STI/PUB/1239

INTERNATIONAL ATOMIC ENERGY AGENCY

VIENNA

ISBN

92–0–114705–8

ISSN 1020–6566

The explosion on 26 April 1986 at the Chernobyl nuclear

power plant and the consequent reactor fire resulted in

an unprecedented release of radioactive material from

a nuclear reactor and adverse consequences for the

public and the environment. Although the accident

occurred nearly two decades ago, controversy still

surrounds the real impact of the disaster. Therefore

the IAEA, in cooperation with the Food and Agriculture

Organization of the United Nations, the United

Nations Development Programme, the United Nations

Environment Programme, the United Nations Office for

the Coordination of Humanitarian Affairs, the United

Nations Scientific Committee on the Effects of Atomic

Radiation, the World Health Organization and the World

Bank, as well as the competent authorities of Belarus,

the Russian Federation and Ukraine, established the

Chernobyl Forum in 2003. The mission of the Forum

was to generate “authoritative consensual statements”

on the environmental consequences and health effects

attributable to radiation exposure arising from the

accident as well as to provide advice on environmental

remediation and special health care programmes, and

to suggest areas in which further research is required.

This report presents the findings and recommendations

of the Chernobyl Forum concerning the environmental

effects of the Chernobyl accident.

Report of the

Chernobyl Forum Expert Group ‘Environment’

Environmental Consequences

of the Chernobyl Accident

nd their Remediation:

wenty Years of Experience

RADIOLOGICAL

AS SES SM E NT

R E P O R T S

S E R I E S

Environmental Consequences of the Chernobyl Accident and their Remediation: Twenty Years of Experience

9.9 mm

180 pages

P1239_covI+IV.indd 1 2006-03-30 14:41:37

ENVIRONMENTAL CONSEQUENCES

OF THE CHERNOBYL ACCIDENT

AND THEIR REMEDIATION:

TWENTY YEARS OF EXPERIENCE

Report of the Chernobyl Forum Expert Group ‘Environment’

The following States are Members of the International Atomic Energy Agency:

The Agency’s Statute was approved on 23 October 1956 by the Conference on the Statute of the IAEA

held at United Nations Headquarters, New York; it entered into force on 29 July 1957. The Headquarters of the

Agency are situated in Vienna. Its principal objective is “to accelerate and enlarge the contribution of atomic

energy to peace, health and prosperity throughout the world’’.

AFGHANISTAN

ALBANIA

ALGERIA

ANGOLA

ARGENTINA

ARMENIA

AUSTRALIA

AUSTRIA

AZERBAIJAN

BANGLADESH

BELARUS

BELGIUM

BENIN

BOLIVIA

BOSNIA AND HERZEGOVINA

BOTSWANA

BRAZIL

BULGARIA

BURKINA FASO

CAMEROON

CANADA

CENTRAL AFRICAN

REPUBLIC

CHAD

CHILE

CHINA

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CÔTE D’IVOIRE

CROATIA

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CZECH REPUBLIC

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OF THE CONGO

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ECUADOR

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ERITREA

ESTONIA

ETHIOPIA

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HAITI

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ICELAND

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INDONESIA

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THAILAND

THE FORMER YUGOSLAV

REPUBLIC OF MACEDONIA

TUNISIA

TURKEY

UGANDA

UKRAINE

UNITED ARAB EMIRATES

UNITED KINGDOM OF

GREAT BRITAIN AND

NORTHERN IRELAND

UNITED REPUBLIC

OF TANZANIA

UNITED STATES OF AMERICA

URUGUAY

UZBEKISTAN

VENEZUELA

VIETNAM

YEMEN

ZAMBIA

ZIMBABWE

ENVIRONMENTAL CONSEQUENCES

OF THE CHERNOBYL ACCIDENT

AND THEIR REMEDIATION:

TWENTY YEARS OF EXPERIENCE

Report of the Chernobyl Forum Expert Group ‘Environment’

RADIOLOGICAL ASSESSMENT REPORTS SERIES

INTERNATIONAL ATOMIC ENERGY AGENCY

VIENNA, 2006

IAEA Library Cataloguing in Publication Data

Environmental consequences of the Chernobyl accident and their

remediation : twenty years of experience / report of the Chernobyl

Forum Expert Group ‘Environment’. — Vienna : International

Atomic Energy Agency, 2006.

p. ; 29 cm. — (Radiological assessment reports series, ISSN

1020

6566)

STI/PUB/1239

ISBN 92–0–114705–8

Includes bibliographical references.

1. Chernobyl Nuclear Accident, Chornobyl, Ukraine, 1986 —

Environmental aspects. 2. Radioactive waste sites — Cleanup.

I. International Atomic Energy Agency. II. Series.

IAEAL 06–00424

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STI/PUB/1239

FOREWORD

The explosion on 26 April 1986 at the Chernobyl nuclear power plant, which is located 100 km from Kiev

in Ukraine (at that time part of the USSR), and the consequent reactor fire, which lasted for 10 days, resulted in

an unprecedented release of radioactive material from a nuclear reactor and adverse consequences for the public

and the environment.

The resulting contamination of the environment with radioactive material caused the evacuation of more

than 100 000 people from the affected region during 1986 and the relocation, after 1986, of another 200 000

people from Belarus, the Russian Federation and Ukraine. Some five million people continue to live in areas

contaminated by the accident. The national governments of the three affected countries, supported by

international organizations, have undertaken costly efforts to remediate the areas affected by the contamination,

provide medical services and restore the region’s social and economic well-being.

The accident’s consequences were not limited to the territories of Belarus, the Russian Federation and

Ukraine, since other European countries were also affected as a result of the atmospheric transfer of radioactive

material. These countries also encountered problems in the radiation protection of their populations, but to a

lesser extent than the three most affected countries.

Although the accident occurred nearly two decades ago, controversy still surrounds the real impact of the

disaster. Therefore the IAEA, in cooperation with the Food and Agriculture Organization of the United Nations

(FAO), the United Nations Development Programme (UNDP), the United Nations Environment Programme

(UNEP), the United Nations Office for the Coordination of Humanitarian Affairs (OCHA), the United Nations

Scientific Committee on the Effects of Atomic Radiation (UNSCEAR), the World Health Organization (WHO)

and the World Bank, as well as the competent authorities of Belarus, the Russian Federation and Ukraine,

established the Chernobyl Forum in 2003. The mission of the Forum was — through a series of managerial and

expert meetings — to generate “authoritative consensual statements” on the environmental consequences and

health effects attributable to radiation exposure arising from the accident, as well as to provide advice on

environmental remediation and special health care programmes, and to suggest areas in which further research

is required. The Forum was created as a contribution to the United Nations’ ten year strategy for Chernobyl,

launched in 2002 with the publication of Human Consequences of the Chernobyl Nuclear Accident — A Strategy

for Recovery.

Over a two year period, two groups of experts from 12 countries, including Belarus, the Russian Federation

and Ukraine, and from relevant international organizations, assessed the accident’s environmental and health

consequences. In early 2005 the Expert Group ‘Environment’, coordinated by the IAEA, and the Expert Group

‘Health’, coordinated by the WHO, presented their reports for the consideration of the Chernobyl Forum. Both

reports were considered and approved by the Forum at its meeting on 18–20 April 2005. This meeting also

decided, inter alia, “to consider the approved reports… as a common position of the Forum members, i.e., of the

eight United Nations organizations and the three most affected countries, regarding the environmental and

health consequences of the Chernobyl accident, as well as recommended future actions, i.e., as a consensus within

the United Nations system.”

This report presents the findings and recommendations of the Chernobyl Forum concerning the

environmental effects of the Chernobyl accident. The Forum’s report considering the health effects of the

Chernobyl accident is being published by the WHO. The Expert Group ‘Environment’ was chaired by L. Anspaugh

of the United States of America. The IAEA technical officer responsible for this report was M. Balonov of the

IAEA Division of Radiation, Transport and Waste Safety.

EDITORIAL NOTE

Although great care has been taken to maintain the accuracy of information contained in this publication, neither the

IAEA nor its Member States assume any responsibility for consequences which may arise from its use.

The use of particular designations of countries or territories does not imply any judgement by the publisher, the IAEA, as

to the legal status of such countries or territories, of their authorities and institutions or of the delimitation of their boundaries.

The mention of names of specific companies or products (whether or not indicated as registered) does not imply any

intention to infringe proprietary rights, nor should it be construed as an endorsement or recommendation on the part of the

IAEA.

CONTENTS

1. SUMMARY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

1.1. Introduction. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

1.2. Radioactive contamination of the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

1.2.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

1.2.1.1. Radionuclide release and deposition. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

1.2.1.2. Urban environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

1.2.1.3. Agricultural environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

1.2.1.4. Forest environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

1.2.1.5. Aquatic environment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

1.2.2. Recommendations for future research and monitoring . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

1.2.2.1. General . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

1.2.2.2. Practical. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

1.2.2.3. Scientific . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

1.2.2.4. Specific recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

1.3. Environmental countermeasures and remediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

1.3.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

1.3.1.1. Radiological criteria. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

1.3.1.2. Urban countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

1.3.1.3. Agricultural countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

1.3.1.4. Forest countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

1.3.1.5. Aquatic countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

1.3.2. Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

1.3.2.1. Countries affected by the Chernobyl accident . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

1.3.2.2. Worldwide. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

1.3.2.3. Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

1.4. Human exposure. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

1.4.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

1.4.2. Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

1.5. Radiation induced effects on plants and animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

1.5.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

1.5.2. Recommendations for future research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

1.5.3. Recommendations for countermeasures and remediation . . . . . . . . . . . . . . . . . . . . . . . . . . 13

1.6. Environmental and radioactive waste management aspects of the dismantling of the

Chernobyl shelter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

1.6.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

1.6.2. Recommendations for future actions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Reference to Section 1. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

2. INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

2.1. Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

2.2. Objectives of the Chernobyl Forum. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

2.3. Method of operation and output of the Chernobyl Forum . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

2.4. Structure of the report . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

References to Section 2. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

3. RADIOACTIVE CONTAMINATION OF THE ENVIRONMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

3.1. Radionuclide release and deposition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

3.1.1. Radionuclide source term . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

3.1.2. Physical and chemical forms of released material . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

3.1.3. Meteorological conditions during the course of the accident. . . . . . . . . . . . . . . . . . . . . . . . 21

3.1.4. Concentration of radionuclides in air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22

3.1.5. Deposition of radionuclides on soil surfaces . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

3.1.6. Isotopic composition of the deposition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

3.2. Urban environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27

3.2.1. Deposition patterns . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27

3.2.2. Migration of radionuclides in the urban environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28

3.2.3. Dynamics of the exposure rate in urban environments. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

3.3. Agricultural environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

3.3.1. Radionuclide transfer in the terrestrial environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

3.3.2. Food production systems affected by the accident . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

3.3.3. Effects on agriculture in the early phase. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

3.3.4. Effects on agriculture in the long term phase. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

3.3.4.1. Physicochemistry of radionuclides in the soil–plant system . . . . . . . . . . . . . . . . . 32

3.3.4.2. Migration of radionuclides in soil. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33

3.3.4.3. Radionuclide transfer from soil to crops. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34

3.3.4.4. Dynamics of radionuclide transfer to crops . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

3.3.4.5. Radionuclide transfer to animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

3.3.5. Current contamination of foodstuffs and expected future trends . . . . . . . . . . . . . . . . . . . . 40

3.4. Forest environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

3.4.1. Radionuclides in European forests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

3.4.2. Dynamics of contamination during the early phase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

3.4.3. Long term dynamics of radiocaesium in forests. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

3.4.4. Uptake into edible products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

3.4.5. Contamination of wood. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

3.4.6. Expected future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

3.4.7. Radiation exposure pathways associated with forests and forest products . . . . . . . . . . . . 46

3.5. Radionuclides in aquatic systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

3.5.1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

3.5.2. Radionuclides in surface waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

3.5.2.1. Distribution of radionuclides between dissolved and

particulate phases. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

3.5.2.2. Radioactivity in rivers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

3.5.2.3. Radioactivity in lakes and reservoirs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

3.5.2.4. Radionuclides in freshwater sediments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

3.5.3. Uptake of radionuclides to freshwater fish. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

3.5.3.1. Iodine-131 in freshwater fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

3.5.3.2. Caesium-137 in freshwater fish and other aquatic biota . . . . . . . . . . . . . . . . . . . . 53

3.5.3.3. Strontium-90 in freshwater fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

3.5.4. Radioactivity in marine ecosystems. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

3.5.4.1. Distribution of radionuclides in the sea . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

3.5.4.2. Transfers of radionuclides to marine biota. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

3.5.5. Radionuclides in groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

3.5.5.1. Radionuclides in groundwater: Chernobyl exclusion zone. . . . . . . . . . . . . . . . . . 56

3.5.5.2. Radionuclides in groundwater: outside the Chernobyl exclusion zone. . . . . . . . 58

3.5.5.3. Irrigation water. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 58

3.5.6. Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 58

3.5.6.1. Freshwater ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 58

3.5.6.2. Marine ecosystems. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

3.6. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . . . . . . 60

3.7. Further monitoring and research needed . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

References to Section 3. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

4. ENVIRONMENTAL COUNTERMEASURES AND REMEDIATION. . . . . . . . . . . . . . . . . . . . . . . 69

4.1. Radiological criteria. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69

4.1.1. International radiological criteria and standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69

4.1.2. National radiological criteria and standards. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71

4.2. Urban decontamination. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72

4.2.1. Decontamination research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73

4.2.2. Chernobyl experience . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73

4.2.3. Recommended decontamination technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

4.3. Agricultural countermeasures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75

4.3.1. Early phase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75

4.3.2. Late phase. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77

4.3.3. Countermeasures in intensive agricultural production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78

4.3.3.1. Soil treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79

4.3.3.2. Change in fodder crops grown on contaminated land. . . . . . . . . . . . . . . . . . . . . . 80

4.3.3.3. Clean feeding . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 80

4.3.3.4. Administration of caesium binders . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 81

4.3.4. Summary of countermeasure effectiveness in intensive production . . . . . . . . . . . . . . . . . . 81

4.3.5. Countermeasures in extensive production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 81

4.3.6. Current status of agricultural countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 83

4.3.7. A wider perspective on remediation, including socioeconomic issues . . . . . . . . . . . . . . . . 83

4.3.8. Current status and future of abandoned land. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84

4.3.8.1. Exclusion and resettlement zones in Belarus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84

4.3.8.2. Rehabilitation of contaminated lands in Ukraine . . . . . . . . . . . . . . . . . . . . . . . . . 85

4.3.8.3. Abandoned zones in the Russian Federation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 86

4.4. Forest countermeasures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 86

4.4.1. Studies on forest countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87

4.4.2. Countermeasures for forests contaminated with radiocaesium . . . . . . . . . . . . . . . . . . . . . . 87

4.4.2.1. Management based countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87

4.4.2.2. Technology based countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87

4.4.3. Examples of forest countermeasures. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89

4.5. Aquatic countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90

4.5.1. Measures to reduce doses at the water supply and treatment stage . . . . . . . . . . . . . . . . . . 90

4.5.2. Measures to reduce direct and secondary contamination of surface waters. . . . . . . . . . . . 91

4.5.3. Measures to reduce uptake by fish and aquatic foodstuffs . . . . . . . . . . . . . . . . . . . . . . . . . . 92

4.5.4. Countermeasures for groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

4.5.5. Countermeasures for irrigation water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

4.6. Conclusions and recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

4.6.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

4.6.2. Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 94

4.6.2.1. Countries affected by the Chernobyl accident . . . . . . . . . . . . . . . . . . . . . . . . . . . . 94

4.6.2.2. Worldwide. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95

4.6.2.3. Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95

References to Section 4. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96

5. HUMAN EXPOSURE LEVELS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

5.1. Introduction. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

5.1.1. Populations and areas of concern. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

5.1.2. Exposure pathways . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

5.1.3. Concepts of dose . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. . . . . . . . 101

5.1.4.

Background radiation levels . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 101

5.1.5. Decrease of dose rate with time. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102

5.1.6. Critical groups . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102

5.2. External exposure. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103

5.2.1. Formulation of the model of external exposure. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103

5.2.2. Input data for the estimation of effective external dose . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103

5.2.2.1. Dynamics of external gamma dose rate over open undisturbed soil . . . . . . . . . . 103

5.2.2.2. Dynamics of external gamma dose rate in anthropogenic areas . . . . . . . . . . . . . 105

5.2.2.3. Behaviour of people in the radiation field . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 105

5.2.2.4. Effective dose per unit gamma dose in air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106

5.2.3. Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106

5.2.3.1. Dynamics of external effective dose . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106

5.2.3.2. Measurement of individual external dose with

thermoluminescent dosimeters. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106

5.2.3.3. Levels of external exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 108

5.3. Internal dose . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109

5.3.1. Model for internal dose. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109

5.3.2. Monitoring data as input for the assessment of internal dose . . . . . . . . . . . . . . . . . . . . . . . 109

5.3.3. Avoidance of dose by human behaviour. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110

5.3.4. Results for doses to individuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110

5.3.4.1. Thyroid doses due to radioiodines. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110

5.3.4.2. Long term internal doses from terrestrial pathways . . . . . . . . . . . . . . . . . . . . . . . 112

5.3.4.3. Long term doses from aquatic pathways. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 115

5.4. Total (external and internal) exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 116

5.5. Collective doses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 118

5.5.1. Thyroid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 118

5.5.2. Total (external and internal) dose from terrestrial pathways. . . . . . . . . . . . . . . . . . . . . . . . 118

5.5.3. Internal dose from aquatic pathways. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119

5.6. Conclusions and recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119

5.6.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119

5.6.2. Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

References to Section 5. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

6. RADIATION INDUCED EFFECTS ON PLANTS AND ANIMALS . . . . . . . . . . . . . . . . . . . . . . . . . 125

6.1. Prior knowledge of radiation effects on biota. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 125

6.2. Temporal dynamics of radiation exposure following the Chernobyl accident . . . . . . . . . . . . . . . . 127

6.3. Radiation effects on plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128

6.4. Radiation effects on soil invertebrates. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 130

6.5. Radiation effects on farm animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 131

6.6. Radiation effects on other terrestrial animals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132

6.7. Radiation effects on aquatic organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132

6.8. Genetic effects in animals and plants. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 133

6.9. Secondary impacts and current conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 135

6.10. Conclusions and recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137

6.10.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137

6.10.2. Recommendations for future research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137

6.10.3. Recommendations for countermeasures and remediation . . . . . . . . . . . . . . . . . . . . . . . . . . 138

References to Section 6. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138

7. ENVIRONMENTAL AND RADIOACTIVE WASTE MANAGEMENT ASPECTS OF THE

DISMANTLING OF THE CHERNOBYL SHELTER . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 141

7.1. Current status and the future of unit 4 and the shelter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

. 141

7.1.1.

Unit 4 of the Chernobyl nuclear power plant after the accident . . . . . . . . . . . . . . . . . . . . . 141

7.1.2. Current status of the damaged unit 4 and the shelter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 142

7.1.3. Long term strategy for the shelter and the new safe confinement. . . . . . . . . . . . . . . . . . . . 144

7.1.4. Environmental aspects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145

7.1.4.1. Current status of the shelter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145

7.1.4.2. Impact on air. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145

7.1.4.3. Impact on surface water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145

7.1.4.4. Impact on groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 148

7.1.4.5. Impacts of shelter collapse without the new safe confinement . . . . . . . . . . . . . . 148

7.1.4.6. Impacts of shelter collapse within the new safe confinement. . . . . . . . . . . . . . . . 150

7.1.5. Issues and areas for improvement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151

7.1.5.1. Influence of the source term uncertainty on environmental decisions . . . . . . . . 151

7.1.5.2. Characterization of fuel-containing material . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151

7.1.5.3. Removal of fuel-containing material concurrent with development

of a geological disposal facility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151

7.2. Management of radioactive waste from the accident. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151

7.2.1. Current status of radioactive waste from the accident . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 153

7.2.1.1. Radioactive waste associated with the shelter . . . . . . . . . . . . . . . . . . . . . . . . . . . . 153

7.2.1.2. Mixing of accident related waste with operational radioactive waste . . . . . . . . . 154

7.2.1.3. Temporary radioactive waste storage facilities . . . . . . . . . . . . . . . . . . . . . . . . . . . 154

7.2.1.4. Radioactive waste disposal facilities . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 155

7.2.2. Radioactive waste management strategy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 156

7.2.3. Environmental aspects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 157

7.2.4. Issues and areas of improvement. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159

7.2.4.1. Radioactive waste management programme for the exclusion zone

and the Chernobyl nuclear power plant . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159

7.2.4.2. Decommissioning of unit 4 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159

7.2.4.3. Waste acceptance criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159

7.2.4.4. Long term safety assessment of existing radioactive waste storage sites . . . . . . 160

7.2.4.5. Potential recovery of temporary waste storage facilities located in the

Chernobyl exclusion zone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 160

7.3. Future of the Chernobyl exclusion zone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 160

7.4. Conclusions and recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 161

7.4.1. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 161

7.4.2. Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 162

References to Section 7. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 163

CONTRIBUTORS TO DRAFTING AND REVIEW . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 165

1. SUMMARY

1.1. INTRODUCTION

This report provides an up to date evaluation

of the environmental effects of the accident that

occurred on 26 April 1986 at the Chernobyl nuclear

power plant. Even though it is now nearly 20 years

after the accident, there are still many conflicting

reports and rumours concerning its consequences.

