volume additions/d; continuous ointermittent (e.g., one volume addition every 12 h) 

Number of Replicates: 

10 (required minimum) 16 (12 at Day -1 and 4 for auxiliary 

males on Day 10) 

12 (4 for 28-d survival and growth and 8 for 35- and 42-d survival, growth, and reproduction) 

Test Chamber: 30-mL borosilicate glass beakers or 

disposable polystyrene cups (recommended because they fit in the viewing field of most stereoscopes). 

300-mL high-form lipless beaker 300-mL high-form lipless beaker 

Test Volume: 15 mL (minimum) 100 mL sediment, 175 overlying 

water 

100 mL sediment, 175 mL overlyinwater in the sediment exposure froDay 0 to Day 28 (175 to 275 mL in the water-only exposure from Day 28 to Day 42) 

Number of organisms/rep: 

1 Assigned using blocking by known parentage 

12 10 

Other Setup Notes: New test solutions are prepared 

daily, and the test organisms are transferred to the freshly prepared solutions 

Test Conditions: Light: Ambient laboratory illumination 10–

20 μE/m2/s, or 50–100 ft-c 

Wide-spectrum fluorescent lights; About 100 to 1000 lux 

Wide-spectrum fluorescent lights; About 100 to 1000 lux 

Photoperiod: 16L: 8D 16L: 8D 16L: 8D Temperature: 25 °C±1 °C 23 °C ± 1 °C 23 °C ± 1 °C pH: Dissolved Oxygen: Aeration: 

4.0 mg/L used only as a last resort 

>2.5 mg/L if DO level falls below 2.5 mg/L: 1 bubble/second 

>2.5 mg/L if DO level falls below 2.5 mg/L: 1 bubble/second 
 73

US EPA-821-R-02-013 

October 2002 

US EPA/600/R-99/064 

March 2000 

US EPA/600/R-99/064 

March 2000 

Salinity: N/A N/A N/A Monitoring: WQ Frequency: DO, temperature, pH: daily, prior to 

renewals in at least one test chamber at each concentration and control pH is measured in the effluent sample each day before new test solutions are made Conductivity, alkalinity, and hardness measured in each new sample and in the control 

Hardness, alkalinity, conductivity, and ammonia at the beginning, on Day 20 and at the end of the test. Temperature daily (ideally continuously). DO and pH three times/week. Conductivity weekly. Measure DO more often if DO has declined by more than 1 mg/L since previous measurement. 

Hardness, alkalinity, conductivity, and ammonia at the beginning and end of a sediment exposure (day 0 and 28). Temperature daily. Conductivity weekly. DO and pH three times/week. Measure DO more often if DO has declined by more than 1 mg/L since the previomeasurement. 

Observation Frequency: 

Daily Daily to assess test organism 

behavior such as sediment avoidance 

Daily to assess test organism behavior such as sediment avoidan

Feeding: 0.1 mL each of YCT and algal 

suspension per test chamber daily 

1.5 mL of Tetrafin® (4 mg/mL dry solids) to each beaker. If fungal or bacterial growth develop from excess food, feeding should be suspended for one or more days. If feeding is suspended in one treatment, it should be suspended in all treatments. 

YCT food, fed 1.0 mL (1800 mg/L stock) daily to each test chamber. If fungal or bacterial growth develfrom excess food, feeding should bsuspended for one or more days. Ifeeding is suspended in one treatment, it should be suspended iall treatments. 

Other Monitoring Notes: 

 At 20 d, 4 of the initial 12 reps. are 

selected for use in growth and survival measurements. AFDW of midges should be determined for the growth endpoint. Emergence traps are placed on the reproductive replicates on day 20 
ENV/JM/MONO(2006)22 

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US EPA-821-R-02-013 

October 2002 

US EPA/600/R-99/064 

March 2000 

US EPA/600/R-99/064 

March 2000 

Termination Notes: 

Because of the rapid rate of development of Ceriodaphnia dubia, at test termination all observations on organism survival and numbers of offspring should be completed within two hours. 

Clean sediment will typically require 40–50 d from initial setup to completion (emergence). Environmental stressors will reduce growth and delay emergence. For treatments in which emergence has occurred, the treatment (not the entire test) is ended when no further emergence is recorded over a period of 7 d. 

Growth can be reported as either length or weight Length should be measured ± 0.1 mm from the base of the first antenna to the tip of the third uropoalong the curve of the dorsal surfacDry weight of amphipods in each replicate can be determined on Day 28 and 42. 

Test Validity Criteria: 

Mean control survival ≥ 80%; 60% of surviving control females must produce at least three broods, with an average of 15 or more young per surviving female. 

Average size of C. tentans in the control at 20 d ≥ 0.6 mg/surviving 

organism dry weight or 0.48 mg/ surviving organism AFDW. Emergence ≥ 50%; mean number of eggs/ egg case ≥ 800; percent hatch ≥ 80% Hardness, alkalinity, and ammonia in overlying water should not vary by more than 50% during the test, DO > 2.5 mg/L 

Mean control survival ≥ 80% on D28. Hardness, alkalinity, and ammonia the overlying water typically shoulnot vary by more than 50% during the sediment exposure, and dissolvoxygen should be maintained abov2.5 mg/L in the overlying water. 
 ENV/JM/MONO(2006)22 

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6.0 DESCRIPTION OF ASSAY ENDPOINTS REFLECTIVE OF REPRODUCTIVE AND 

DEVELOPMENTAL IMPAIRMENT 

171. In considering the list of potential endpoints presented in the following sections, it is desirable 

that the measured response is directly related to exposure to the potential EDC and that the response is the 

result of interruption of endocrine function (Ingersoll et al. 1999). Although many, if not all of the 

endpoints described in the following sections could indicate a response to an EDC, most also will vary in 

response to exposure to other stressors. Further confounding the interpretation of testing results is the 

interrelatedness of some endpoint measurements. Thus, a stressor could act on a test organism to reduce its 

swimming ability, thereby reducing its ability to feed, which then reduces its growth. Similarly, reduced 

fecundity could be an expression of reduced growth. Often, exposure to a compound that results in 

reduced fecundity also results in reduced growth (Nimmo et al. 1981); however, this is not always the case. 

Lussier et al. (1999) reported fecundity effects without reduced growth, and McKenney and Celestial 

(1996) found reduced reproductive measures at concentrations of methoprene below that which resulted in 

reduced growth. Various endpoint measures can have differing sensitivities to stressors. For example, 

reproductive parameters are often, but not always more sensitive measures of contaminant toxicity than is 

simple survival (Lussier et al. 1985). Hollister et al. (1980) found no reproductive impairment among test 

organisms, in spite of significant mortality due to exposure to alkaline effluent. Additionally, stressors can 

affect some reproductive measures, whereas others cannot (e.g., Hollister et al. 1980, Lussier et al. 1985). 

Clearly, the possible interrelationships among potential endpoints and the varying sensitivities shown by 

some endpoints underscore the need to use multiple endpoints in EDC testing. Examples of assay 

endpoints reflective of reproductive and developmental impairment are given below. The list is not 

complete but rather involves general endpoints identified in the protocols described in section 5. 

6.1 GROWTH, MORPHOLOGICAL AND BEHAVIORAL ALTERATIONS 

172. Measures of growth, morphological changes, and changes in test organism behavior as indicators 

of EDC effects are reviewed below. 

6.1.1 Growth 

173. Molting is one of the key arthropod physiological processes that are under hormonal control, and 

therefore, it is susceptible to the negative effects of EDCs. Molting is regulated primarily by the 

interaction of molt stimulating hormones (ecdysteroids) and nervous system secretions produced in the 

cephalothorax with molt inhibiting hormones produced in the eyestalks (e.g., Cuzin-Roudy and Saleuddin 

1989, Fingerman 1997, Subramoniam 2000, Zou and Fingerman 1997). Molting is either directly or 

indirectly involved in the expression of the various endpoints that may be examined through toxicological 

testing. Because noticeable growth can only occur as a result of molting, any disruption of molting could 

result in alterations in growth. Many pesticides, generally classed as Insect Growth Regulators (IGR), 

have been developed that target insect development. IGRs function as ecdysone agonists, JH or anti-JH 

analogs and as chitin synthesis inhibitors. All can have detrimental effects on crustaceans, especially 

through interruption of the molt cycle, most likely by impacting endocrine function (Touart 1982). For this 

reason, the estimation of molting frequency may be a useful endpoint relevant to mysid and other 

crustaceans for EDC testing. The duration of the molt cycle of adult mysids (A. bahia) was determined to 

be about 6.6 d (Touart 1982), with the duration for male being slightly less and that for females being 

slightly more. Juveniles are expected to molt within the first 24 h post release, but may delay molting until 

about 48 h (Touart 1982). Successive juvenile molts occur at increasingly longer durations. Touart (1982) 

found that at test conditions of 22oC and 20‰, sexual characters were noticeable after the fourth molt (9–

12 d) and that mating occurred after the fifth or sixth molts (17–19 d). Higher test temperatures would 
ENV/JM/MONO(2006)22 

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likely shorten the duration between molts. Molt staging relies on changes in the integument and is most 

commonly divided into four stages: postmolt, intermolt, premolt, and molt (Verslycke et al. 2004a). 

Winkler and Greve (2002) found that temperature, body size, and age all are factors contributing to mysid 

growth rates. Touart also found that the pesticide diflubenzuron (Dimilin®) increased the duration of the 

molt cycle in A. bahia, probably acting on the mysid endocrine system as a molt inhibitor. This was 

demonstrated by his experiment in which eyestalk-ablated mysids that were then exposed to dimilin had 

molt cycle durations that did not differ from those in control animals. In effect, the Dimilin may have 

replaced or mimicked the “natural” mysid molt inhibitor synthesized in the eyestalks. Zou and Fingerman 

(1997) suggest a certain degree of overlap between molting hormones and estrogen-mimicking 

xenobiotics. They tested this hypothesis using two compounds, DES and endosulfan. Molting time for 

daphnids exposed to DES and endosulfan was delayed because these compounds blocked the ecdysteroid 

receptors, preventing molting hormones from binding to the receptor and turning the receptor on, thereby 

slowing the molting process. However, it is unclear whether the delayed molting process was due to the 

exposure to these contaminants or just a general response to stressors. More pharmacokinetic studies are 

underway to investigate delayed molting further. 

174. Because there are current immunoassay methods for quantifying ecdysteroids and also methods 

for molt staging, evaluations of potential interactions of EDCs and molting should be possible (Block et al. 

2003; Verslycke et al. 2004a). 

175. Two direct measures of growth may be used in the assessment of sublethal effects on crustaceans. 

Probably the most common measurement is the determination of simple dry weight. At test termination, 

animals are briefly rinsed in deionized water to remove salt, and dried to constant weight. In published 

studies that employ this method, drying temperatures ranged from 60oC (e.g., McKenney 1982) to 105oC 

(Khan et al. 1992) and drying times ranged from 16 h (Khan et al. 1992) to 72 h (UCSC 1998). ASTM 

(1997) specified that growth should be measured by drying animals at 60oC for 72 h to 96 h, or to constant 

weight. ASTM also recommended that separate determinations be made for males and females. Ash-free 

dry weight (AFDW) is often used as the appropriate weight measurement for many invertebrates, because 

the technique reduces any inaccuracies introduced by inorganic constituents in the organism’s body. 

Inorganic components can originate from processes such as the development of skeletal components or 

feeding (the ingestion of sediment). As with some other small crustaceans, the small size of the mysids 

and their planktonic diet may make the removal of ash from the dry weight measurement an unnecessary 

step that would not greatly improve the accuracy of the measurement (US EPA 2000). 