For this reason the Chernobyl Forum was initiated

by the IAEA in cooperation with the Food and

Agriculture Organization of the United Nations

(FAO), the United Nations Development

Programme (UNDP), the United Nations

Environment Programme (UNEP), the United

Nations Office for the Coordination of Humani-

tarian Affairs (OCHA), the United Nations

Scientific Committee on the Effects of Atomic

Radiation (UNSCEAR), the World Health Organi-

zation (WHO) and the World Bank, as well as the

competent authorities of Belarus, the Russian

Federation and Ukraine. The first organizational

meeting of the Chernobyl Forum was held on

5 February 2003, at which time the decision was

taken to establish the Forum as an ongoing entity of

the above named organizations.

The Chernobyl Forum was established as a

series of managerial, expert and public meetings

with the purpose of generating authoritative

consensual statements on the health effects attrib-

utable to radiation exposure arising from the

accident and the environmental consequences

induced by the released radioactive material,

providing advice on remediation and special health

care programmes and suggesting areas in which

further research is required. The terms of

reference of the Forum as approved at the meeting

were:

(a) To explore and refine the current scientific

assessments on the long term health and

environmental consequences of the

Chernobyl accident, with a view to producing

authoritative consensus statements focusing

on:

(i) The health effects attributable to

radiation exposure caused by the

accident;

(ii) The environmental consequences

induced by the radioactive material

released due to the accident (e.g. contam-

ination of foodstuffs);

(iii) The consequences attributable to the

accident but not directly related to the

radiation exposure or radioactive

contamination.

(b) To identify gaps in scientific research relevant

to the radiation induced or radioactive

contamination induced health and environ-

mental impacts of the accident, and to suggest

areas in which further work is required based

on an assessment of the work done in the past

and bearing in mind ongoing work and

projects.

mentation of, scientifically sound programmes

on mitigation of the accident consequences,

including possible joint actions of the organi-

zations participating in the Forum, such as:

(i) Remediation of contaminated land, with

the aim of making it suitable for normal

agricultural, economic and social life

under safe conditions;

(ii) Special health care of the affected

population;

(iii) Monitoring of long term human exposure

to radiation;

(iv) Addressing the environmental issues

pertaining to the decommissioning of the

Chernobyl shelter and the management

of radioactive waste originating from the

Chernobyl accident.

The Chernobyl Forum is a high level organi-

zation of senior officials of United Nations agencies

and the three most affected countries. The technical

reports of the Forum were produced by two expert

groups: Expert Group ‘Environment’ (EGE) and

Expert Group ‘Health’ (EGH). The membership of

the two groups comprised recognized international

scientists and experts from the three most affected

countries. Through the work of these two groups

and their subworking groups, the technical

documents were prepared. The EGE was

coordinated by the IAEA and the EGH was

coordinated by the WHO.

In all cases, the scientists of the EGE and

EGH were able to reach consensus on the contents

of their respective technical documents. The

technical reports were finally approved by the

Chernobyl Forum itself. This report, on the environ-

mental consequences, is published by the IAEA; the

report on the health consequences will be published

by the WHO.

1.2. RADIOACTIVE CONTAMINATION OF

THE ENVIRONMENT

The Chernobyl accident caused a large

regional release of radionuclides into the

atmosphere and subsequent radioactive contami-

nation of the environment. Many European

countries were affected by the radioactive contami-

nation; among the most affected were three former

republics of the Soviet Union, now Belarus, the

Russian Federation and Ukraine. The deposited

radionuclides gradually decayed and moved within

and among the environments — atmospheric,

aquatic, terrestrial and urban.

1.2.1. Conclusions

1.2.1.1. Radionuclide release and deposition

Major releases from unit 4 of the Chernobyl

nuclear power plant continued for ten days, and

included radioactive gases, condensed aerosols and

a large amount of fuel particles. The total release of

radioactive substances was about 14 EBq

(as of

26 April 1986), which included 1.8 EBq of

131

0.085 EBq of

137

Cs and other caesium radioisotopes,

0.01 EBq of

Sr and 0.003 EBq of plutonium

radioisotopes. The noble gases contributed about

50% of the total release of radioactivity.

Large areas of Europe were affected to some

degree by the Chernobyl releases. An area of more

than 200 000 km

in Europe was contaminated with

radiocaesium (above 0.04 MBq of

137

Cs/m

), of

which 71% was in the three most affected countries

(Belarus, the Russian Federation and Ukraine). The

deposition was highly heterogeneous; it was

strongly influenced by rain when the contaminated

air masses passed. In the mapping of the deposition,

137

Cs was chosen because it is easy to measure and is

of radiological significance. Most of the strontium

and plutonium radioisotopes were deposited close

(less than 100 km) to the reactor, due to their being

contained within larger particles.

Much of the release comprised radionuclides

with short physical half-lives; long lived radio-

nuclides were released in smaller amounts. Thus

many of the radionuclides released by the accident

have already decayed. The releases of radioactive

iodines caused concern immediately after the

accident. Owing to the emergency situation and the

short half-life of

131

I, few reliable measurements

were made of the spatial distribution of deposited

radioiodine (which is important in determining

doses to the thyroid). Current measurements of

129

may assist in estimating

131

I deposition better and

thereby improve thyroid dose reconstruction.

After the initial period,

137

Cs became the

nuclide of greatest radiological importance, with

Sr being of less importance. For the first years

134

Cs was also important. Over the longer term

(hundreds to thousands of years), the only radio-

nuclides anticipated to be of interest are the

plutonium isotopes and

241

Am.

1.2.1.2. Urban environment

In urban areas, open surfaces such as lawns,

parks, streets, roads, squares, roofs and walls

became contaminated with radionuclides. Under

dry conditions, trees, bushes, lawns and roofs

became more contaminated; under wet conditions,

horizontal surfaces such as soil plots, lawns, etc.,

received the highest contamination. Particularly

high

137

Cs activity concentrations were found

around houses where rain had transported the

radioactive material from the roofs to the ground.

The deposition in urban areas in the nearest city of

Pripyat and surrounding settlements could have

initially given rise to substantial external radiation

doses, but this was partially averted by the

evacuation of the people. The deposited radioactive

material in other urban areas has given rise to

exposure of the public in the subsequent years and

continues to do so.

Due to wind and rain and human activities,

including traffic, street washing and cleanup, surface

contamination by radioactive material was reduced

significantly in inhabited and recreational areas

during 1986 and afterwards. One of the conse-

quences of these processes has been the secondary

contamination of sewage systems and sludge

storage areas.

At present, in most of the settlements

subjected to radioactive contamination, the air dose

rate above solid surfaces has returned to the pre-

accident background level. The elevated air dose

1 EBq = 10

Bq (becquerel).

rate remains mainly over undisturbed soil in

gardens, kitchen gardens and parks.

1.2.1.3. Agricultural environment

In the early phase, direct surface deposition of

many different radionuclides dominated the

contamination of agricultural plants and the animals

consuming them. The release and deposition of

radioiodine isotopes caused the most immediate

concern, but the problem was confined to the first

two months, because of the short physical half-life

(eight days) of the most important iodine isotope,

131

I. The radioiodine was rapidly transferred to milk

at a high rate in Belarus, the Russian Federation

and Ukraine, leading to significant thyroid doses to

those consuming milk, especially children. In the

rest of Europe the consequences of the accident

varied; increased levels of radioiodine in milk were

observed in some contaminated southern areas

where dairy animals were already outdoors.

Different crop types, in particular green leafy

vegetables, were also contaminated with radio-

nuclides to varying degrees, depending on the

deposition levels and the stage of the growing

season. Direct deposition on to plant surfaces was of

concern for about two months.

After the early phase of direct contamination,

uptake of radionuclides through plant roots from

soil became increasingly important and showed

strong time dependence. Radioisotopes of caesium

(

137

Cs and

134

Cs) were the nuclides that led to the

greatest problems, and after the decay of

134

Cs,

137

remains to cause problems in some Belarusian,

Russian and Ukrainian areas. In addition,

causes problems in the near field of the reactor, but

at longer distances the deposition levels were too

low to be of radiological significance. Other radio-

nuclides, such as plutonium isotopes and

241

Am,

either were present at very low deposition levels or

were not very available for root uptake, and

therefore did not cause real problems in

agriculture.

In general, there was an initial substantial

reduction in the transfer of radionuclides to

vegetation and animals, as would be expected, due

to weathering, physical decay, migration of radio-

nuclides down the soil column and reduction in

radionuclide bioavailability in soil. Particularly in

contaminated intensive agricultural systems, mostly

in the former USSR, there was substantial

reduction in the transfer of

137

Cs to plants and

animals, especially in the first few years. However,

in the past decade there has been little further

obvious decline, and long term effective half-lives

have been difficult to quantify with precision.

The radiocaesium activity concentrations in

foodstuffs after the early phase were influenced not

only by deposition levels but also by soil types,

management practices and types of ecosystem. The

major and persistent problems in the affected areas

occur in extensive agricultural systems with soils

with a high organic content and where animals graze

on unimproved pastures that are not ploughed or

fertilized. In particular, this affects rural residents in

the former USSR, who are commonly subsistence

farmers with privately owned dairy cattle.

In the long term,

137

Cs in meat and milk, and

to a lesser extent

137

Cs in vegetables, remains the

most important contributor to human internal dose.

As its activity concentration, in both vegetable and

animal foods, has been decreasing during the past

decade very slowly, at 3–7%/a, the contribution of

137

Cs to dose will continue to dominate for decades

to come. The contribution of other long lived radio-

nuclides,

Sr, plutonium isotopes and

241

Am, to

human dose will remain insignificant.

1.2.1.4. Forest environment

Following the Chernobyl accident, vegetation

and animals in forests and mountain areas showed a

particularly high uptake of radiocaesium, with the

highest recorded

137

Cs activity concentrations being

found in forest products, due to the persistent

recycling of radiocaesium in forest ecosystems.

Particularly high

137

Cs activity concentrations have

been found in mushrooms, berries and game, and

these high levels have persisted since the accident.

Thus, while there has been a general decline in the

magnitude of exposures due to the consumption of

agricultural products, there have been continued

high levels of contamination in forest food products,

which still exceed intervention limits in many

countries. This can be expected to continue for

several decades to come. Therefore, the relative

importance of forests in contributing to the

radiation exposures of the populations of several

affected countries has increased with time. It will be,

primarily, the combination of downward migration

in the soil and the physical decay of

137

Cs that

contribute to any further reduction in the contami-

nation of forest food products.