176. Another direct measurement of growth is body length. Body length has been measured as the 

distance from the base of the eyestalks to the tip of the telson (Hunt et al. 1997; UCSC 1998) or to the tip 

of the exopod (Langdon et al. 1996; Winkler and Greve 2002). ASTM (1997) suggests that body length 

should be measured along the midline of the body from the tip of the carapace to the tip of the exopod 

(excluding the terminal setae). ASTM notes that it is difficult to make this measurement on preserved 

animals because of the body curvature that results from the fixation process. Langdon et al. (1996) reduced 

the potential inaccuracy in this measurement by relaxing the mysids prior to fixation and then determining 

length as the sum of a series of relatively straight-line measurements. Dahl et al. (in press) defined the total 

body length (μm) of the harpacticoid copepod N. spinipes exposed to sublethal concentrations of the drug 

Simvastatin as the distance between the anterior part of prosome to the posterior part of the last urosomite. 

The study showed that the body length was a less sensitive endpoint than development rate, growth rate 

and RNA content. Winkler and Greve (2002) anaesthetized mysids in soda water to relax organisms prior 

to collecting length measurements. Age-specific growth rates were calculated using duration of the 

successive intermolt periods in days and the growth factor at each molt. Comparisons of different 

temperatures (10°C and 15°C) different size classes, and other mysids provided a combination of results. 

The mysid N. integer reached maturity in a shorter time and at a smaller size than P. flexuosus. P. 
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flexuosus grew faster at all stages but matured much later. Both species had temperature-dependent growth 

curves but in opposite ways. N. integer had a lower growth rate at 10°C than at 15°C. 

177. The most common response to exposure to toxicants is reduced growth (several studies) although 

it may not be a particularly sensitive endpoint (Lussier et al. 1999). McKenney and Celestial (1996) 

reported reduced growth after exposure to methoprene, a juvenile hormone analog. Reduced growth has 

important connections to reproductive success because the size of females is directly related to fecundity 

(Verslycke et al. 2004a). However, reduced growth occurred only at the highest concentration in their 

exposure series. It may not be useful to measure growth for all treatments. Because growth is a sublethal 

endpoint, its use has sometimes been restricted to treatments for which survival was not significantly less 

than that in control treatments (Hunt et al. 1997). Hunt et al. (1997) observed that individuals that survive 

high toxicant concentrations may often be larger than average. In a study conducted by Hunt et al. (2002), 

survival of the mysid, Mysidopsis intii, was the most sensitive of five endpoints measured during a 28-day 

chronic exposure to nickel. Growth is less sensitive in such cases because surviving mysids are large, 

more tolerant of toxicants, and use cannibalism as a nutritional source. The effect of contaminant exposure 

on growth may be related to test organism age in addition to contaminant concentration. McKenney (1986) 

found that older (9–16 d) juvenile mysids showed significantly reduced growth at a lower fenthion 

concentration than that found for younger (0–9 d) juveniles. Reduced growth in mysids possibly results 

from a transfer of energy from growth mechanisms as an organism attempts to counteract stress 

(McKenney 1985, 1989, McKenney and Matthews 1990; Section 6.4.3 below). Reduced growth has strong 

implications for reproductive success in mysids as several studies (e.g., Mees et al. 1994, Turpen et al. 

1994) have shown that fecundity is directly related to female body size. 

178. Molt stage has become an increasingly used technique for measuring growth. If molt stages are 

classified, duration of different stages under normal laboratory conditions, and then environmental effects 

on relative duration of stages can be evaluated, using the molt-stage technique (Gorokhova 2002). Molt 

cycle chronology is a prerequisite for the use of molt staging in growth studies. One study was conducted 

using laboratory reared juveniles and subadults of Mysis mixta and Neomysis integer; molt staging 

techniques were used to determine the main molt stages for each species under different regimes of 

temperature and feeding. Field application of the molt stage technique was applied for Mysis mixta, to 

determine molt cycle duration. 

6.1.2 Morphology 

179. Changes in morphology resulting from exposure to contaminants have been documented for 

many taxa, including arthropods. For example, chironomid (midge) larvae, which are commonly used in 

toxicological testing, can develop mouthpart deformities when exposed to chemicals that affect endocrine 

systems (Meregalli et al. 2001, Watts et al. 2003). The deformities result from physiological disruption 

during molting. Meregalli et al. (2001) found that sublethal concentrations of 4-n-nonylphenol, a known 

endocrine disruptor, caused mouthpart deformities in Chironomus riparius. Relatively low concentrations 

of 17α-ethinylestradiol and bisphenol-A (10 ng/L) caused mouthpart deformities. Typically, deformities 

include, loss, addition, fusing, and splitting and these are assigned a numerical score. When more than one 

deformity was observed, the scores were summed to provide an overall total score. Deformities were 

assessed relative to the normal arrangement of each of three structures (mandibles, the mentum, and the 

ephipharyngeal pectin). Observations were made by using a scanning electron microscope, and the most 

notable deformities were associated with the mentum. In the same study molting in response to these 

chemicals was also recorded and showed no statistically significant differences at the higher 

concentrations. This difference in endpoint sensitivity highlights the need to select multiple endpoints for 

assessment of potential EDCs and the need to understand the mechanism involved in selected effects noted 

for one mouth part over others such that screening tools could be developed for compounds such as EDCS 

that may demonstrate a specific deformity in response to those contaminants. Despite the similarities in 
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molting physiology between insects and crustaceans, it does not appear that morphology has been widely 

considered as an appropriate and measurable endpoint in crustaceans toxicological studies. A laboratory 

study showed that exposure to cadmium resulted in the development of abnormal genitalia in male A. 

bahia and malformed carapaces in males and females (Gentile et al. 1982). The time to the first 

appearance of the abnormalities was positively related to cadmium concentration. One study based on 

field-collected data found that four populations of N. integer contained many individuals with abnormal 

telson morphology (Mees et al. 1995). Such abnormalities could arise through physiological perturbations 

that occur during molting, and they could provide a quantifiable measure of disruption in endocrine-related 

functions. 

180. The determination of the degree of fluctuating asymmetry (FA) found among mysids could 

provide a quantifiable, repeatable measure of morphological perturbations resulting from exposure to 

possible EDCs. FA is the asymmetric development of a normally symmetrical bilateral structure in which 

there is no tendency for one side of the structure to have a larger value than the other (Palmer and Strobeck 

1986, Palmer 1994, Leung et al. 2000). FA is thought to arise because of environmental or genetic stress 

during development (Clarke 1993, Leung and Forbes 1996) and may result from a shift in metabolic 

energy from systems that maintain developmental stability to those that help organism compensate for 

increased stress (Sommer 1996). Measurement of FA, once appropriate characters have been determined, 

is relatively easy and requires only use of a microscope. Characters that have been measured are readily 

observable and include midge larvae teeth (Chironomus, Clarke 1993), copepod body spines (Tisbe, Clarke 

1992), and shrimp antennae (Palaemon, Clarke 1993). However, Leung et al. (2000) mentioned that the 

measurement of single characters does not always reliably indicate environmental stress, and they 

suggested that an approach involving the use of composite indices of FA could increase the probability of 

detecting environmental stress. Although some studies of FA have been done on crustaceans (e.g., Clarke 

1992, 1993), none has focused on mysids. Before FA could be used to detect mysid developmental 

abnormalities resulting from exposure to EDCs, preliminary studies examining several suites of potential 

characters would have to be performed. 

181. Laufer et al. (2002), using male spider crabs Libinia emarginata, showed that ecdysteroids and 

low methyl farnesoate concentration (controlled by standard eyestalk ablation experiments) promoted 

allometric growth (disproportionate growth of body parts; for example, propodus of male spider crabs 

becomes disproportionately larger then the carapace during reproduction) while ecdysone and higher 

concentrations of methyl farnesoate inhibited allometric growth. Again, researchers need to understand the 

mechanistic causes before these changes can be attributed to effects. 

182. The use of histology for assessing impacts associated with EDCs has been largely unexplored as 

an endpoint for crustaceans. One study, using the amphipod Gammarus pulex, examined impacts 

associated with a sewage treatment plant. A highly significant number of females collected displayed 

abnormal structure of oocytes in vitellogenesis. Particularly noted were uneven ooplasma, few yolk 

bodies, and lipid globules that were present and unevenly distributed. In addition, both the males and 

females had decreased body size possibly related to interference with the ecdysteroid receptor, which could 

have impacts on the molt cycle and interfere with vitellogenesis (Gross et al. 2001). 

6.1.3 Behavior 

183. The disruption of crustacean swimming behavior is one endpoint that has recently been 

investigated, particularly in mysid, as potentially informative in documenting sublethal exposure to 

contaminants. In a series of studies, Roast et al. (1998, 2000a, 2000b, 2001) determined that sublethal 

concentrations of cadmium and chlorpyrifos (an organophosphorus pesticide) significantly altered the 

swimming behavior of N. integer. Perturbations included decreased ability to swim against a current, an 

increase in general activity with no improvement in ability to swim against the current, and a reduced 
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tendency to maintain a position near the bottom. These changes in behavior could have important 

ecological impacts to the animals by causing them to be moved to an unfavorable habitat or by increasing 

their susceptibility to predation, thus indirectly resulting in a lethal response to contaminants at sublethal 

concentrations. However, if the disruption of swimming behavior is to be used as an endpoint in endocrine 

disruption studies, it should be applied with caution. Roast et al. speculated that the disruption of 

swimming that they observed was probably a nervous system effect related to interference with cholinergic 

pathways. Additionally, the time required to make behavioural observations is very long, making such 

endpoint very labour intensive; it might not be considered worthwile for a regulatory test. 

184. Reduced feeding has been noted as a sublethal response to exposure to some contaminants. For 

example, Nimmo et al. (1981) noted, but did not quantify, reduced mysid feeding in response to exposure 

to some pesticides. Although difficult to accurately quantify, reduced feeding no doubt is an important 

sublethal effect of contaminant exposure that may be expressed in many of the life cycle parameters 

described below. Reduced feeding could lead to reduced growth, reduced time to maturity, and reduced 

egg production, among other factors. 

185. One study using the amphipod Gammarus pulex looked at the combined effects of water and 

dietary uptake of 4-nonylphenol. Most feeding studies usually focus on one route of exposure and assume a 

constant feeding rate during the course of the study. The exposure route for benthic invertebrates includes 

direct contact with the sediment and pore water, uptake from the water column, and ingestion of food and 

other particles. 4-nonylphenol is more likely to be absorbed to the sediment particles because of its 

hydrophobic nature. The results of this study showed that feeding rates can be determined in a variety of 

ways depending upon how the data are analyzed (variable feeding rate versus mean feeding rate). The 

study showed that significant uptake of 4-nonylphenol occurred from food. 

186. Schmitt and Seuront (2001) performed a 3D copepod trajectory using Temora longicornis, 

recorded in the laboratory. They observed a so-called multifractal random walk, which was considered as 

an inborn behavior, and not induced by uneven distribution of phytoplankton. They also identified three 

different modes of swimming, which may reflect the effect of gravity, predator avoidance, and feeding 

switch between two kinds of different food sources. 

6.2 MEASURES OF REPRODUCTIVE PERFORMANCE 

187. There are several measures of reproductive performance that can be used to assess sublethal 

response. For example, sexual maturity, the time to first brood release, the time required for egg 

development, fecundity, and alterations in reproductive characteristics in populations have all been used as 

endpoints. Zou and Fingerman (1997) showed that conditions of crowding, food shortage, and change in 

length of day could trigger a change from asexual to sexual reproduction, with males being produced 

initially. Females receive environmental cues, which trigger the ovaries to lay male-producing eggs. 

Exogenous agents may interfere with male differentiation and may affect this process of environmental 

cues to the ovaries. 

6.2.1 Sexual Maturity 

188. Khan et al. (1992) asserted that sexual maturity, which they described as the presence of gonads 

or a brood pouch, is a feasible endpoint for reproductive tests, because gonad maturation is essentially the 

first step toward reproductive output. Maturity allows for the measurement of effects to both males and 

females. Khan et al. used a dissection microscope to examine live mysids for the presence of gonads at the 

test termination. They quantified maturity as the ratio of the number of sexually mature mysids to the 

number of surviving mysids in each test replicate, and demonstrated that maturity was a sensitive endpoint 

for tests of 96-h and 7-day duration that were initiated with 8-day-old mysids. Others used the time 
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required for mysids to reach sexual maturity as a test endpoint (e.g., Gentile et al. 1983, Lussier et al. 

1985). Gentile et al. (1983) used the development of testes or the presence of eggs in the oviduct to 

determine when sexual maturity was reached in male and female mysids, respectively. They reported that 

exposure to high levels of mercury significantly lengthened the time required for mysids to reach maturity, 

which then was expressed as delay in the appearance of eggs in the brood pouch and the release of young. 

189. Exposure of H. azteca to 17α-ethinylestradiol showed several reproductive and morphological 

effects including: smaller gnathopods in females; skewed sex ratios favoring females; histological 

aberrations of the reproductive system, such as indications of hermaphroditism, disturbed maturation of 

germ cells; and disturbed spermatogenesis of the post F1 generation males (Vandenbergh et al. 2003). For 

both gnathopod growth and secondary sex characteristics, more pronounced effects were noted at the lower 

concentration (U-shaped curve), which may suggest a receptor-mediated response. It could be that 17αethinylestradiol causes disturbance of the androgen gland through interaction with the AGH or hormone 

metabolizing enzymes with subsequent changes in AGH activity. 

6.2.2 Time to First Brood Release 

190. In uncontaminated systems, the length of time to the release of the first brood is primarily related 

to environmental temperature, with some influence by salinity and an interaction between the two factors 

(McKenney 1996). McKenney (1996) determined that the shortest time to release of the first brood for A. 

bahia was about 16 days at a temperature of 28oC and a salinity of 28‰. The time to release of the first 

brood increases with decreasing temperature and salinity. Because this parameter is measured as the 

number of days from hatching of the mysids used in the test until they release their first brood, it can 

represent the expression of more than one factor, including the length of time it takes a mysid to reach 

sexual maturity and the time required for eggs to develop in the brood pouch before being released. Most 

contaminant effects are likely to lengthen the time to release of the first brood. Lussier et al. (1985) found 

that several metals (mercury, zinc, nickel) significantly increased the time to first brood release, whereas 

others (e.g., cadmium, copper, silver) did not. Gentile et al. (1983) found that very high levels of mercury 

caused mysids to abort eggs that had been deposited in the brood pouch. 

191. The time to first brood release is also one of the main endpoints often used in copepod life cycle 

tests aiming at identifying effects of environmental pollutants with a suspected specific biological activity 

(e.g., Breitholtz and Bengtsson 2001; Chandler et al. 2004). 

6.2.3 Egg Development Time 

192. Egg development time is measured as the number of days between the first appearance of eggs in 

the brood pouch and the first release of juveniles. Gentile et al. (1983) found that mercury did not 

significantly affect the brood duration in mysids, although several other reproductive parameters were 

affected. Winkler and Greve (2002) calculated incubation time for N. integer as the difference between the 

date of laying eggs in the brood pouch and the date of release of the juveniles from the marsupium. Data 

were collected by removing females with fertilized eggs from culture jars and placing them in individual 

containers. Temperature affected the start of maturation. At 15°C, development was much shorter (1.5 

months) than at 10°C (3 months). The reproduction rate of N. integer increased at higher temperatures 

because of declining incubation periods plus an increasing number of neonates released per brood (more 

juveniles released as the female gets larger). Also, at higher temperatures the number of released juveniles 

per brood was highly variable, most likely due to the successive oviposition of the females. Overall, N. 

integer had double the reproduction success of P. flexuosus because of the longer incubation times and 

release of fewer juveniles. 
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193. Similarly, in copepods, the mean number of days from brood sac extrusion to naupliar hatching 

has been used to study egg development time (Cary et al. 2004). 

6.2.4 Brood Size (Fecundity) 

194. Brood size can be measured as the number of eggs per brood (Khan et al. 1992) or as the number 

of young produced, expressed either as total young per female or as young produced per available female 

reproductive-day (e.g., Gentile et al. 1982, Lussier et al. 1985, Baldwin et al. 1995, Hutchinson et al. 1999, 

Breitholtz and Bengtsson 2001, Bejarano and Chandler 2003, Chandler et al. 2004). The latter measure 

may be used to normalize differences in the number of females available per test concentration (Gentile et 

al. 1982). The number of available female reproductive-days is calculated by multiplying the number 

mature females by the number of days survived. McKenney (1996) showed that the number of eggs in the 

first brood of mysids was related to salinity and temperature, and that the largest number of eggs was 

produced at a temperature of 25oC and 31‰ salinity. Because it is an important measure of reproductive 

success, any reduction in brood size can be interpreted as an indication of reproductive toxicity (Khan et al. 

1992). However, brood size is also directly related to female size. Therefore, reduced fecundity in 

response to exposure to EDCs needs to be carefully evaluated to distinguish direct interruption of 

reproductive processes from a simple reduction in growth. Khan et al. (1992) also stated that the use of 

fecundity without supporting parameters to indicate reproductive impairment is not advisable, because 

fecundity is labor-intensive to determine, requires trained personnel, and ignores toxic effects on males. 

The most likely effect of contaminants is a reduction in fecundity (Hollister et al. 1980, Lussier et al. 

1985), which in some cases is the only response to contaminant exposure (Lussier et al. 1999). 

Contaminant exposure can also result in the abortion of broods (Gentile et al. 1983). Lussier et al. (1999) 

reported, but could not explain, a seasonal difference in fecundity: more eggs were produced in the fall 

than in winter or summer. This phenomenon should be considered when comparing tests conducted at 

different times of the year. Lussier et al. (1999) concluded that fecundity was nonetheless a sensitive and 

useful endpoint. 

6.2.5 Intersexuality and Sex Determination 

195. Exposure to EDCs can result in profound alterations in the reproductive characteristics of 

populations, expressed as physiological or morphological changes in individuals. For example, the most 

commonly reported phenomenon is a condition, pseudohermaphroditism, in which female mollusks 

develop male reproductive structures in response to exposure to tributyltin (LeBlanc et al. 1999). Among 

the Crustacea, cases in which individuals showed intersexuality have been reported for several different 

taxa (see references cited in LeBlanc et al. 1999). Mees et al. (1995) reported intersexuality in natural 

populations of N. integer collected in northern Europe. Ford et al. (2004a, 2004b) found intersexuality in 

males and females of Echinogammarus marinus, an estuarine amphipod. Ford et al. (2004a) found that the 

polluted sites had a higher incidence of intersexuality than reference sites in Scotland. When compared to 

“normal” individuals, intersex amphipods matured more slowly and showed reduced fertility and fecundity 

(Ford et al. 2004b). Reduced fitness was attributed to difficulties in mating resulting from the larger size 

of intersex individuals and the abnormal morphology associated with the condition. However, links 

between such phenomena in crustaceans and EDCs have not been established (LeBlanc et al. 1999, Ford et 

al. 2004a, 2004b). Regardless, Ford et al. (2004a) identified a distinctive morphometry associated with 

intersex males and suggested that it might be a useful biomarker of endocrine disruption. In some cases, 

abnormal sex ratios could be the result of EDC exposure, as has been seen in wild populations of copepods 

(Moore and Stevenson 1991, 1994). Another field observation was made by Bareau and Grecian (2003), 

who observed intersexuality in the amphipod Corophium volutator (Pallas) at four sites of the Bay of 

Fundy, Canada. They found that intersex individuals had both male and female sexual appendages, e.g. the 

second antenna (which is longer in males than in females) was of intermediate appearance, and that no 

intersex individuals carried eggs. The origin of the intersexuality in this study was unknown. Further, 
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studies of daphnids have demonstrated that the population sex ratios found under good environmental 

conditions may be altered by juvenile hormone analogs (Olmstead and LeBlanc 2002, Tatarazako et al. 

(2003). Daphnids reproduce parthenogenetically, producing female young when environmental conditions 

are favorable, but produce male offspring when conditions become unfavorable. Five juvenile hormone 

analogs were shown to alter normal sex ratios in D. magna by stimulating the production of males 

(Tatarazako et al. 2003). 

196. Relatively few field studies have been conducted addressing endocrine disruption to organisms 

found in estuaries, which have been shown to contain a variety of contaminants including sewage, 

chlorinated hydrocarbons, metals, and radioactive materials. Some of these contaminants have endocrine 

disruption potential (e.g., sewage effluents containing steroidal estrogens, DDT and its metabolites, and 

TBT) (Oberdörster and Cheek, 2001). 
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7.0 DESCRIPTION OF POTENTIAL MODES OF ACTION REFLECTIVE OF 

REPRODUCTIVE AND DEVELOPMENTAL IMPAIRMENT 

7.1 RESPONSE TO ECDYSTEROID AGONISTS AND ANTAGONISTS 

197. Concern has often been expressed in recent years about the disruption of endocrine systems in 

aquatic organisms by the action of organic and inorganic contaminants (e.g., Snyder and Mulder 2001; 

Depledge and Billinghurst 1999; Fingerman 1997). In a review by Hutchinson et al. (1999a), it was 

suggested that based on estimated figures, the concentration of EDCs expected to be introduced to the 

United States’ aquatic environment could be as high as 2.16 ng/L for 17α-ethinylestradiol-derived 

contraceptives, and 41.5 ng/L for conjugated estrogens used in hormone-replacement therapy. Although 

there has been considerable research conducted on the health of fish exposed to EDCs, there is little 

information available for crustaceans (Baldwin et al. 2001, Hutchinson et al. 1999a, b). 

198. Developmental abnormalities and toxicity to daphnid embryos were noted (at levels far below 

concentrations causing toxicity to maternal organisms) when the maternal organisms were continuously 

exposed and also when the embryos were collected from unexposed parents and exposed directly to 

testosterone. These developmental abnormalities and delays in molt frequency of neonates were mitigated 

when the daphnids were co-exposed to 20-hyrdroxyecdysone. These findings suggest that testosterone 

may function as an anti-ecdysteroid in crustaceans (Mu and LeBlanc 2002a). This study suggests that 

testosterone was able to block the activity of 20-hydroxyecdysone when both steroids were provided. 

Ecdysteroids for the early embryo stage originate from the mother and are transferred to the egg. 

Ecdysteroids present in the late embryo are in part synthesized from the embryo. Thus, when the embryo 

is exposed to an anti-ecdysteroid antagonist it could affect both pools (maternal and embryo) of 

ecdysteroids. This could be manifested by early- and late-stage abnormalities. When direct exposure of 

embryos is conducted, it would have no impact on the maternal pool of ecdysteroid resident in the embryo, 

but it would affect production of ecdysteroids within the embryo itself, most likely noted by late-stage 

developmental abnormalities. 

199. In the environment most chemicals are present as mixtures, yet little research has been conducted 

to evaluate the potential endocrine disruption of complex mixtures. Toward that end, the toxicity of a 

chemical mixture of fenarimol and testosterone was evaluated using Daphnia magna to ascertain if 

combined exposure would result in greater than additive toxicity. By itself, fenarimol causes late-stage 

developmental abnormalities in embryos while testosterone interferes with early and late-stage embryo 

abnormalities (Mu and LeBlanc 2004). Fenarimol is a known ecdysteroid synthesis inhibitor of 

endogenous hormones; when combined with testosterone, a known ecdysteroid antagonist, fenarimol 

effectively inhibited hormone synthesis, paving the way for testosterone to bind to ecdysteroid receptors. 