The high transfer of radiocaesium in the

lichen–reindeer meat–humans pathway was demon-

strated after the Chernobyl accident in the Arctic

and sub-Arctic areas of Europe. The Chernobyl

accident led to considerable contamination of

reindeer meat in Finland, Norway, the Russian

Federation and Sweden, and caused significant

problems for the Sami people.

The use of timber and associated products

makes only a small contribution to the exposure of

the general public, although wood ash can contain

high amounts of

137

Cs and could potentially give rise

to higher doses than other uses of wood. Caesium-

137 in timber is of minor importance, although

doses in the wood pulp industry have to be

considered.

Forest fires increased air activity concentra-

tions in 1992, but not to a high extent. The possible

radiological consequences of forest fires have been

much discussed, but these are not expected to cause

any problems of radionuclide transfer from contam-

inated forests, except, possibly, in the nearest

surroundings of the fire.

1.2.1.5. Aquatic environment

Radionuclides from Chernobyl contaminated

surface water systems not only in areas close to the

site but also in many other parts of Europe. The

initial contamination of water was due primarily to

direct deposition of radionuclides on to the surfaces

of rivers and lakes and was dominated by short lived

radionuclides (most importantly

131

I). In the first

few weeks after the accident, activity concentrations

in drinking water from the Kiev reservoir were a

particular concern.

The contamination of water bodies decreased

rapidly during the weeks after fallout through

dilution, physical decay and absorption of radio-

nuclides by catchment soils. For lakes and

reservoirs, the settling of suspended particles to the

bed sediments also played an important role in

reducing radionuclide levels in water. Bed

sediments are an important long term sink for

radionuclides.

The initial uptake of radioiodine by fish was

rapid, but activity concentrations declined quickly,

due primarily to physical decay. Bioaccumulation of

radiocaesium in the aquatic food chain led to

significant concentrations in fish in the most

affected areas, and in some lakes as far away as

Scandinavia and Germany. Owing to generally

lower fallout and lower bioaccumulation,

activity concentrations in fish were not a significant

contributor to human dose in comparison with

radiocaesium, particularly since

Sr is accumulated

in bone rather than in edible muscle.

In the long term, secondary contamination by

wash-off of long lived

137

Cs and

Sr from contami-

nated soils and remobilization from bed sediments

continues (at a much lower level) to the present day.

Catchments with a high organic content (peat soils)

release much more radiocaesium to surface waters

than those with mostly mineral soils. At present,

surface water activity concentrations are low;

irrigation with surface water is therefore not

considered to be a problem.

Fuel particles deposited in the sediments of

rivers and lakes close to the Chernobyl nuclear

power plant show significantly lower weathering

rates than the same particles in terrestrial soils. The

half-life of these particles is roughly the same as the

physical half-life of the radionuclides

Sr and

137

Cs.

While

137

Cs and

Sr activity concentrations in

the water and fish of rivers, open lakes and

reservoirs are currently low, the most contaminated

lakes are those few lakes with limited inflowing and

outflowing streams (‘closed’ lakes) in Belarus, the

Russian Federation and Ukraine that have a poor

mineral nutrient status. Activity concentrations of

137

Cs in fish in some of these lakes will remain for a

significant time into the future. In a population

living next to a closed lake system (e.g. Lake

Kozhanovskoe in the Russian Federation),

consumption of fish has dominated the total

137

ingestion for some people.

Owing to the large distance of the Black and

Baltic Seas from Chernobyl, and the dilution in

these systems, activity concentrations in sea water

have been much lower than in fresh water. The low

radionuclide concentrations in the water combined

with the low bioaccumulation of radiocaesium in

marine biota has led to activity concentrations in

marine fish that are not of concern.

1.2.2. Recommendations for future research and

monitoring

1.2.2.1. General

Various ecosystems considered in this report

have been intensively monitored and studied during

the years after the Chernobyl accident, and the

transfer and bioaccumulation of the most important

long term contaminants,

137

Cs and

Sr, are now

generally well understood. There is, therefore, little

urgent need for major new research programmes on

radionuclides in ecosystems; there is, however, a

requirement for continued, but more limited,

targeted monitoring of the environments, and for

further research in some specific areas, as detailed

below.

Long term monitoring of radionuclides

(especially

137

Cs and

Sr) in various environmental

compartments is required to meet the general

practical and scientific needs described below.

1.2.2.2. Practical

The practical needs are to:

(a) Assess current and predict future levels of

human exposure and contamination of foods

in order to justify remedial actions and long

term countermeasures.

(b) Inform the general public in affected areas

about the persistence of radioactive contami-

nation in food products and its seasonal and

annual variability in natural food products

gathered by themselves (such as mushrooms,

game, freshwater fish from closed lakes,

berries, etc.) and give advice on dietary and

food preparation methods to reduce radionu-

clide intake by humans.

about changing radiological conditions in

order to relieve public concerns.

1.2.2.3. Scientific

The scientific needs are to:

(a) Determine the parameters of the long term

transfer of radionuclides in various

ecosystems and different natural conditions in

order to improve predictive models both for

use in Chernobyl affected areas and for

application to potential future radioactive

releases.

(b) Determine mechanisms of radionuclide

behaviour in less studied ecosystems (e.g. the

role of fungi in forests) in order to

understand the mechanisms determining the

persistence of radionuclides in these

ecosystems and to explore possibilities for

remediation, with special attention to be paid

to processes of importance for contribution

to human and biota doses.

As activity concentrations in environmental

compartments are now in quasi-equilibrium and

changing slowly, the number and frequency of

sampling and measurements performed in

monitoring and research programmes can be

substantially reduced compared with the early years

after the Chernobyl accident.

The deposits of

137

Cs and a number of other

long lived radionuclides in the 30 km zone should be

used for radioecological studies of the various

ecosystems located in this highly contaminated

area. Such studies are, except for very small scale

experiments, not possible or difficult to perform

elsewhere.

1.2.2.4. Specific recommendations

Updated mapping of

137

Cs deposition in

Albania, Bulgaria and Georgia should be

performed in order to complete the study of the

post-Chernobyl contamination of Europe.

Improved mapping of

131

I deposition, based

both on historical environmental measurements

carried out in 1986 and on recent measurements of

129

I in soil samples in areas where elevated thyroid

cancer incidence has been detected after the

Chernobyl accident, would reduce the uncertainty

in thyroid dose reconstruction needed for the deter-

mination of radiation risks.

Long term monitoring of

137

Cs and

activity concentrations in agricultural plant and

animal products produced in areas with various soil

and climate conditions and different agricultural

practices should be performed in the next decades,

in the form of limited target research programmes

on selected sites, to determine parameters for the

modelling of long term transfer.

Studies of the distribution of

137

Cs and

plutonium radionuclides in the urban environment

(Pripyat, Chernobyl and some other contaminated

towns) at long times after the accident would

improve modelling of human external exposure and

inhalation of radionuclides in the event of a nuclear

or radiological accident or malicious action.

Continued long term monitoring of specific

forest products, such as mushrooms, berries and

game, should be carried out in those areas in which

forests were significantly contaminated and where

the public consumes wild foods. The results from

such monitoring are being used by the relevant

authorities in the affected countries to provide

advice to the general public on the continued use of

forests for recreation and the gathering of wild

foods.

In addition to the general monitoring of forest

products, required for radiation protection, more

detailed, scientifically based, long term monitoring

of specific forest sites is required to provide an

ongoing and improved understanding of the

mechanisms, long term dynamics and persistence of

radiocaesium contamination and its variability. It is

desirable to explore further the key organisms, for

example fungi, and their role in radiocaesium

mobility and long term behaviour in forest

ecosystems. Such monitoring programmes are being

carried out in the more severely affected countries,

such as Belarus and the Russian Federation, and it

is important that these continue into the foreseeable

future if the current uncertainties on long term

forecasts are to be reduced.

Aquatic systems have been intensively

monitored and studied during the years after the

Chernobyl accident, and transfers and

bioaccumulation of the most important long term

contaminants,

Sr and

137

Cs, are now well

understood. There is, however, a requirement for

continued (but perhaps more limited) monitoring of

the aquatic environment, and for further research in

some specific areas, as detailed below.

Although there is currently no need for major

new research programmes on radioactivity in

aquatic systems, predictions of future contami-

nation of aquatic systems by

Sr and

137

Cs would be

improved by continued monitoring of radioactivity

in key systems (the Pripyat–Dnieper system, the

seas, and selected rivers and lakes in the most

affected areas and western Europe). This would

continue the excellent existing time series measure-

ments of activity concentrations in water, sediments

and fish, and enable the refinement of predictive

models for these radionuclides.

Although they are currently of minor radio-

logical importance in comparison with

Sr and

137

Cs, further studies of transuranic elements in the

Chernobyl zone would help to improve predictions

of environmental contamination in the very long

term (hundreds to thousands of years). Further

empirical studies of transuranic radionuclides and

Tc are unlikely to have direct implications for

radiological protection in the Chernobyl affected

areas, but would add to knowledge of the environ-

mental behaviour of these very long lived radio-

nuclides.

Future plans to reduce the water level of the

Chernobyl cooling pond will have significant

implications for its ecology and the behaviour of

radionuclides/fuel particles in newly exposed

sediments. Specific studies on the cooling pond

should therefore continue. In particular, further

study of fuel particle dissolution rates in aquatic

systems such as the cooling pond would improve

knowledge of these processes.

1.3. ENVIRONMENTAL

COUNTERMEASURES AND

REMEDIATION

After the Chernobyl accident, the authorities

in the USSR introduced a range of short term and

long term countermeasures to reduce the effects of

the environmental contamination. The counter-

measures consumed a great amount of human,

economic and scientific resources. Unfortunately,

there was not always openness and transparency in

the actions of the authorities, and information was

withheld from the public. This can, in part, explain

some of the problems experienced later in commu-

nication with the public, and the public’s mistrust of

the authorities. Similar behaviour in many other

countries outside the Russian Federation, Belarus

and Ukraine led to a distrust in authority that, in

many countries, prompted investigations on how to

deal with such major accidents in an open and

transparent way and on how the affected people can

be involved in decision making processes.

The unique experience of countermeasure

application after the Chernobyl accident has

already been widely used both at the national and

international levels in order to improve prepar-

edness against future nuclear and radiological

emergencies.

1.3.1. Conclusions

1.3.1.1. Radiological criteria

At the time of the Chernobyl accident, well

developed international and national guidance on

general radiation protection of the public and

specific guidance applicable to major nuclear

emergencies was in place. The basic methodology of

the guidance used in the former USSR was different

from that of the international system, but the dose

limits of the radiation safety standards were similar.

The then available international and national

standards were widely applied for the protection of

the populations of the European countries affected

by the accident.