By exposing only embryos to fenarimol, this would result in perturbations in late embryo development 

since maternal ecdysteroids would be present and active, whereas exposure of testosterone to isolated 

embryos would cause both early- and late-stage developmental abnormalities. Results showed that 

fenarimol increased the toxicity of testosterone, while testosterone had no effect on the toxicity of 

fenarimol. Additionally, a model was used to predict combined effects using algorithms. The predictive 

model was very effective in estimating the joint toxicity of these compounds for the independent action 

and synergy of both compounds. Further studies are needed to evaluate the role of complex mixtures on 

crustacean endocrine systems. 

200. The endocrine and reproductive effects of EDCs mimic the effects of natural hormones, 

antagonize the effects of hormones, alter the pattern of synthesis and/or metabolism of hormones, and/or 

modify hormone receptor levels (Depledge and Billinghurst 1999). The ability of some environmental 
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contaminants to bind to steroid hormone receptors as agonists or antagonists in a recognized mechanism of 

toxicity to endocrine-related processes has been documented (LeBlanc and McLachlan 1999). 

201. The literature is vague with respect to gender differences from exposure to ecdysteroids. CuzinRoudy and Saleuddin (1989) discussed possible differences in effects to male and female mysids, Siriella 

armata. This study showed that secondary vitellogenesis starts at the beginning of the molt cycle for this 

organism, when ecdysteroid levels are low. There is a striking difference between males and females at 

this point: in females, ecdysteroid levels were 10 times higher than those in males, but the response of the 

epidermis for molt preparation was the same. Females also had much higher levels of 20-E, ecdysone, and 

high polarity products, which are probably linked to the storage of ecdysteroids in oocytes during 

secondary vitellogenesis. Embryonic and post-embryonic development occurs in the marsupium of the 

females. Juveniles are released shortly before ecdysis, after which the adult female lays a new batch of 

eggs in the marsupium. A secondary vitellogenic cycle starts for a new batch of oocytes on the second day 

of the female molt cycle. Secondary vitellogenesis is strictly linked to the molt cycle. During 

development, gonads and gonoducts differentiate before the appearance of secondary sexual characteristics 

(Cuzin-Roudy and Saleuddin 1989). 

202. Macrocrustaceans are in general fast-growing and slow-breeding organisms. Integration between 

molting and reproduction is a physiological necessity in females. Ecdysteroid, the chief hormone in 

molting, is thought to be involved with control of female reproductive activities. However, this is 

controversial. Investigations using amphipods have shown that levels of vitellogenin fluctuate with 

hemolymph ecdysteroid levels (see for example Cuzin-Roudy and Saleuddin 1989, Depledge and 

Billinghurst 1999). 

203. In crustacean females, sequestered ecdysteroids may be passed on to the eggs for possible 

elimination and to function as morphogenetic hormones partaking in the control of embryogenesis and 

early development. The ovary in many crustaceans accumulates ecdysteroid for possible use during 

embryogenesis (Subramoniam 2000). Molting and reproduction are more evident in the female, because 

vitellogenesis is the central event of the female reproductive cycle along with secretion of a new cuticle 

during molting. Hormones play a role in the nutritive supply for molting and vitellogenesis. The ovaries 

eliminate ecdysteroids by forming ecdysonic acid as a necessary way to eliminate ecdysteroid in the eggs 

and embryos. They also form conjugates as a means of elimination. In embryos, there are concentrations 

of the three ecdysteroids – i.e., ecdysone, 20-E, and PoA – and their conjugates. There are fluctuations in 

embryonic ecdysteroids, as evidenced by one shrimp species, Sicyonia ingentis, in which the eggs after 

spawning contain low levels of ecdysteroid. The levels then rise through development, probably by the 

synthesis of this hormone by the embryo’s Y-organ. The endogenous accumulation of ecdysteroid within 

the ovary is also known to function in the induction of meiotic maturation of the oocyte (Subramoniam 

2000). 

204. In one experiment, Subramoniam (2000) removed eggs from the pleopods of the freshwater 

prawn, Macrobrachium nobilii. The release of eggs quickened the next molting and reproductive cycle. 

In another experiment, Subramoniam (2000) found that although the ovarian cycle begins during the 

intermolt stage, vitellogenesis (serum levels) progresses into the next premolt stage. Premolt starts with 

the release of the larvae, and the next spawning occurs after ecdysis. Among penaeid shrimp, free 

spawning occurs during the premolt stage, followed by ecdysis (Subramoniam 2000). There are few in 

vitro studies available that focus on specific mechanisms involved in disruption in arthropods. The 

ecdysone receptor is in the same gene family as the thyroid receptor found in vertebrates. 

205. In another experiment conducted by Bodar et al. (1990), daphnids were exposed to varying 

concentrations of cadmium and separately to exogenous ecdysone and 20-hydroxyecdysone to monitor any 

changes to the molt or reproductive cycle and to evaluate the role of ecdysteroids relative to molting and 
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reproduction. The study showed dose-dependent effects on molting and reproduction for the ecdysone and 

cadmium exposures. The effects of the higher ecdysteroid concentrations included unsuccessful 

exuviations, incomplete molting, and eventually death such that animals died before the age of potential 

reproduction was reached. The daphnids did not molt after treatment to ecdysteroids, and it was speculated 

that that they spent a disproportionate amount of energy on molting which negatively impacted the 

reproductive physiology. 

206. For the cadmium exposures, a dose-dependent effect of cadmium on ecdysteroid titers was 

observed. At cadmium concentration of 5 μg/L there was an increase in levels above control (~210 pg 

ecdysone eq/mg dry weight) after 2-day exposure; these high levels declined after 8 days to levels 

approaching the control. As the cadmium concentration increased to 20 μg/L, there was a linear increase 

in ecdysone concentrations to around 750 pg ecdysone eq/mg, which is three times higher than controls. A 

pronounced decline in growth occurred under cadmium exposure. Also, a stimulatory effect on steroid 

hormones was seen such that increasing cadmium levels corresponded with increasing hormone titers. 

This stimulatory effect of cadmium on ecdysteroids has been observed for other organisms. Two theories 

are put forth: either the cadmium caused increased ecdysteroid levels which led to molt and reproductive 

impairment, or the cadmium interfered with the metal regulatory system through the metallothioneins and 

metalloenzymes that are involved in the molt cycle. 

207. Incubation of ecdysteroid synthesis tissues in vitro is a method to detect endocrine modulators on 

molting hormone synthesis. This method was investigated as a potential biomarker using the midge C. 

riparius. This method can be used to determine if a particular chemical causes endocrine disruption at the 

sub-organismal level. Detailed methods for conducting ecdysteroid biosynthesis and subsequent 

measurement using radioimmunoassay techniques are described in Hahn and Schulz (2002). Male and 

female midges responded quite differently to this technique such that ecdysteroidogenic activity was 

significantly increased above controls for the males, while the opposite trend was found for the females. 

Also, exposed males developed faster than controls, whereas the treated females showed slower 

development than the controls. Even further, there is speculation that ecdysteroid metabolism is regulated 

by different processes in males and female midges during the fourth stage of larval development. 

208. Block et al. (2003) used two crustaceans—a copepod (Amphiascus tenuiremis) and an amphipod 

(Leptocheirus plumulosus)—to determine ecdysteroid concentrations at different life stages. They also 

used a method known as fluorescence-based enzyme immunoassay (EIA) to quantify and compare 

ecdysteroid titers in such small organisms. Detailed methods for conducting the assay, including tissue 

collection and extraction of small sample volumes, is described. The overall synthesis, regulation, and 

metabolism of ecdysteroids used within and across species is most likely associated with variations in 

growth (molting), mating, and life cycles. Therefore, it is important to have a precise measure of 

ecdysteroid levels capable of detection at the femtomolar (10−15 molar) level. 

209. In summary, detailed studies of crustacean response to ecdysteroids are lacking. Future studies 

that address sequence determination of vitellogenic genes and their hormonal activity could provide 

interesting insight into the vitellogenic process in this taxonomic group. The study of genomic response of 

ovarian maturation following ecdysteroid exposure deserves further attention and research. Synergistic and 

antagonistic actions of the X-organ sinus-molt and gonad-inhibiting neuropeptides, and the mandibular 

organ control over molting and reproduction are other areas requiring further research for use of 

crustaceans for EDC testing in the future. 

7.2 RESPONSE TO VERTEBRATE-TYPE ANDROGENS 

210. Vertebrate-type steroidal androgens have been measured in some crustaceans, but androgen 

receptors have not been documented. Presently, the androgenic gland has been identified only for 
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malacostracans (Block et al. 2003). Steroidal androgens can function directly as hormones in ways that do 

not require receptors, or they can be present as inactive components of steroid metabolic pathways 

(LeBlanc and McLachlan 1999). There is currently only limited published research that evaluates the 

androgenic hormones and their effect specifically on crustaceans. Administration of testosterone to shrimp 

has resulted in hypertrophy and hyperplasia of the androgenic gland. The androgenic gland is associated 

with the testis, and is responsible for the secretion of the androgenic hormone. This hormone is nonsteroidal and is responsible for masculinization. Testosterone administered to shrimp and crab results in 

the increase in testis size and in the conversion from ovaries to testes in females (LeBlanc and McLachlan 

1999). Experiments using Daphnia magna showed that testosterone, acting as an antagonist to 20-

hydroxyecdysone, caused toxicity to neonates by interfering with the control of molting by ecdysteroids 

(Mu and LeBlanc 2002a). These studies, and the recent discovery of endogenous testosterone in Neomysis 

integer (Verslycke et al. 2002), suggested that studies designed to measure androgenic effects in 

crustaceans could be conducted. To date, an androgen receptor has not been identified nor cloned in 

crustaceans; research should be directed at identification and characterization of this receptor to aid in 

determining the usefulness of sex steroids as an evaluation tool for crustaceans (Verslycke et al. 2004a). 

7.2.1 Endpoint Sensitivity 

211. Vertebrate-type steroid hormones are found in the ovaries, testis, mandibular organ, and 

hemolymph of some crustaceans. Many of these steroid hormones exhibit fluctuations during gonadal 

development, suggesting a role in reproduction of crustaceans. 

212. The identification of physiological targets of EDC in invertebrates is the approach taken by 

LeBlanc and McLachlan (1999). One example is that diethylstilbestrol and endosulfan have been shown to 

inhibit molting in immature daphnids, but to have no effect on the mature animals’ fecundity. These 

effects may indicate that chemicals that are estrogenic to vertebrates could affect molting and reproduction 

in crustaceans, interfering with the proper function of the ecdysone receptor. In a study designed to 

examine antiandrogens, Daphnia magna was exposed to the compound cyproterone acetate, to determine 

whether it interferes with the androgen receptor as it does with vertebrates (LeBlanc and McLachlan 

(1999). The results showed impairment to growth. The exposed organisms were smaller, and there was a 

reduction in number of offspring. The latter was most likely due to the smaller size of the organisms, 

which would not have been able to accommodate a more normal number of brood in the pouch. The 

effects of steroidal androgens and chemicals that cause metabolic androgenization are consistent with 

interference to the delivery or packaging of nutrients into the developing eggs. Ecdysteroids, juvenoids, 

progestogen, and crustacean androgens have all been shown to influence vitellogenin or lipid production in 

arthropods. Androgens may interfere with one or more of the hormonally regulated processes that provide 

nutrients to embryos. 