The scale and long term consequences of the

Chernobyl accident required the development of

some additional national and international

radiation safety standards as a result of changing

radiological conditions.

1.3.1.2. Urban countermeasures

Decontamination of settlements was widely

applied as a countermeasure in the contaminated

regions of the USSR during the first years after the

Chernobyl accident as a means of reducing the

external exposure of the public and the inhalation

of resuspended radioactive substances.

Decontamination was cost effective with

regard to reduction of external dose when its

planning and implementation were preceded by a

remediation assessment based on cost–benefit

techniques and external dosimetry data. Since the

areas have been cleaned up, no secondary

contamination of cleaned up plots has been

observed.

The decontamination of urban environments

has produced a considerable amount of low level

radioactive waste, which, in turn, has created a

problem of disposal.

Numerous experimental studies and

associated modelling have been used as the

scientific basis for developing improved recommen-

dations for decontamination of the urban

environment. Such recommendations could be used

in the event of any future large scale radioactive

contamination of urban areas.

1.3.1.3. Agricultural countermeasures

Countermeasures applied in the early phase of

the Chernobyl accident were only partially effective

in reducing radioiodine intake via milk, because of

the lack of timely information about the accident

and guidance on recommended actions, particularly

for private farmers. This led to significant

radioiodine exposure of some people in the affected

countries.

The most effective countermeasures in the

early phase were exclusion of contaminated pasture

grasses from animals’ diets and the rejection of

milk. Feeding animals with clean fodder was

effectively implemented in some countries;

however, this countermeasure was not widely

applied in the USSR, due to a lack of uncontami-

nated feeds. Slaughtering of cattle was often carried

out, but it was unjustified from a radiological point

of view and caused significant hygienic, practical

and economic problems.

Several months after the accident, long term

agricultural countermeasures against radiocaesium

and radiostrontium were effectively implemented in

all contaminated regions; these countermeasures

included feeding animals with clean fodder and

obligatory milk processing. This enabled most

farming practices to continue in affected areas and

resulted in a large reduction in dose. The most

important precondition was the radiation

monitoring of agricultural lands, feeds and

foodstuffs, including in vivo monitoring of caesium

activity concentrations in the muscle of cattle.

The greatest long term problem has been

radiocaesium contamination of milk and meat. In

the USSR, and later in the three independent

countries, this was addressed by the treatment of

land used for fodder crops, clean feeding and the

application of caesium binders to animals. Clean

feeding is one of the most important and effective

measures used in countries where animal products

have

137

Cs activity concentrations exceeding the

action levels. In the long term, environmental

radiation conditions are changing only slowly;

however, the efficiency of environmental counter-

measures remains at a constant level.

The application of agricultural counter-

measures in the three most affected countries has

substantially decreased since the mid-1990s,

because of economic problems. Within a short time

this resulted in an increase of radionuclide content

in plant and animal agricultural products.

There are still agricultural areas in the three

countries that remain out of use. This land could be

used after appropriate remediation, but at present

legal, economic and social constraints make this

difficult.

Where social and economic factors, along with

radiological factors, have been taken into account

during the planning and application of counter-

measures, better acceptability of the counter-

measures by the public has been achieved.

In western Europe, because of the high and

prolonged uptake of radiocaesium in the affected

extensive systems, a range of countermeasures is

still being used for animal products from uplands

and forests.

For the first time, practical, long term agricul-

tural countermeasures have been developed, tested

and implemented on a large scale; these include

radical improvement of meadows, pre-slaughter

clean feeding, the application of caesium binders,

and soil treatment and cultivation. Their implemen-

tation on more than three billion hectares of

agricultural land has made it possible to minimize

the amount of products with radionuclide activity

concentrations above the action levels in all three

countries.

1.3.1.4. Forest countermeasures

The principal forest related countermeasures

applied after the Chernobyl accident were

management based countermeasures (restrictions

of various activities normally carried out in forests)

and technology based countermeasures.

Restrictions widely applied in the three most

affected countries, and partially in Scandinavia,

included the following actions that have reduced

human exposure due to residence in radioactively

contaminated forests and the use of forest products:

(a) Restrictions on public and forest worker

access, as a countermeasure against external

exposure.

(a) Restrictions on the harvesting of food

products such as game, berries and

mushrooms. In the three most affected

countries mushrooms are widely consumed,

and therefore this restriction has been

particularly important.

(b) Restrictions on the collection of firewood by

the public, in order to prevent external

exposures in the home and garden when the

wood is burned and the ash is disposed of or

used as a fertilizer.

avoiding the consumption of meat with high

seasonal levels of radiocaesium.

(d) Fire prevention, especially in areas with large

scale radionuclide deposition, aimed at the

avoidance of secondary contamination of the

environment.

However, experience in the three most

affected countries has shown that such restrictions

can also result in significant negative social conse-

quences, and advice from the authorities to the

general public may be ignored as a result. This

situation can be offset by the provision of suitable

educational programmes targeted at the local scale

to emphasize the relevance of suggested changes in

the use of some forest areas.

It is unlikely that any technology based forest

countermeasures (i.e. the use of machinery and/or

chemical treatments to alter the distribution or

transfer of radiocaesium in the forest) will be

practicable on a large scale.

1.3.1.5. Aquatic countermeasures

Numerous countermeasures were put in place

in the months and years after the accident to protect

water systems from the transfer of radionuclides

from contaminated soils. In general, these measures

were ineffective and expensive and led to relatively

high exposures of the workers implementing the

countermeasures.

The most effective countermeasure was the

early restriction of drinking water abstraction and

the change to alternative supplies. Restrictions on

the consumption of freshwater fish have proved

effective in Scandinavia and Germany; however, in

Belarus, the Russian Federation and Ukraine such

restrictions may not always have been adhered to.

It is unlikely that any future countermeasures

to protect surface waters would be justifiable in

terms of economic cost per unit of dose reduction. It

is expected that restrictions on the consumption of

fish will remain in a few cases (in closed lakes) for

several more decades.

Future efforts in this area should be focused

on public information, because there are still public

misconceptions concerning the perceived health

risks due to radioactively contaminated waters and

fish.

1.3.2. Recommendations

1.3.2.1. Countries affected by the Chernobyl

accident

Long term remediation measures and

countermeasures should be applied in the areas

contaminated with radionuclides if they are radio-

gically justified and optimized.

Members of the general public should be

informed, along with the authorities, about the

existing radiation risk factors and the technological

possibilities to reduce them in the long term via

remediation and countermeasures, and be involved

in discussions and decision making.

In the long term, remediation measures and

countermeasures remain efficient and justified —

mainly in the agricultural areas with poor (sandy

and peaty) soils, where high radionuclide transfer

from soil to plants can occur.

Particular attention must be given to private

farms in several hundred settlements and to about

50 intensive farms in Belarus, the Russian

Federation and Ukraine, where radionuclide

concentrations in milk still exceed the national

action levels.

Among long term remediation measures,

radical improvement of pastures and grasslands, as

well as the draining of wet peaty areas, is highly

efficient. The most efficient agricultural counter-

measures are pre-slaughter clean feeding of animals

accompanied by in vivo monitoring, application of

Prussian blue to cattle and enhanced application of

mineral fertilizers in plant cultivation.

Restricting harvesting by the public of wild

food products such as game, berries, mushrooms

and fish from closed lakes may still be needed in

areas where radionuclide activity concentrations

exceed the national action levels.

Advice should continue to be given on

individual diets, as a way of reducing consumption

of highly contaminated wild food products, and on

simple cooking procedures to remove radioactive

caesium.

It is necessary to identify sustainable ways of

making use of the most affected areas, but also to

revive the economic potential of such areas for

the benefit of the community. Such strategies

should take into account the associated radiation

hazard.

1.3.2.2. Worldwide

The unique experience of countermeasure

application after the Chernobyl accident should be

carefully documented and used for the preparation

of international and national guidance for

authorities and experts responsible for radiation

protection of the public and the environment.

Practically all the long term agricultural

countermeasures implemented on a large scale in

contaminated lands of the three most affected

countries can be recommended for use in the event

of future accidents. However, the effectiveness of

soil based countermeasures varies at each site.

Analysis of soil properties and agricultural practice

before the application of countermeasures is

therefore of great importance.

Recommendations on the decontamination of

the urban environment in the event of large scale

radioactive contamination should be distributed to

the management of nuclear facilities that have the

potential for substantial accidental radioactive

release (nuclear power plants and reprocessing

plants) and to authorities in adjacent regions.

1.3.2.3. Research

Generally, the physical and chemical

processes involved in environmental counter-

measures and remediation technologies, both of a

mechanical nature (radionuclide removal, mixing

with soil, etc.) or of a chemical nature (soil liming,

fertilization, etc.), or their combinations, are

understood well enough to be modelled and applied

in similar circumstances worldwide. Much less well

understood are the biological processes that could

be used in environmental remediation (e.g.

reprofiling of agricultural production, bioremedi-

ation, etc.). These processes require more research.

An important issue that requires more socio-

logical research is the perception by the public of

the introduction, performance and withdrawal of

countermeasures in the event of an emergency, as

well as the development of social measures aimed at

involving the public in these processes at all stages,

beginning with the decision making process.

There is still substantial diversity in the inter-

national and national radiological criteria and safety

standards applicable to the remediation of areas

affected by environmental contamination with

radionuclides. The experience of radiological

protection of the public after the Chernobyl

accident has clearly shown the need for further

international harmonization of appropriate radio-

logical criteria and safety standards.

1.4. HUMAN EXPOSURE

Following the Chernobyl accident, both

workers and the general public were affected by

radiation that resulted, or can result, in adverse

health effects. In this report consideration is given

primarily to the exposure patterns of members of

the general public exposed to radionuclides

released to the environment. Information on doses

received by members of the general public, both

those evacuated from the accident area and those

who l

ive permanently in contaminated areas, is

required for the following health related purposes:

(a) Substantiation of countermeasures and

remediation programmes;

(b) Forecast of expected adverse health effects

and justification of corresponding health

protection measures;

(d) Epidemiological and other medical studies of

radiation induced adverse health effects.

The results of post-accident environmental

monitoring indicate that the most affected countries

were Belarus, the Russian Federation and Ukraine.

Much of the information on doses from the

Chernobyl accident relates to these countries.

There were four main mechanisms for

delivering radiation dose to the public: external

dose from cloud passage, internal dose from

inhalation of the cloud and resuspended material,

external dose from radioactive material deposited

on soil and other surfaces, and internal dose from

the ingestion of food products and water. Except for

unusual circumstances, the latter two pathways

were the more important. External dose and

internal dose tended to be approximately equally

important, although this general conclusion is

subject to large variation, due to the shielding

afforded by buildings and the soil from which crops

were grown.