213. Baldwin (1997, 1998) conducted a series of experiments using the daphnid. During one 

experiment, the daphnid was exposed to 4NP, which resulted in changes in rates of elimination of 

testosterone and a corresponding decrease in glucose-conjugated testosterone, and an increase in the rate of 

production of various androgenic derivatives of testosterone. This is called metabolic androgenization, 

which is found to reduce fecundity of exposed daphnids associated with developmental abnormalities and 

high mortality of offspring. Results from a separate experiment with exposure to NP revealed no 

significant evidence of changes in steroid elimination processes, except at the highest concentration, which 

reduced elimination of glucose- and sulfate-conjugates and increased elimination of oxido-reduced 

derivatives. Effects were seen at sublethal levels for 4NP and approaching acute levels for nonylphenol. It 

has not yet been determined whether there is an androgen receptor in crustaceans; therefore, more studies 

are needed to determine the functional role of steroidal androgens. 
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214. It is possible that endogenous androgens may be precursors to other hormones, and that large 

doses of exogenously added androgens could elicit activity through other receptors. In crustaceans, 

testosterone is converted to androstenedione at various rates (LeBlanc and McLachlan 1999). Future 

studies may reveal that the conversion is affected by age, reproductive state, or photoperiod. It is possible 

that alteration in testosterone metabolism could serve as a biomarker, because effects are observed at 

concentrations less that those eliciting reproductive response. 

215. Verslycke et al. (2003b) examined the elimination rates of testosterone by monitoring a specific 

set of metabolites (polar hydroxylated, nonpolar oxido-reduced, and glucose- and sulfate-conjugated). 

Various theories surround imposex of neogastropods exposed to TBT: inhibition of the cytochrome P450 

system; blocking phase II sulfate conjugation; and interference directly with neurohormonal system, 

leading to changes in steroid titers. Further work is needed to confirm which mechanism(s) is responsible. 

7.2.2 Gender Differences 

216. Currently, there is no documented research that discerns gender differences in mysids or other 

crustaceans as a result of androgenic-type hormone response. Detailed mechanistic and anatomical studies 

would need to be conducted on mysids to ascertain whether differences in gender relative to EDCs can be 

observed. 

217. In the 1980s, the condition of imposex (the imposition of male sex organs including a penis and 

vas deferens) was observed with increasing frequency on marine gastropods exposed to tributyltin (TBT) 

(Depledge and Billinghurst 1999). The mode of action of TBT giving rise to imposex is currently under 

investigation. Female snails exposed to TBT have elevated testosterone in the hemolymph, and injections 

of TBT into females induced penis formation (Depledge and Billinghurst 1999). Lee (1991) thought that 

many of the observed effects in mollusks are related to enzymes involved in TBT metabolism. Inhibition 

of a cytochrome P450-dependent aromatase (which normally converts 17α-estradiol to testosterone) could 

result in the accumulation of testosterone, which would otherwise be metabolized. 

218. Studies of the shrimp, Palaemon serratus, showed that eyestalk ablation resulted in rapid 

maturation of the ovaries (reviewed by Fingerman et al. 1998). It was later shown that this effect is caused 

by the sinus gland containing a gonad-inhibiting hormone (GIH). This system is present in male 

crustaceans as well, and eyestalk ablation to induce gonadal maturation is a common practice on shrimp 

farms worldwide (Fingerman et al. 1998). The presence of gonad-stimulating hormones (GSH) was 

demonstrated in decapod crustaceans. In female crustaceans, the GIH and GSH acted directly on ovaries, 

which then secreted the ovarian hormone. Ovaries are a source of ovarian hormone, which induces the 

development of secondary female sexual characteristics. In male crustaceans, GIH and GSH acted on the 

androgenic gland. Two experiments were conducted to determine the role of the androgen gland using 

Macrobrachium rosenbergii. When the androgen gland was removed, the male became feminized, and 

when the androgen gland was implanted into a female, the female became masculinized (Fingerman et al. 

1998). 

219. In their review, Fingerman et al. (1998) reported that parasitism of crustaceans by rhizocephalans 

induced castration. The castration of the males often involved additional impairment to testicular function 

by modification of the secondary sexual characteristics, causing the males to take on female appearance. 

For example, the narrow male abdomen of crabs became wider, resembling that of a female. Several 

authors, such as Fingerman et al. (1998) have reported that in male shore crabs, Carcinus maenas, 

spermatogenesis occurred nonetheless in the testes of specimens found with feminized abdomens. 
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7.3 RESPONSE TO VERTEBRATE-TYPE ESTROGENS 

220. Billinghurst et al. (1998, 2000, 2001) examined the effects of two estrogens, 4-n-nonylphenol and 

17β-estradiol on larval settlement and the production of a larval storage protein (cypris major protein, 

CMP) in Balanus amphitrite. Cyprids use CMP during settlement and the early post-settlement 

development. Because CMP is structurally related to vitellin, which is analogous to vitellogenin, it can be 

used as a biomarker of estrogen exposure in lower vertebrates. The expectation in these studies was that 

cyprid settlement might be affected by the stimulation of CMP synthesis after larval exposure to 

environmental estrogens. The results of the 1998 study, however, showed reduced settlement after 

exposure to both estrogens, but that the cause was not related to endocrine disruption. The second study 

(Billinghurst et al. 2000) measured levels of CMP and found that they were elevated after exposure of 

nauplii to low levels of the estrogens. The third study (2001) measured effects of 4-n-nonylphenol and 

17β-estradiol to larvae of Eliminius modestus (nauplii and cyprids). Specifically a disruption of the timing 

of larval development was noted, but this disruption was not consistent for different trials of this same 

experiment. This is in contrast to the 1998 study, but the studies were carried out at different times of year 

and with species that have different reproductive cycles. The variable response of different species to the 

same chemical reinforces the concept that development of larval crustaceans is subject to precise 

mechanisms and exposure to contaminants may depend on season and species. Further, this study showed 

that varying the timing of exposure of 4-n-nonylphenol and 17β-estradiol or the duration was critical. For 

example, organisms exposed for 12 months were significantly smaller than control organisms. As reported 

for other studies, Billinghurst et al. (2001) found that intermediate concentrations of NP are more 

disruptive than higher or lower concentrations. 

221. Hutchinson et al. (1999a, 1999b) found that exposure to several steroids had no effect on the 

survival and development of copepod (Tisbe battagliai) nauplii and cautioned against extending the 

reported effects of steroid exposure in some species of crustaceans to the group as a whole. At about the 

same time, Bechmann (1999) showed that high levels (>62 μg/L) of nonylphenol were acutely toxic to T. 

battagliai, but that exposure to a low level (31 μg/L) did not affect any of the measured life-table 

parameters (survival, sex ratio, fecundity) measured. Breitholtz and Bengtsson (2001) did not find 

evidence of endocrine disruption in the harpacticoid copepod Nitocra spinipes after exposure to the 

estrogens 17β-estradiol, 17α-ethinylestradiol, and diethylstilbestrol. This contrasts with another study that 

exposed Hyalella azteca to 17α-ethinylestradiol at concentrations of 0.1 μg/L to 10 μg/L in a 

multigenerational experiment (Vandenbergh et al. 2003). Results showed that F1 males exposed from 

gametogenesis to adulthood developed significantly smaller second gnathopods; this response generated a 

U-shaped dose response curve suggesting a receptor-mediated response because effects were found at 

lower concentrations but masked at higher concentrations. Post F1-generation males exhibited histological 

aberrations of the reproductive tract (i.e., hermaphroditism, disturbed maturation of germ cells, and 

disturbed spermatogenesis); again these responses were more pronounced at the lower concentrations, 

suggesting a receptor-mediated response. Also noted, but not statistically significant, was that the 

populations exposed to 17α-ethinylestradiol for more than two generations tended to favor females. 

222. Additional experiments conducted by Oberdörster et al. (2000) using Palaemonetes pugio in a 6-

week pyrene exposure showed a significant increase in vitellin at 63 ppb and a significant increase in 

embryo mortality at 63 ppb. The increase in VTN could be a countermeasure against lipophilic 

compounds such as pyrene, because vitellins may be able to bind lipophilic compounds and transfer them 

to developing embryos. 

223. In the shrimp, Penaeus monodon, 17α-estradiol and progesterone in free and conjugated forms 

increase in the ovary during vitellogenesis (Fairs et al. 1990). Metabolic precursors such as pregnenolone 

and dehydroepiandrosterone also increase and show a peak during the major vitellogenic stages, suggesting 

a pathway in crustaceans that is similar to that in vertebrates. Fairs et al. (1990) also reported that 17α-
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estradiol and progesterone levels in the hemolymph showed fluctuations resembling that of serum 

vitellogenin levels during ovarian maturation. Estrogen could possibly control the stimulation of yolk 

synthesis, whereas the progesterone could control the prophasic meiotic maturation, causing germinal 

vesicle breakdown in the post vitellogenic oocytes. Exogenous injections of steroidal hormones induced 

vitellogenesis in the prawn, P. japonicus. In a study of a marine shrimp, P. semisulcatus, it has been 

shown that both the vitellogenin synthesis in the hepatopancreas and vitellin synthesis on the oocytes are 

coded by one gene (Subramoniam 2000). 

224. Female crustaceans synthesize and secrete the protein vitellogenin into the hemolymph at the 

onset of oogenesis. There are a limited number of enzyme-linked immunosorbent assays (ELISA) that 

have been developed for measurement of vitellogenin or vitellin (VTN) in crustaceans but all are designed 

for detection in larger decapod species. A new method has been proposed by Volz and Chandler (2004) 

for measurement of VTN in small microquantities in a sediment-dwelling copepod, Amphiascus 

tenuiremis. This ELISA uses VTN-specific polyclonal antibodies from Leptocheirus plumulosus, which 

show specificity toward female copepod proteins. Quantities of purified VTN, used as a standard, were 

collected from grass shrimp eggs because it can be collected in relatively large quantities and it reacts well 

with anti-VTN antibodies from L. plumulosus. The results using this ELISA showed significant 

discrimination between gravid females and male samples. The final working range for the ELISA was 

from 31.25 to 1000 ng/mL with intra-assay and interassay CVs of 3.9 and 16.8%. Further, the detection 

limit was 2 ng/mL and the ability to quantify VTN in small numbers (four or more) copepods makes this 

ELISA a promising tool for further research for monitoring endocrine activity of toxicants to copepods and 

other crustaceans. 

225. Studies (Pounds et al. 2002) to determine mode of action of several selected natural and synthetic 

steroids and xenoestrogens were examined using a combination of the Tisbe battagliai life-cycle test and 

also the BII haemocyte line of Drosophila melanogaster, which is a screening assay to examine the agonist 

and antagonist effect of compounds to the cell line. One steroid tested, 20 HE, demonstrated reproductive 

effects on Tisbe battagliai at 26.9 μg/L and also demonstrated agonist activity to the ecdysteroid receptor, 

suggesting that the in vivo response was mediated via the receptor. The other steroids tested showed no 

response to either assay, indicating that even at high concentrations the synthetic and natural vertebrate 

steroids do not interact with the ecdysteroid receptor. Receptor-binding assays provide information 

relative to specific interaction of a compound to the endocrine system but do not give predictive 

information about how this compound will affect the whole organism. 

7.4 RESPONSE TO OTHER HORMONAL DISTURBANCES 

226. In their review paper, Fingerman et al. (1998) described other hormonal responses and 

disturbances in crustaceans, such as color-changing hormones, retinal pigment hormones, pericardial 

hormones, and blood glucose hormones. Each of these will be briefly described below, relative to 

crustaceans in general. Specific hormone disturbances to mysids as well as endpoint sensitivity and gender 

differences await further study. 

227. The sinus gland is the storage and release site for color-change hormones, among others. The 

sinus gland is located proximal to the eye and lies next to the large hemolymph sinus (Fingerman 1997). 