Estimates of doses to individual members of

population groups were based on millions of

measurements of concentrations of radioactive

material in air, soil, foods, water, human thyroids

and the whole body contents of humans. In

addition, many measurements were made of the

external gamma exposure rate over undisturbed

and disturbed fields, and external doses to humans

were measured with individual thermoluminescent

dosimeters. Thus the results of estimated doses are

firmly based upon measurements and tend to be

realistic rather than conservative.

As the major health effect of the Chernobyl

accident for the general public was an elevated

thyroid cancer incidence in children and adoles-

cents, much attention has been paid to the

dosimetry of the thyroid gland. The assessment of

thyroid doses resulting from the intake of

131

I is

based on the results of 350 000 human measure-

ments and a few thousand measurements of

131

I in

milk performed in Belarus, the Russian Federation

and Ukraine within a few weeks of the accident.

Doses to humans were reduced significantly

by a number of countermeasures. Official counter-

measures included evacuation and relocation of

persons, the blockage of contaminated food

supplies, the removal of contaminated soil, the

treatment of agricultural fields to reduce the uptake

of radionuclides, the substitution of foods and the

prohibition of the use of wild foods. Unofficial

countermeasures included the self-initiated

avoidance of foods judged to be contaminated.

1.4.1. Conclusions

The collective effective dose (not including

dose to the thyroid) received by about five million

residents living in the areas of Belarus, the Russian

Federation and Ukraine contaminated by the

Chernobyl accident (

137

Cs deposition on soil

>37 kBq/m

) was approximately 40 000 man Sv

during the period 1986–1995. The groups of exposed

persons within each country received an approxi-

mately equal collective dose. The additional amount

of collective effective dose projected to be received

between 1996 and 2006 is about 9000 man Sv.

The collective dose to the thyroid was nearly

2×10

man Gy, with nearly half received by persons

exposed in Ukraine.

The main pathways leading to human

exposure were external exposure from radio-

nuclides deposited on the ground and the ingestion

of contaminated terrestrial food products.

Inhalation and ingestion of drinking water, fish and

products contaminated with irrigation water were

generally minor pathways.

The range in thyroid dose in different

settlements and in all age–gender groups is large,

between less than 0.1 Gy and more than 10 Gy. In

some groups, and especially in younger children,

doses were high enough to cause both short term

functional thyroid changes and thyroid cancer in

some individuals.

The internal thyroid dose from the intake of

131

I was mainly due to the consumption of fresh

cow’s milk and, to a lesser extent, of green

vegetables; children, on average, received a dose

that was much higher than that received by adults,

because of their small thyroid masses and

consumption rates of fresh cow’s milk that were

similar to those of adults.

For populations permanently residing in

contaminated areas and exposed predominantly via

ingestion, the contribution of short lived radio-

iodines (i.e.

132

133

I and

135

I) to thyroid dose was

minor (i.e. about 1% of the

131

I thyroid dose), since

short lived radioiodines decayed during transport of

the radioiodines along the food chains. The highest

relative contribution (20–50%) to the thyroid doses

to the public from short lived radionuclides was

received by the residents of Pripyat through

inhalation; these residents were evacuated before

they could consume contaminated food.

Both measurement and modelling data show

that the urban population was exposed to a lower

external dose by a factor of 1.5–2 compared with

the rural population living in areas with similar

levels of radioactive contamination. This arises

because of the better shielding features of urban

buildings and different occupational habits. Also,

as the urban population depends less on local

agricultural products and wild foods than the rural

population, both effective and thyroid internal

doses caused predominantly by ingestion were

lower by a factor of two to three in the urban than

in the rural population.

The initial high rates of exposure declined

rapidly due to the decay of short lived radionuclides

and to the movement of radiocaesium into the soil

profile. The latter caused a decrease in the rate of

external dose due to increased shielding. In

addition, as caesium moves into the soil column it

binds to soil particles, which reduces the availability

of caesium to plants and thus to the human food

chain.

The great majority of dose from the accident

has already been accumulated.

Persons who received effective doses (not

including dose to the thyroid) higher than the

average by a factor of two to three were those who

lived in rural areas in single storey homes and who

ate large amounts of wild foods such as game meats,

mushrooms and berries.

The long term internal doses to residents of

rural settlements strongly depend on soil properties.

Contributions due to internal and external exposure

are comparable in areas with light sandy soil, and

the contribution of internal exposure to the total

(external and internal) dose does not exceed 10% in

areas with predominantly black soil. The contri-

bution of

Sr to the internal dose, regardless of

natural conditions, is usually less than 5%.

The long term internal doses to children

caused by ingestion of food containing caesium

radionuclides are usually lower by a factor of about

1.1–1.5 than those to adults and adolescents.

Both accumulated and predicted mean doses

in settlement residents vary in the range of two

orders of magnitude, depending on the radioactive

contamination of the area, predominant soil type

and settlement type. In the period 1986–2000 the

accumulated dose ranged from 2 mSv in towns

located in black soil areas up to 300 mSv in villages

located in areas with podzol sandy soil. The doses

expected in the period 2001–2056 are substantially

lower than the doses already received (i.e. in the

range of 1–100 mSv).

If countermeasures had not been applied, the

populations of some of the more contaminated

villages could have received lifetime (70 years)

effective doses of up to 400 mSv. Intensive

application of countermeasures such as settlement

decontamination and agricultural countermeasures

has substantially reduced the doses. For

comparison, a worldwide average lifetime dose

from natural background radiation is about

170 mSv, with a typical range of 70–700 mSv in

various regions of the world.

The vast majority of the approximately five

million people residing in the contaminated areas of

Belarus, the Russian Federation and Ukraine

currently receive annual effective doses of less than

1 mSv (equal to the national action levels in the

three countries). For comparison, a worldwide

average annual dose from natural background

radiation is about 2.4 mSv, with a typical range of 1–

10 mSv in various regions of the world.

The number of residents of the contaminated

areas in the three most affected countries that

currently receive more than 1 mSv annually can be

estimated to be about 100 000 persons. As the future

reduction of both the external dose rate and the

radionuclide (mainly

137

Cs) activity concentrations

in food is predicted to be rather slow, the reduction

in the human exposure levels is also expected to be

slow (i.e. about 3–5%/a with current counter-

measures).

Based upon available information, it does not

appear that the doses associated with hot particles

were significant.

The assessment of the Chernobyl Forum

agrees with that of UNSCEAR [1.1] in terms of the

dose received by the populations of the three most

affected countries: Belarus, the Russian Federation

and Ukraine.

1.4.2. Recommendations

Large scale monitoring of foodstuffs, whole

body counting of individuals and provision of

thermoluminescent dosimeters to members of the

general public are no longer necessary. The critical

groups in areas of high contamination and/or high

transfer of radiocaesium to foods are known.

Representative members of these critical groups

should be monitored with dosimeters for external

dose and with whole body counting for internal

dose.

Sentinel or marker individuals in more highly

contaminated areas not scheduled for further

remediation might be identified for continued

periodic whole body counting and monitoring for

external dose. The goal would be to follow the

expected continued decrease in external and

internal dose rate and to determine whether such

decreases are due to radioactive decay alone or to

further ecological elimination.

1.5. RADIATION INDUCED EFFECTS ON

PLANTS AND ANIMALS

The biological effects of radiation on plants

and animals have long been of interest to scientists;

in fact, much of the information on the effects on

humans has evolved from experimental studies on

plants and animals. Additional research followed

the development of nuclear energy and concerns

about the possible impacts of radioactive releases

into the terrestrial and aquatic environments. By

the mid-1970s, a large amount of information had

been accrued on the effects of ionizing radiation on

plants and animals.

The Chernobyl nuclear accident in April 1986

occurred not in a desert or ocean but in a territory

with a temperate climate and flourishing flora and

fauna. Both acute radiation effects (radiation death

of plants and animals, loss of reproduction, etc.) and

long term effects (change of biodiversity,

cytogenetic anomalies, etc.) have been observed in

the affected areas. Biota located in the area nearest

to the source of the radioactive release, the 30 km

zone or Chernobyl exclusion zone (CEZ), were

most affected. As a result, in this area population

and ecosystem effects on biota, caused, on the one

hand, by high radiation levels, and, on the other

hand, by plant succession and animal migration due

to intraspecific and interspecific competition, have

occurred.

The plant and animal conditions in the CEZ

changed rapidly during the first months and years

after the accident and later arrived at a quasi-

stationary equilibrium. At present, traces of adverse

radiation effects on biota can hardly be found in the

near vicinity of the radiation source (a few

kilometres from the damaged reactor), and on the

rest of the territory both wild plants and animals are

flourishing because of the removal of the major

natural stressor: humans.

1.5.1. Conclusions

Radiation from radionuclides released by the

Chernobyl accident caused numerous acute adverse

effects in the biota located in the areas of highest

exposure (i.e. up to a distance of a few tens of

kilometres from the release point). Beyond the

CEZ, no acute radiation induced effects on biota

have been reported.

The environmental response to the Chernobyl

accident was a complex interaction among radiation

dose, dose rate and its temporal and spatial

variations, and the radiosensitivities of the different

taxons. Both individual and population effects

caused by radiation induced cell death have been

observed in plants and animals as follows:

(a) Increased mortality of coniferous plants, soil

invertebrates and mammals;

(b) Reproductive losses in plants and animals;

(mammals, birds, etc.).

No adverse radiation induced effects have

been reported in plants and animals exposed to a

cumulative dose of less than 0.3 Gy during the first

month after the radionuclide fallout.

Following the natural reduction of exposure

levels due to radionuclide decay and migration,

populations have been recovering from the acute

radiation effects. By the next growing season after

the accident, the population vi

ability of plants and

animals substantially recovered as a result of the

combined effects of reproduction and immigration.

A few years were needed for recovery from the

major radiation induced adverse effects in plants

and animals.

The acute radiobiological effects observed in

the Chernobyl accident area are consistent with

radiobiological data obtained in experimental

studies or observed in natural conditions in other

areas affected by ionizing radiation. Thus rapidly

developing cell systems, such as meristems of plants

and insect larvae, were predominantly affected by

radiation. At the organism level, young plants and

animals were found to be the most sensitive to the

acute effects of radiation.

Genetic effects of radiation, in both somatic

and germ cells, were observed in plants and animals

in the CEZ during the first few years after the

accident. Both in the CEZ and beyond, different

cytogenetic anomalies attributable to radiation

continue to be reported from experimental studies

performed on plants and animals. Whether the

observed cytogenetic anomalies have any

detrimental biological significance is not known.

The recovery of affected biota in the CEZ has

been confounded by the overriding response to the

removal of human activities (e.g. termination of

agricultural and industrial activities and the

accompanying environmental pollution in the most

affected area). As a result, the populations of many

plants and animals have expanded, and the present

environmental conditions have had a positive

impact on the biota in the CEZ.

1.5.2. Recommendations for future research

In order to develop a system of environmental

protection against radiation, the long term impact

of radiation on plant and animal populations should

be further investigated in the CEZ; this is a globally

unique area for radioecological and radiobiological

research in an otherwise natural setting.

In particular, multigenerational studies of

radiation effects on the genetic structure of plant

and animal populations might bring fundamentally

new scientific information.

There is a need to develop standardized

methods for biota–dose reconstruction, for example

in the form of a unified dosimetric protocol.

1.5.3. Recommendations for countermeasures

and remediation

Protective actions for farm animals in the

event of a nuclear or radiological emergency should

be developed and internationally harmonized based

on modern radiobiological data, including the

experience gained in the CEZ.

It is likely that any technology based

remediation actions aimed at improving the radio-

logical conditions for plants and animals in the CEZ

would have adverse impacts on biota.

1.6. ENVIRONMENTAL AND

RADIOACTIVE WASTE MANAGEMENT

ASPECTS OF THE DISMANTLING OF

THE CHERNOBYL SHELTER

1.6.1. Conclusions

The accidental destruction of unit 4 of the

Chernobyl nuclear power plant resulted in

extensive radioactive contamination and the

generation of large amounts of radioactive waste in

the unit, the Chernobyl nuclear power plant site and

the surrounding area (CEZ). Construction of the

shelter between May and November 1986 was

aimed at environmental containment of the

damaged reactor, reduction of radiation levels on

the site and the prevention of further release of

radionuclides off the site.

The shelter was erected in an extremely short

period of time under conditions of severe radiation

exposure of personnel. As a result, the measures

taken to save time and reduce dose during the

construction led to imperfection in the newly

constructed shelter as well as to a lack of compre-

hensive data on the stability of the damaged unit 4

structures. In addition to uncertainties on stability

at the time of its construction, structural elements of

the shelter have degraded as a result of moisture

induced corrosion during the two decades that have

passed since the shelter was erected. The main

potential hazard associated with the shelter is a

possible collapse of its top structures and release of

radioactive dust into the environment.

In order to avoid the potential collapse of the

shelter in the future, measures are planned to

strengthen the unstable structures of the shelter. In

addition, a new safe confinement (NSC) with more

than 100 years of service life is planned to be built as

a cover over the existing shelter as a longer term

solution. The construction of the NSC is expected to

allow for the dismantlement of the current shelter,

removal of highly radioactive fuel-containing

material (FCM) from unit 4 and eventual

decommissioning of the damaged reactor.

In the course of remediation activities, both at

the Chernobyl nuclear power plant site and in its

vicinity, large volumes of radioactive waste were

generated as a result of the cleanup of contaminated

areas and placed in temporary near surface waste

storage and disposal facilities. Facilities of the

trench and landfill type were created from 1986 to

1987 in the CEZ at distances of 0.5–15 km from the

nucl

ear power plant site with the intention of

avoiding dust spread, reducing the radiation levels

and enabling better working conditions at unit 4 and

in its surroundings. These facilities were established

without proper design documentation, engineered

barriers or hydrogeological investigations and do

not meet current waste safety requirements.

During the years following the accident, large

resources were expanded to provide a systematic

analysis and an acceptable strategy for the

management of the existing radioactive waste.

However, to date, a broadly accepted strategy for

radioactive waste management at the Chernobyl

nuclear power plant site and the CEZ, and

especially for high level and long lived waste, has

not been developed. The main reason for this is the

large number of radioactive waste storage and

disposal facilities, of which only half are well studied

and inventoried. This results in large uncertainties

on the radioactive waste inventories.

More radioactive waste is expected in the

years to come, generated during the construction of

the NSC, the possible shelter dismantling, FCM

removal and the decommissioning of unit 4. This

waste, belonging to different categories, must be

properly managed.

According to the Ukrainian national

programme on radioactive waste management,

there are different options for the different waste

categories. The planned options for low level

radioactive waste are to sort the waste according to

its physical characteristics (e.g. soil, concrete, metal)

and possibly decontaminate and/or condition it for

beneficial reuse (reuse of soil for NSC foundations,

melting of metal pieces, etc.), or send it for disposal.

The long lived waste is planned to be placed in

interim storage. Different storage options are being

considered, and a decision has not yet been made.

After construction of the NSC and decommis-

sioning of the shelter facilities, it is envisaged that

shelter dismantling and further removal of FCM

will occur. High level radioactive waste is planned

to be partially processed in place and then stored at

a temporary storage site until a deep geologic

disposal site is ready.

Such a strategic approach is foreseen by the

Comprehensive Programme on Radioactive Waste

Management, which was approved by the

Ukrainian Government in 1996 and confirmed in

2004. According to this concept, it is considered

reasonable to begin a specific investigation for

exploring the most appropriate geological site in

this area in 2006. Following such planning, the

construction of a deep geologic disposal facility

might be completed before 2035–2040.

The future development of the CEZ as an

industrial site or nature reserve depends on the

future strategy for the conversion of unit 4 into an

ecologically safe system (i.e. the development of the

NSC, the dismantlement of the current shelter, the

removal of FCM and the eventual decommissioning

of the unit 4 reactor site). Currently units 1, 2 and 3

(1000 MW RBMK (high power channel type)

reactors) are shut down with a view to being decom-

missioned, and two additional reactors (units 5 and

6) that had been near completion were abandoned

in 1986 following the accident.

There are uncertainties related to the current

radioactive material inventory at the shelter and

also at the waste storage and disposal sites within

the CEZ. This situation affects not only safety

assessments and environmental analyses but also

the design of remediation actions and the criteria

for new facilities.

1.6.2. Recommendations for future actions

Recognizing the ongoing effort on improving

safety and addressing the aforementioned uncer-

tainties in the existing input data, the following

main recommendations are made regarding the

dismantling of the shelter and the management of

the radioactive waste generated as a result of the

accident.

Since individual safety and environmental

assessments have been performed only for

individual facilities at and around the Chernobyl

nuclear power plant, a comprehensive safety and

environmental impact assessment, in accordance

with international standards and recommendations,

that encompasses all activ

ities inside the entire

CEZ, should be performed.

During the preparation and construction of

the NSC and soil removal, special monitoring wells

are expected to be destroyed. Therefore, it is

important to maintain and improve the environ-

mental monitoring strategies, methods, equipment

and staff qualification needed for the adequate

performance of monitoring of the conditions at the

Chernobyl nuclear power plant site and the CEZ.

Development of an integrated radioactive

waste management programme for the shelter, the

Chernobyl nuclear power plant site and the CEZ is

needed to ensure application of consistent

management approaches and sufficient facility

capacity for all waste types. Specific emphasis needs

to be given to the characterization and classification

of waste (in particular waste with transuranic

elements) from all remediation and decommis-

sioning activities, as well as to the establishment of

sufficient infrastructure for the safe long term

management of long lived and high level waste.

Therefore, development of an appropriate waste

management infrastructure is needed in order to

ensure sufficient waste storage capacity; at present,

the rate and continuity of remediation activities at

the Chernobyl nuclear power plant site and in the

CEZ are being limited.

A coherent and comprehensive strategy for

the rehabilitation of the CEZ is needed, with

particular focus on improving the safety of the

existing waste storage and disposal facilities. This

will require development of a prioritization

approach for remediation of the sites, based on

safety assessment results, aimed at making decisions

about those sites at which waste will be retrieved

and disposed of and those sites at which the waste

will be allowed to decay in situ.

REFERENCE TO SECTION 1

[1.1] UNITED NATIONS, Sources and Effects of

Ionizing Radiation (Report to the General

Assembly, with Scientific Annexes), Vol. II, Scien-

tific Committee on the Effects of Atomic

Radiation (UNSCEAR), UN, New York (2000).

2. INTRODUCTION

2.1. BACKGROUND

The Chernobyl Forum was initiated by the

IAEA in cooperation with the Food and

Agriculture Organization of the United Nations

(FAO), the United Nations Development

Programme (UNDP), the United Nations

Environment Programme (UNEP), the United

Nations Office for the Coordination of Humani-

tarian Affairs (OCHA), the United Nations

Scientific Committee on the Effects of Atomic

Radiation (UNSCEAR), the World Health Organi-

zation (WHO) and the World Bank, as well as the

competent authorities of Belarus, the Russian

Federation and Ukraine. The organizational

meeting for the Chernobyl Forum was held on 3–

5 February 2003, at which time the decision was

taken to establish the Forum as an ongoing entity of

the above named organizations.

The background for the establishment of the

Forum dates back to 2000, when UNSCEAR

published its 2000 report to the United Nations

General Assembly [2.1]. In this report it was stated

that, apart from the very early deaths due to

extreme overexposure, the only clearly indicated

health effect on the population that could be

attributed to radiation exposure was an increased

rate in the diagnosis of thyroid cancer among

persons who were young children at the time of

exposure. The political representatives of Belarus,

the Russian Federation and Ukraine had strong

reservations regarding the report. These reserva-

tions seem to have had two bases:

(a) The statement on health effects was widely

divergent from what was being reported in the

popular press and even by some other

members of the United Nations family;

(b) The political representatives felt that the

views of scientists from the three affected

countries had not been considered by

UNSCEAR.

Subsequently, during his visit to Belarus and in

meetings with the Belarusian authorities and its

scientific community, the Director General of the

IAEA, M. ElBaradei, noted that “a lack of trust still

prevails among the people of the region…, due in

part to the contradictory data and reports — on the

precise environmental and health impacts of the

accident — among national authorities, as well as

among the relevant international organizations.”

This was in general agreement with the stated views

of the political authorities of the three countries. It

was evident that the authorities desired a new

opportunity for an exchange of views and for the

discussion of issues such as optimization of activities

related to the remediation of contaminated land

and the provision of health care to those affected by

the accident. During meetings with these represent-

atives, the Director General of the IAEA indicated

his support for the concept of a Chernobyl Forum as

a joint activity of the United Nations family and the

three affected countries.

2.2. OBJECTIVES OF THE CHERNOBYL

FORUM

At the organizational meeting of the Forum it

was decided to establish the Chernobyl Forum as a

series of managerial, expert and public meetings for

the purpose of generating authoritative consensual

statements on the health effects attributable to

radiation exposure arising from the accident and on

the environmental consequences induced by the

released radioactive material, to provide advice on

remediation and special health care programmes,

and to suggest areas in which further research is

required.