For sessile crustaceans, the sinus gland is located in the head close to the optic centers. Investigators 

believe that 90% of axonal terminals that compose the sinus gland belong to neurons whose cell bodies lie 

in the medulla terminalis X-organ. Therefore, the medulla terminalis X-organ sinus-gland complex is 

similar to the vertebrate hypothalamo-neurohypophyseal complex (Fingerman 1997). 

228. Color change is affected by cells called chromatophores, which are located in the integument. 

They are responsible for color change through their dispersion and aggregation. In an early experiment, 
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the hemolymph of a dark prawn specimen was transferred to a pale one (Fingerman 1997). When this 

organism was kept on a white background, it turned dark. The researcher then cut through the exoskeleton 

to sever any peripheral nerves that should innervate the chromatophores to determine whether color change 

was related to the endocrine system or to the nervous system. The incision had no effect on color change. 

Histological examination failed to show any innvervation of the chromatophores, which lead to the 

conclusion that color changes of this prawn are due to hemolymph-borne pigment concentrating 

substances. 

229. For a variety of decapod crustaceans, chromatophores that cause integumentary color changes are 

controlled by antagonistically acting pigment-dispersing and pigment-concentrating neurohormones. For 

example, in the fiddler crab Uca pugilator, the neurohormone, 5-HT-serotonin triggers the release of red 

pigment-dispersing hormone, but has no effect on the black chromatophores (Fingerman 1997). The black 

chromatophores are triggered by norepinephrine, which releases a black pigment-dispersing hormone 

(BPDH). Studies have shown that the eyestalks of U. pugilator contained four times as much BPDH as did 

the control after exposure to naphthalene, due to naphthalene’s inhibition of norepinephrine release. The 

opposite mechanism was observed for cadmium (Fingerman 1997). The eyestalks of control organisms 

contained three times more BPDH than did the cadmium-exposed crabs, which indicated that cadmium 

inhibited the synthesis of BPDH. 

230. Retinal tissue contains pigments that control the amount of light striking the rhabdom (the 

photosensitive part of each ommatidium that composes the compound eye) through changes in position. 

Three types of retinal pigments have been categorized: the distal, proximal, and reflecting (Fingerman 

1997). Migration of the distal pigment is controlled by the light-adapting hormone (LAH) and the darkadapting hormone (DAH). Most studies of retinal pigments use the distal pigment, because techniques for 

its use are noninvasive. Several studies were conducted using Palaemonetes vulgaris. In one experiment 

conducted by Fingerman in 1959, this species was kept under constant illumination and then injected with 

extracts of eyestalks or sinus glands (described in Fingerman 1997). Because there is an initial lightadapted response followed by a dark-adapted response, Fingerman was able to induce a dark-adapting 

response from an organism kept under constant illumination, suggesting that eyestalks contain both LAH 

and DAH. 

231. The sinus gland contains the source of CHH, which causes elevation of blood glucose levels for 

crustaceans. CHH is similar in structure to MIH. It has been determined that MIH and CHH also show 

similar activity (Fingerman 1997). Future studies should address the specific roles of these hormones in 

mysids and other crustaceans. Exposure of the freshwater prawn Macrobrachium kistnensis, and several 

species of crabs, to cadmium caused hyperglycemia. Similarly, exposure of U. pugilator to naphthalene 

caused hyperglycemia, although the mode of action is apparently different for the two compounds. 

Cadmium inhibits CHH synthesis, whereas naphthalene stimulates CHH synthesis; 5-HT apparently 

triggers release of CHH (Fingerman 1997). 

232. Pericardial organs lie in the venous sinus that surrounds the heart, and the axon terminals could 

be part of the neuroendocrine system that releases hormones affecting the heart. Experiments showed that 

the hearts of three species, Cancer pagurus, Homarus vulgaris, and Squilla mantis responded to pericardial 

organ extracts with increase in both frequency and amplitude of the heart beat. Efforts to identify 

substances in the pericardial organs have revealed 5-HT, dopamine, and octopamine (Fingerman 1997). 
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7.5 HOW TO CONFIRM MODES OF ACTION – THE USE OF BIOCHEMICAL 

MEASURES 

233. A number of biochemical measures are explored below as possible endpoints for EDC-exposure 

tests: metabolic disruption, steroid metabolism, vitellogenin induction, hormone levels, RNA:DNA ratios, 

protein levels, and the levels of cytochrome P450 enzymes and of blood glucose. Care must be taken when 

interpreting results of bioenergetic biomarkers such as those mentioned here, because many abiotic as well 

as toxic stressors affect metabolic processes. 

7.5.1 Metabolic Disruption (O:N ratios) 

234. Many of the perturbations expressed in the parameters described in Sections 4.2.4 could be 

related to changes in energy pathways resulting from chronic exposure to contaminants. McKenney (1985, 

1989) and McKenney et al. (1991) showed that exposure to contaminants caused increased respiration rates 

in juvenile mysids, often with as little as 24-h exposure. The increased general metabolic demands related 

to contaminant exposure reduced growth by decreasing the amount of energy available to produce new 

somatic tissues. Normally, energy allocations to metabolism, growth, and reproduction occur, so that if 

there are changes in allocations of energy this may be an indicator of stress attributable to a toxicant or an 

environmental change (e.g., temperature, oxygen levels) (Verslycke et al. 2004a). It is important that 

bioenergetic endpoints include a baseline assessment of what is considered normal for crustacea. Further, 

alterations to energy metabolism can be assessed by using tools like the cellular energy allocation assay. 

Changes in metabolic usage can also be monitored using oxygen and nitrogen ratios, protein and lipid 

content, or the carbon to nitrogen ratios. 

235. One way of predicting effects of abiotic stress on energy metabolism is the cellular energy 

allocation (CEA) assay. Verslycke et al. (2003b) conducted a study using Neomysis integer exposed to 

varying concentrations of tributyltin chloride (10, 100, and 1000 ng TBT Cl/L). Effects were noted using 

the CEA assay such as less overall consumption of energy and lower respiration rates than control mysids. 

These results show that TBT interferes with the energy metabolism by disrupting energy production 

process. Overall there was a loss of protein, lipids, and sugars in the organisms exposed to higher 

concentrations of TBT versus the control. Later, Verslycke et al. (2004b) showed that exposure to 

sublethal doses of nonylphenol and methoprene significantly increased energy consumption, resulting in 

less energy available for reproduction and growth. Using the CEA assay to assess the energy budget 

quantitatively for a particular organism helps to elucidate potential modes of action for toxicants. Methods 

for CEA measurements include spectrophotometric measures of lipid, protein, and sugar (Verslycke and 

Janssen 2002). Additionally, electron transport activities were measured by using INT (piodonitrotetrazolium violet) as an electron acceptor mimic in the electron transport system, which can 

provide a measure of oxygen consumption (Owens and King 1975 as cited by Verslycke and Janssen 

2002). 

236. Increased metabolic demands caused by exposure to contaminants could also impair mysid 

growth and reproductive capability. Exposure to relatively high concentrations (0.072 μg/L, 0.1 μg/L) of 

the pesticide chlorpyrifos caused increased oxygen consumption rates and reduced egestion rates in 

Neomysis integer (Roast et al. 1999). These responses resulted in reduced energy available for growth 

(i.e., lower “scope for growth”). Young mysids typically use high-energy lipids to meet metabolic 

demands, but change to metabolize proteins as they mature, thereby leaving more lipid material available 

for the production of gametes (McKenney 1989). Gorokhova and Hansson (2000) found that a 6% 

decrease in the carbon:nitrogen ratios occurred as juvenile Mysis mixta matured to gravid adults and 

asserted that this was evidence that maturation and reproduction are fueled primarily by lipids. Increased 

metabolic demands caused by exposure to contaminants is met by greater lipid metabolism, which reduces 

the lipids available to meet reproductive needs (McKenney 1985, 1989). These changes in metabolic 
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substrate usage can be measured by monitoring the oxygen : nitrogen (O:N) ratio of test organisms. The 

O:N ratio indicates the relationship between the amount of oxygen consumed by an organism to the 

amount of nitrogen excreted, and shows the relative role protein catabolism plays in the organism’s energy 

budget (Carr et al. 1985; McKenney 1985). Americamysis bahia showed a change toward lipid 

metabolism after only 4 days of exposure to high concentrations of cadmium (Carr et al. 1985). Mysids 

showed increased metabolic demands after four days of exposure to the herbicide thiobencarb (McKenney 

1985) or five days of exposure to the defoliant DEF (McKenney et al. 1991). High O:N ratios occurred 

among maturing mysids exposed to the compound, indicating a shift to lipid metabolism that would have 

reduced the lipids available for gamete production. All three studies concluded that changes the O:N ratio 

measured among test mysids was a sensitive indicator that could provide for the relatively early detection 

of reproductive impacts by contaminants. 

7.5.2 Steroid Metabolism 

237. Although the role estrogens play in crustacean reproduction is not known (Baldwin et al. 1995), 

these steroids are important in other invertebrate groups. Because of the likelihood that crustaceans could 

be exposed to environmental estrogens such as dichlorodiphenyl trichloroethane (DDT), polychlorinated 

biphenyls (PCBs), and nonylphenols (nonionic surfactants), there is the potential for these compounds to 

disrupt steroid metabolism. Baldwin et al. (1995, 1997, 1998), studied the effects of three environmental 

estrogens, diethylstilbestrol (DES), 4-nonylphenol (4NP), and nonylphenol polyethoxylate (NPPE), on the 

steroid metabolism of the freshwater daphnid, Daphnia magna. Their work focused on the disruption of 

the metabolic elimination of testosterone after short- and long-term exposure to the test compounds and 

sought to determine whether such an analysis could be used as an early indication of reproductive 

impairment. They measured differences in the glucose conjugation, sulfate conjugation, hydroxylated and 

reduced/dehydrogenated metabolites of 14C-labeled testosterone in daphnids exposed to sublethal 

concentrations of the test compounds. They found that the different compounds had different effects on 

testosterone metabolism. For example, DES increased glucose conjugation, but did not affect sulfate 

conjugation, whereas 4NP reduced both of these elimination processes. In their two earlier studies 

(Baldwin et al. 1995, 1997), Baldwin’s group proposed that changes in testosterone metabolism could 

provide an early indication of potential reproductive toxicity after sublethal exposure to suspected EDCs. 

However, in their 1998 paper, Baldwin et al. studied NNPE, a nonionic surfactant that degrades to 

nonylphenol, and did not find significant disruption of steroid metabolism after short-term exposure. They 

did report some effects after chronic exposure and postulated that those could have resulted from the 

degradation of NPPE to NP. Therefore, they cautioned that use of short-term exposures as an early 

warning indicator might underestimate chronic effects resulting from bioaccumulation and bioactivation of 

the test compounds. 

238. Verslycke et al. (2002) studied testosterone metabolism in Neomysis integer. Significantly, they 

detected endogenously-produced testosterone in male and female mysids. They also found an anabolic 

steroid, β-boldenone—the first known occurrence of the compound in an invertebrate—in mysids exposed 

to testosterone added to the test medium, although the metabolic pathway by which it is formed was not 

discovered. A vertebrate estrogen, 17α-estradiol was not detected. Verslycke et al. (2002) showed that 

testosterone metabolism in mysids involves phase I (oxido-reduced/hydroxylated) and phase II 

(conjugated) derivatives. The study proposed that changes in invertebrate steroid metabolism might be 

used to indicate exposure to EDCs. To evaluate this idea, Verslycke et al. (2003b) examined the effect of 

exposure to tributyltin on the elimination of testosterone by the mysids. The authors summarized the effect 

of TBT exposure on testosterone elimination by calculating the metabolic androgenization ratio—the ratio 

of the eliminated oxido-reduced products to the polar products (hydroxylated plus conjugated). TBT 

exposure changed testosterone metabolism by inducing reductase activity at low, but not high, 

concentrations and reducing sulfate conjugation, although the mechanisms by which these occurred were 

not identified. To further refine the idea that changes in testosterone metabolism could be used to indicate 
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exposure to an EDC, Verslycke et al. (2004b) subjected individuals of N. integer to seven compounds 

(testosterone, flutamide, 17α-ethinestradiol, precocene, nonylphenol, fenoxycarb, methoprene) suspected 

of having endocrine disruption properties. All seven were very toxic to N. integer with 96-h LC50 values 

ranging from 0.32 mg/L (methoprene) to 1.95 mg/L (testosterone). The effects of sublethal doses of 

nonylphenol and methoprene on testosterone metabolism were investigated and found to be significant. 

Methoprene and nonylphenol affected phase I testosterone metabolism in a manner similar to that of TBT 

(induced reductase activity at low concentrations of the chemical). The two compounds had opposite 

effects on phase II metabolism. Glycosylation was increased at a high (100 μg/L) dose of nonylphenol, but 

was significantly reduced at the same dose of methoprene. The metabolic androgenization ratio showed a 

dose-dependant increase after exposure to methoprene, but after exposure to nonylphenol was significantly 

increased only at the lowest concentration tested, i.e., 10 ng/L. 

7.5.3 Vitellogenin 

239. Depledge (unpublished, cited in Depledge and Billinghurst 1999) found that exposure to 4-nnonylphenol induced the production of vitellogenin in decapods. Vitellogenin production is most likely 

controlled by the hormones primarily involved in molting, although in some crustacean groups this is 

likely not the case (Subramoniam 2000). The molt-inhibiting hormone and 20-hydroxyecdysone (20HE) 

are involved in the production and regulation of vitellins and there are several feedback loops (positive and 

negative) involved. To date, because of the complexity of vitellogenesis, it has not been used as a 

biomarker of endocrine disruption (Oberdörster and Cheek 2001). Little research has been conducted on 

vitellin expression in crustaceans, possibly due to the lack of antibodies, which have been shown to have 

cross-reactivity with other species. Most often polyclonal antibodies are used and available for crustaceans 

with the drawback that these antibodies may cross-react with other nonvitellin proteins if developed from 

impure vitellogenin. One such study, using grass shrimp exposed to pyrene, showed up-regulation of 

vitellin and, when the females were transferred to clean seawater, there was an increase in embryo 

mortality, suggesting that the vitellin can transport lipophilic contaminants to developing embryos. 

240. As pointed out by Verslycke et al. (2004a), to understand the potential effects of xenobiotics on 

reproduction, there must be accurate measures of vitellogenin and vitellin in crustaceans. A few studies 

have been done to characterize and purify vitellin (e.g., Volz et al. 2002), and to develop vitellin ELISAs 

(e.g., Volz and Chandler 2004). Additional studies need to be conducted to evaluate the utility of such 

techniques for ED evaluation. 

7.5.4 Hormone levels 

241. Since crustaceans need to molt to grow and achieve a full adult body size (Charmantier et al. 

1997, Chang 1997), the crustacean body goes through several morphological and functional changes, life 

events which may be sensitive to disruption of chemical substances. As indicated above (section 7.1), 

Block et al. (2003) determine ecdysteroid concentrations at different life stages of the copepod 

(Amphiascus tenuiremis) and the amphipod (Leptocheirus plumulosus). A fluorescence-based enzyme 

immunoassay (EIA) was used to quantify and compare ecdysteroid titers in such small organisms as 

copepods. Ecdysteroids may also be detected by Radioimmunoassay (RIA), a method employing 

radioactive isotopes to label either an antigen or an antibody. The isotope emits gamma rays, which are 

usually measured following removal of unbound (free) radiolabel. This has for instance been reported by 

Chang and O ́Connor (1979) and since then been used for e.g. amphipod ecdysteroid level determinations 

(Jacobson and Sundelin 2006). The detailed methods for conducting the assay, including tissue collection 

and extraction of small sample volumes, could be useful for ED evaluation also in other crustacean species. 

242. Methyl farnesoate (MF) is the unepoxidated form of the insect juvenile hormone (JHIII) and is 

produced in the mandibular organ. Recent studies have suggested that MF functions in a similar manner in 
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crustaceans as JH does in insects (Laufer and Biggers 2001). MF is involved in reproductive maturation 

and in morphological changes most notable from late juvenile stages to adult. MF has high potential as a 

biomarker of chemical activity at the endocrine level, but it is extremely lipophilic, and difficult to isolate 

and assay routinely. Various investigators have conducted eyestalk ablation studies in which crustaceans 

such as shrimp, crayfish, and fiddler crabs were stimulated to reach ovarian maturation. JH analogs such 

as methoprene have inhibited early larval and post larval development in the shrimp Palaemonetes pugio; 

they actually enhanced premetamorphic stage development. 

7.5.5 RNA:DNA ratios 

243. Identification of biochemical changes, such as RNA contents, which are related to growth and 

metabolism (Dahlhoff 2004), can be used to determine if an organism has been exposed to a stress, 

including contaminants (Yang et al. 2002). The rationale is based on the fact that the RNA content of 

tissues or whole organisms consists primarily of ribosomal RNA (rRNA). Consequently, concentration of 

rRNA, at any given time is directly related to the protein synthesis of a cell (Elser et al. 2000). The quantity 

of RNA is directly linked to the growth of the individual (Saiz et al. 1998). For example, in small 

metazoans, which have high metabolic rates of biosynthesis, a high portion of RNA is required (Brown et 

al. 2004). DNA exists in a quasi-constant quantity in a somatic cell and therefore might be used as an index 

of the number of cells (Buckley et al. 1999). Based on this, it is becoming increasingly common to assess 

growth rates in a variety of aquatic animals by using the RNA:DNA ratio (e.g., lobsters: Rosa and Nunes 

2003; krill: Cullen et al. 2003; copepods: Saiz et al. 1998, Campbell et al. 2001, Gorokhova 2003; 

daphniids: Vrede et al. 2002; cirripeds: Desai and Anil 2002). However, in small crustaceans, such as 

daphnids, copepods and decapod larvae, RNA content alone may be a more sensitive endpoint than the 

RNA:DNA ratio due to the growth- and ontogeny-related fluctuations in DNA content (Gorokhova and 

Kyle 2002, Rosa and Nunes 2003, Gorokhova 2003). Also, Dahl et al. (in press) found that in the 

harpacticoid copepod Nitocra spinipes exposed to the pharamaceutical substance Simvastatin, RNA levels 

correlated with regular life table endpoints (i.e. development rate). 

7.5.6 Protein levels 

244. Crustacean larvae show dramatic growth and morphogenetic changes, which obviously may be 

affected by environmental conditions or chemical substances. Studies on the cellular response to conditions 

of physical/chemical stress have played a significant role in diverse areas. Variations in protein contents 

may be due to stress factors, e.g. variations in environmental factors such as temperature (e.g. Verslycke 

and Janssen 2002). Verslycke and Jenssen (2002) found that a decline in protein content during a 4-day 

study in the mysid N. integer was caused by the ability of N. integer to use its protein content as a source 

during starvation or stress. Verslycke and Janssen (2002) also state that the exoskeleton of N. integer

contains large amounts of proteins and consequently large variations in protein contents could be a result 

of molting. Torres et al. (2002) investigated the effects of reduced salinity on larvae of four crustacean 

speices (Cancer pagurus, Homarus gammarus, Carcinus maenas, Chasmagnathus granulata). They found 

that the protein contents were generally lower in the exposed groups in contrast to the control group. Some 

exposed groups even had lower protein contents than at the beginning of the experiment. When exposed to 

stress, an organism may react by producing stress proteins. Stress proteins are a ubiquitous family of 

proteins, present within cells at constitutive levels. Upon exposure to stress, they are induced in order to 

provide protection against cellular damage. These stress proteins interact with other cellular proteins; often 

damaged as a result of stress, and in doing so, minimize chances for these proteins to further engage in 

synthetic or regulatory processes (Feder and Hoffman 1999). Hence, protein levels are highly correlated to 

stress factors of an organism, and may therefore be a useful complement to more traditional 

ecotoxicological endpoints as well as other biochemical markers for ED evaluation. 
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7.5.7 Cytochrome P450 Enzymes 

245. Cytochrome P450 enzymes (CYPs) are commonly occurring proteins that are involved in the 

metabolism (i.e., detoxification) of many exogenous and endogenous compounds (Snyder 2000, Snyder 

and Mulder 2001, Verslycke et al. 2004a). Snyder and Mulder (2001) measured CYP45 levels, a family of 

P450 proteins found in the lobster Homarus americanus, and thought they must be involved in the molting 

cycle, in response to exposure to the pesticide heptachlor, a known EDC. They found that peak levels of 

ecdysteroid hormones, and accordingly, molting, occurred later in heptachlor-exposed larvae than in 

control larvae, indicating that heptachlor disrupts steroid molting hormone metabolism. They also found 

CYP45 levels in lobster larvae exposed to heptachlor on Days 1, 2, or 3 after hatching to be 15 times 

higher than they were in those exposed to control solutions. Levels of CYP45 typically peaked 1 to 2 days 

after exposure, and then decreased. Snyder and Mulder suggested that it could be a useful early biomarker 

of exposure to EDCs, because it showed a dramatic and rapid increase in levels after exposure to 

heptachlor. In their testosterone metabolism study, Verslycke et al. (2002) showed that Neomysis integer

has many P450 enzymes that comprise its complex steroid hydroxylation system. They suggested that 

changes in P450 activity could be used as a biomarker indicating exposure of this mysid to EDCs. 

246. It is known that most P450 activities occur in the hepatopancreas, but other tissues, such as the 

gills, stomach, intestines, and antennal glands, have demonstrated some P450-like roles. Oxidative 

metabolism, most notably of PAHs, has been demonstrated in many crustacean species. A detailed review 

of studies conducted examining the metabolic changes (mostly caused by benzo(a)pyrene) in crustaceans 

can be found in James and Boyle (1998). This paper also offers suggestions of P450 involvement and or 

metabolism relative to drugs, steroids, and pesticides. Several studies have shown the presence of 

testosterone and progesterone in gonadal tissues and hemolymph. The source of these steroids is not well 

known but it is likely that several cytochrome P450-dependent steps are involved. Future directions should 

include genomic information on the crustacean P450 system, the definitive identification of an aH receptor, 

and a best understanding of the regulation of the P450 system, particularly for steroid synthesis. 

247. Changes in cytochrome P450 enzymes, which function in the detoxification of many exogenous 

and endogenous compounds, may be associated with disruption of the hormonally regulated molting 

process and are therefore appropriate to measure in EDC studies. The general procedure for determining 

levels of cytochrome P450 in crustacea involves homogenization of whole animals, centrifugation, and the 

collection of the resulting supernatant. Quantification of the levels occurs via gel electrophoresis. Snyder 

and Mulder (2001) described a specific method for determining CYP45 (a family of CP450 enzymes) 

levels in daphnid crustaceans. 

7.5.8 Glucose Levels 

248. Levels of glucose in crustacean blood (i.e. haemolymph) are regulated by a hormone, crustacean 

hyperglycemic hormone (CHH), which is produced in the sinus gland (Fingerman et al. 1998). Release of 

CHH increases blood glucose levels. Some exogenous compounds have been shown to affect the levels of 

glucose in the bloodstreams of several crustacean taxa, probably be stimulating (e.g., naphthalene) or 

inhibiting (e.g., cadmium) CHH synthesis (Fingerman et al. 1998). Measurement of changes in blood 

glucose levels in crustacea exposed to potential EDCs could be indicative of hormonal perturbation other 

than that associated directly with reproduction or molting. Hyperglycemia is a common response to 

environmental or toxicant interactions, and changes in blood glucose levels in response to presence of 

EDCs may be an indication of interference with hormonal activities (Verslycke et al. 2004a). However, 

many abiotic and toxic responses affect energy metabolism, and researchers must be able to make 

distinctions between natural variations in hormone levels and variations that are caused by EDCs. 
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8.0 RESEARCH NEEDS AND OTHER CONSIDERATIONS 

8.1 RESEARCH NEEDS 

249. The last decade has witnessed several arthropod studies where e.g. vertebrate steroids or 

industrial pollutants with known endocrine potency in mammals have had little or no effects related to 

endocrine disruption (Caspers 1998; Bechmann 1999; Hutchinson et al. 1999a, b, Breitholtz and Bengtsson 

2001). Although e.g. estrogens have been identified in crustaceans (e.g. Novak et al. 1990), their 

functional role has not been established. This is, however, not surprising since arthropods have endocrine 

systems that differ widely from the vertebrates’ (e.g. Hasegawa et al., 1993; Subramoniam, 2000). It has 

therefore been recommend focusing on the unique invertebrate endocrine functions rather than having 

man’s hormone system as a reference when carrying out ecotoxicological research to find potential 

endocrine disruptors in environmental risk assessments (Breitholtz et al. 2006). In fact, it is possible that 

systematic studies including endocrine irregularities in arthropods will reveal further effects by other 

substances than those that give effects in vertebrates. 

250. In a recent review, Sumpter and Johnson (2005) have indicated ten major lessons from endocrine 

disruption in the aquatic environment, which we should learn from in our way forward. At least three 

lessons are of special interest for the current DRP. First, we need to find out what is normal before we can 

say what is abnormal. In other words, we need to obtain baseline information on e.g. endocrinology in 

those groups of species we wish to protect. Second, it must be clear that one animal’s poison may not be 

another’s. The challenge is to predict which groups are the most likely to be selectively targeted by a given 

substance. Third, the last decade has shown that acute toxicity testing cannot detect endocrine disruption. 

Clearly, chronic toxicity tests can give messages entirely different from that of acute toxicity. As a 

consequence, the authors highlight the work conducted within regulatory bodies, in particular OECD, to 

establish and validate new testing procedures. This illustrates the need for development and establishment 

of new arthropod life cycle guidelines, as has been discussed and highlighted in the current DRP. 

251. Hutchinson (2002) has also recommended that for environmental assessments of potential 

invertebrate endocrine disrupters, attention should be given to the use of reference compounds to validate 

new invertebrate test methods, evaluating in vitro data from both mammalian-type and arthropod hormone 

receptor assays, utilize small-scale test methods which do not generate effluent handling problems, using 

relevant environmental concentrations during testing and considering using improved test methods for 

which molecular tools are available. In this context, the rapid progress in the field of genomics, i.e. the 

study of how an individual’s entire genome translates into biological functions, is beginning to provide 

tools that may assist our understanding of how chemicals impact ecosystem health. The term 

ecotoxicogenomics has recently also been proposed to describe the integration of genomics and 

ecotoxicology (Snape et al. 2004). Moreover, since proteins direct biological processes, the physiological 

state of an individual organism is perhaps better reflected by the proteome than the genome. Recently, new 

computational tools have facilitated dramatic advances in proteomics, which provides a comprehensive and 

quantitative picture of protein expression and its changes in an organism. For example, proteomic tools 

have been successfully applied in several invertebrate species, including nematodes (Schrimpf et al. 2001) 

and fruit flies (Vierstraete et al. 2003). The development of proteomics in standard invertebrate species, 

such as D. magna, has been identified as being of major importance for future environmental assessment of 

chemicals (Snape et al. 2004). In fact, increased knowledge about protein expression and its changes upon 

toxic exposure will in itself facilitate the mechanistic understanding of the toxic response, including 

endocrine irregularities. 
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252. Cellular processes, such as protein expression, inevitably determine individual and population 

rates. Hence, the concerted study of possible links between effects at the cellular (i.e. proteomics), 

individual (i.e. life table variables such as growth and fecundity) and population level (i.e. predictions of 

growth rate by the use of life table data) may result in powerful tools to be used for environmental 

assessments of potential endocrine disruptors. Rotchell and Ostrander (2003) have also pointed at that 

these and other molecular techniques, which are routinely used in medical research to identify specific 

genes and proteins, also could be used in the search for endocrine disruptors among aquatic organisms. 

Such tools will meet demands of both mechanistic understanding and ecological relevance (Snape et al. 

2004). Hopefully, if clear links may be established between different levels of biological organization, this 

would also facilitate extrapolations, from lower to higher order of levels, and vice versa. In this context, 

arthropod life cycle tests may be the optimal platform to conduct such extrapolations, which could 

substantially enhance the regulation of potential endocrine disruptors in the future. 

253. Chemicals that interfere with the biosynthetic and/or biotransformation pathways of arthropod 

hormones (such as ecdysteroids) would consequently have impact on individual and population growth 

(Verslycke et al. 2002). Any progress in the development of molecular techniques (both in vitro and in 

vivo) that may help revealing modes of action of chemicals in arthropods should be welcomed. Hutchinson 

(2002) claims that such a mode-of-action approach in e.g. arthropods could have both cost-effective and 

ethical benefits. In this context, it would be necessary to develop biomarkers for the proposed species; 

promising biomarkers could be the induction of vitellogenesis in males, and the inhibition of aromatase in 

females (Depledge unpublished, cited in Depledge and Billinghurst 1999). As more evidence for 

endocrine disruption responses is gathered from experimental research, mechanistic studies would be 

required to determine the specific ways in which chemicals can disturb hormones. That is, it becomes 

important to distinguish between endocrine disruption and metabolic toxicity and to distinguish between 

primary and secondary effects. The final step in the strategy would be to conduct field surveys to detect 

and confirm that endocrine disruption effects actually occur in situ (Depledge and Billinghurst 1999). 

8.2 REGULATORY CONSIDERATIONS 

254. In the EU the legally binding standardized Testing Methods to determine the hazardous 

properties of chemicals are contained in Annex V of Dir 67/548/EEC on the Classification, Packaging and 

Labelling of Dangerous Substances. They play a central role in the EU policy on chemicals control and 

they are referred to in many other pieces of EU legislation (e.g. those related to dangerous preparations, 

pesticides, cosmetics and biocides also refer to these methods). To have standardized Testing Methods for 

chemicals at the EU level is also important for ensuring the Single Market & Free Trade inside the Union 

because the free movement of goods between the Member States is based on the mutual acceptance of the 

risk evaluation made by each country and this, in turn, relays on the mutual acceptance of the data 

generated when testing the chemicals. Obviously the same can be said at a global level, for this reason, the 

work is closely linked and co-ordinated with the parallel OECD Test Guidelines Programme. In general, 

the EU adopts the OECD Test Guidelines principally unchanged. Today there are no internationally agreed 

(i.e. OECD, EU) test methods to determine the potential environmental effects of endocrine disrupters. 

Under the existing legislation within the EU, the assessment of the toxic potential of a chemical is 

dependent on the type of chemical. For example, pesticides and biocides have separate legislation 

governing their testing and assessment. This existing legislation, while accounting for detrimental 

endocrine related effects on reproduction or disease states such as cancer, does not use disruption to the 

endocrine system as an endpoint per se. However, in 1999 the European Commission adopted a 

Community Strategy for Endocrine Disrupters (COM(1999)706) containing a number of short- and longterm actions. Research and development of agreed test methods for endocrine disrupters are prioritized in 

the Strategy. Furthermore, in the proposal for a new EU regulatory framework for chemicals, REACH 

(Registration, Evaluation and Authorisation of Chemicals) an authorisation system for uses of substances 

and the placing on the market of substances for such uses is established for the substances of very high 
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concern. Substances of very high concern are defined as: substances that are category 1 and 2 carcinogens 

or mutagens; substances that are toxic to the reproductive system of category 1 and 2; substances that are 

persistent, bioaccumulative and toxic or very persistent and very bioaccumulative; and substances such as 

endocrine disrupters, which are demonstrated to be of equivalent concern. The authorisation provisions 

require those using or making available substances with properties of very high concern to apply for an 

authorisation of each use within deadlines set by the European Commission. The aim of the authorization 

is to ensure the good functioning of the internal market while assuring that the risks from substances of 

very high concern are properly controlled and that these substances are eventually replaced by suitable 

alternative substances or technologies where these are economically and technically viable. The proposed 

new EU legislation is expected to come into force in 2007. 

255. Chemical substances in the United States are regulated under several statutes. Predominant 

among these for use in requesting data and collecting risk information include the Federal Insecticide, 

Fungicide and Rodenticide Act (FIFRA) for pesticide active ingredients, the Toxic Substances Control Act 

(TSCA) for general chemicals, the Federal Food, Drug ad Cosmetic Act (FFDCA) for food additives, 

pharmaceuticals and personal care products, and the Clean Water Act (CWA) for surface water and 

drinking water contaminants. Any of these authorities may be employed singly or in combination in 

obtaining hazard data relevant to potential endocrine disrupting substances. Existing test guidelines 

employed for aquatic invertebrates include various acute and chronic methods across several taxa, 

including daphnids, copepods, mysids, penaeids, chironomids, and pelecypods. The primary chronic 

methods for aquatic arthropods, which are routinely required for outdoor use pesticides, are the OPPTS 

850.1300 Daphnia chronic toxicity test (which is consistent with the OECD TG 211) and the OPPTS 

850.1350 Mysid chronic toxicity test (no OECD counterpart). Neither of these test methods is considered 

sufficient for fully addressing potential endocrine disrupting substances. As part of the Endocrine 

Disruptor Screening Program (EDSP), a two-generation mysid test procedure (described in Section 5.1.1.1 

and Annex 1) is being developed to provide a fully comprehensive second tier evaluation of substances 

identified as possible endocrine disruptors in first tier studies. Although the focus of the EDSP tier one is 

on vertebrate relevant endocrine processes, the mysid test in tier two is deemed necessary to assess the 

potential adverse consequence of tier one identified putative disrupting agents to invertebrate endocrine 

regulated processes. 

256. In Japan, chemical substances (including potential endocrine disruptors) are under the control of 

the amended Chemical Substances Control Law. This law concerns production, import and export of 

chemical substances. In spring 2002, the Ministry of the Environment also formed a task force on 

endocrine disrupters testing and assessment for invertebrates (crustaceans). 
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9.0 REFERENCES

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Nervous System. General and Comparative Endocrinology 119(3):317-328. 

2. Andersen, H.R., Wollenberger, L., Halling-Sorensen, B., and Kusk, K.O. 2001. Development of 

Copepod Nauplii to Copepodites: a Parameter for Chronic Toxicity Including Endocrine 

Disruption. Environmental Toxicology and Chemistry 20(12):2821-2829. 

3. Andries, J.C. 2001. Endocrine and Environmental Control of Reproduction in Polychaeta. 

Canadian Journal of Zoology Revue Canadienne de Zoologie 79(2):254-270. 

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ANNEXES 

ANNEX 1 – Mysid life cycle draft protocol (available from the OECD upon request, please contact 

env.tgcontact@oecd.org) 

ANNEX 2 – Daphnia life cycle draft protocol (available from the OECD upon request, please contact 

env.tgcontact@oecd.org)

ANNEX 3 – Copepod life cycle draft protocol (available from the OECD upon request, please contact 

env.tgcontact@oecd.org)

ANNEX 4 – Chironomid life cycle draft protocol (available from the OECD upon request, please 

contact env.tgcontact@oecd.org)