Participants at the organization meeting

accepted the terms of reference of the Forum as

being:

(a) To explore and refine the current scientific

assessments on the long term health and

environmental consequences of the

Chernobyl accident, with a view to producing

authoritative consensus statements focusing

on:

(i) The health effects attributable to

radiation exposure caused by the

accident;

(ii) The environmental consequences

induced by the radioactive material

released due to the accident (e.g. contam-

ination of foodstuffs);

(iii) The consequences attributable to the

accident but not directly related to the

radiation exposure or radioactive

contamination.

(b) To identify gaps in scientific research relevant

to the radiation induced or radioactive

contamination induced health and environ-

mental impacts of the accident, and to suggest

areas in which further work is required based

on an assessment of the work done in the past

and bearing in mind the ongoing work and

projects.

mentation of, scientifically sound programmes

on mitigation of the accident consequences,

including possible joint actions of the organi-

zations participating in the Forum, such as:

(i) Remediation of contaminated land, with

the aim of making it suitable for normal

agricultural, economic and social life

under safe conditions;

(ii) Special health care of the affected

population;

(iii) Monitoring of long term human exposure

to radiation;

(iv) Addressing the environmental issues

pertaining to the decommissioning of the

Chernobyl shelter and the management

of radioactive waste originating from the

Chernobyl accident.

2.3. METHOD OF OPERATION AND

OUTPUT OF THE CHERNOBYL FORUM

The Chernobyl Forum is a high level organi-

zation of senior officials of United Nations agencies

and the three affected countries. The technical

reports of the Forum were produced by two expert

groups: Expert Group ‘Environment’ (EGE) and

Expert Group ‘Health’ (EGH). The membership of

the two groups comprised recognized international

scientists and experts from the three affected

countries. Through the work of these two groups

and their subworking groups, the technical

documents were prepared. The EGE was

coordinated by the IAEA and the EGH was

coordinated by the WHO.

The documents were produced through

meetings of groups of experts on specific topics. The

groups considered in detail the data available from

the literature as well as unpublished data from the

three most affected countries. The documents were

used as the basis for the final reports of the

Chernobyl Forum, which were approved by the

Forum itself.

This report is the Chernobyl Forum’s report

on the environmental consequences of the

Chernobyl accident. The Chernobyl Forum’s report

on the health effects of the Chernobyl accident will

be published by the WHO [2.2].

2.4. STRUCTURE OF THE REPORT

This report consists of seven sections.

Following the Introduction, Section 3 describes the

processes and patterns of radioactive contamination

of the urban, agricultural, forest and aquatic

environments as a result of the deposition of the

Chernobyl release. Section 4 identifies the major

environmental countermeasures and remediation

measures applied to the aforementioned four

environments in order to mitigate the accident’s

consequences and, specifically, to reduce human

exposure. Section 5 deals with an assessment of

human exposure to radiation within the affected

areas, based on data on environmental radioactive

contamination and the countermeasures presented

in Sections 3 and 4. Section 6 presents an overview

of experimental data on the radiation induced

effects in plants and animals observed predomi-

nantly in the near zone of radioactive contami-

nation. Finally, Section 7 discusses the

environmental aspects of the dismantling of the

shelter facility at the Chernobyl site and radioactive

waste management in the CEZ.

Each section is completed with relevant

conclusions and recommendations for future

environmental remediation actions, monitoring and

research. The entire report is preceded by Section 1,

the Summary.

REFERENCES TO SECTION 2

[2.1] UNITED NATIONS, Sources and Effects of

Ionizing Radiation (Report to the General

Assembly, with Scientific Annexes), Vol. II, Scien-

tific Committee on the Effects of Atomic

Radiation (UNSCEAR), UN, New York (2000)

451–566.

[2.2] WORLD HEALTH ORGANIZATION, Health

Effects of the Chernobyl Accident and Special

Health Care Programmes, Report of the

Chernobyl Forum Expert Group “Health”

(EGH), WHO, Geneva (in press).

3. RADIOACTIVE CONTAMINATION

OF THE ENVIRONMENT

The accident at the Chernobyl nuclear power

plant resulted in a substantial release of radionu-

clides to the atmosphere and caused extensive

contamination of the environment. A number of

European countries were subjected to radioactive

contamination; among the most affected were three

former republics of the USSR, now Belarus, the

Russian Federation and Ukraine. The activity levels

of the radionuclides in the environment gradually

declined due to radioactive decay. At the same time,

there was movement of the radionuclides within the

environments — atmospheric, aquatic, terrestrial

and urban — and among the environments. The

processes that determined the patterns of

radioactive contamination in those environments

are presented in this section.

The focus of this section is mainly on

radioactive contamination of the off-site

environment. Significant attention is given in

Section 7 to the Chernobyl nuclear power plant site,

the CEZ and the Chernobyl shelter.

3.1. RADIONUCLIDE RELEASE AND

DEPOSITION

3.1.1. Radionuclide source term

The accident at unit 4 of the Chernobyl

nuclear power plant took place shortly after

midnight on 26 April 1986. Prior to the accident, the

reactor had been operated for many hours in non-

design configurations in preparation for an

experiment on recovery of the energy in the turbine

in the event of an unplanned shutdown. The cause

of the accident is rather complicated, but can be

considered as a runaway surge in the power level

that caused the water coolant to vaporize inside the

reactor. This in turn caused a further increase in the

power level, with a resulting steam explosion that

destroyed the reactor. After the initial explosion,

the graphite in the reactor caught fire. Despite the

heroic efforts of the staff to control the fire, the

graphite burned for many days, and releases of

radioactive material continued until 6 May 1986.

The reconstructed time course of the release of

radioactive material is shown in Fig. 3.1 [3.1–3.3].

The occurrence of the accident was not

immediately announced by the authorities of the

then USSR. However, the releases were so large

that the presence of fresh fission products was soon

detected in Scandinavian countries, and retro-

spective calculations of possible trajectories

indicated that the accident had occurred in the

former USSR. Further details of the accident and its

immediate consequences are available in reports by

the International Nuclear Safety Advisory Group

[3.1], the International Advisory Committee [3.4]

and UNSCEAR [3.5, 3.6].

An early estimate of the amount of

137

released by the accident and deposited in the

former USSR was made based on an airborne

radiometric measurement of the contaminated parts

of the former USSR; this estimate indicated that

about 40 PBq (1 × 10

Ci) was deposited. Estimates

of the releases have been refined over the years, and

the current estimate of the total amount of

137

deposited in the former USSR is about twice the

earlier estimate (i.e. 80 PBq). Current estimates of

the amounts of the more important radionuclides

released are shown in Table 3.1. Most of the radio-

nuclides for which there were large releases have

short physical half-lives, and the radionuclides with

long half-lives were mostly released in small

0.00

0.10

0.20

0.30

0.40

0.50

0.60

0.70

1 2 3 4 5 6 7 8 9 10 11

Days after initiation of the accident on 26 April 1986

Cooldown

period

Heatup period

Sharp drop

Release rate (EBq/d)

0.002–0.006 EBq/d

FIG. 3.1. Daily release rate to the atmosphere of radioac-

tive material, excluding noble gases, during the Chernoby

accident. The values are decay corrected to 6 May 1986 and

are uncertain by ±50% [3.1].

amounts. In the early period after the accident, the

radionuclide of most radiological concern was

131

later, the emphasis shifted to

137

Cs.

By 2005 most of the radionuclides released by

the accident had already decayed below levels of

concern. Interest over the next few decades will

continue to be on

137

Cs and, to a lesser extent,

Sr;

the latter remains more important in the near zone

of the Chernobyl nuclear power plant. Over the

longer term (hundreds to thousands of years), the

only radionuclides anticipated to be of interest are

the plutonium isotopes. The only radionuclide

TABLE 3.1. REVISED ESTIMATES OF THE PRINCIPAL RADIONUCLIDES

RELEASED DURING THE COURSE OF THE CHERNOBYL ACCIDENT

Half-life Activity released (PBq)

Inert gases

Krypton-85 10.72 a 33

Xenon-133 5.25 d 6500

Volatile elements

Tellurium-129m 33.6 d 240

Tellurium-132 3.26 d ~1150

Iodine-131 8.04 d ~1760

Iodine-133 20.8 h 910

Caesium-134 2.06 a ~47.

Caesium-136 13.1 d 36

Caesium-137 30.0 a ~85

Elements with intermediate volatility

Strontium-89 50.5 d ~115

Strontium-90 29.12 a ~10

Ruthenium-103 39.3 d >168

Ruthenium-106 368 d >73

Barium-140 12.7 d 240

Refractory elements (including fuel particles)

Zirconium-95 64.0 d 84

Molybdenum-99 2.75 d >72

Cerium-141 32.5 d 84

Cerium-144 284 d ~50

Neptunium-239 2.35 d 400

Plutonium-238 87.74 a 0.015

Plutonium-239 24 065 a 0.013

Plutonium-240 6 537 a 0.018

Plutonium-241 14.4 a ~2.6

Plutonium-242 376 000 a 0.00004

Curium-242 18.1 a ~0.4

Most of the data are from Refs [3.6, 3.7].

Based on

134

Cs/

137

Cs ratio of 0.55 as of 26 April 1986 [3.8].

Based on fuel particle release of 1.5% [3.9].

expected to increase in its levels in the coming years

241

Am, which arises from the decay of

241

Pu; it

takes about 100 years for the maximum amount of

241

Am to form from

241

Pu.

3.1.2. Physical and chemical forms of released

material

Radionuclides in the releases from the

stricken reactor were in the form of gases,

condensed particles and fuel particles. The presence

of the latter was an important characteristic of the

accident. The oxidation of nuclear fuel was the basic

mechanism of fuel particle formation. Less oxidized

fuel particles were formed as a result of the initial

explosion and were released primarily towards the

western direction. More oxidized and soluble

particles predominated in the remaining fallout,

which was deposited in many other areas.

During oxidation and dispersal of the nuclear

fuel, volatilization of some radionuclides took place.

After the initial cloud cooled, the more volatile of

the released radionuclides remained in the gas

phase, whilst the less volatile radionuclides

condensed on particles of construction material,

soot and dust. Thus the chemical and physical forms

of the radionuclides in the Chernobyl release were

determined by the volatility of their compounds and

the conditions inside the reactor. Radioactive

compounds with relatively high vapour pressure

(primarily isotopes of inert gases and iodine in

different chemical forms) were transported in the

atmosphere in the gas phase. Isotopes of refractory

elements (e.g. cerium, zirconium, niobium and

plutonium) were released into the atmosphere

primarily in the form of fuel particles. Other radio-

nuclides (isotopes of caesium, tellurium, antimony,

etc.) were found in both fuel and condensed

particles. The relative contributions of condensed

and fuel components in the deposition at a given

site can be estimated from the activity ratios of

radionuclides of different volatility classes.

Fuel particles made up the most important

part of the fallout in the vicinity of the release

source. Radionuclides such as

Zr,

Nb,

Mo,

141,144

Ce,

154,155

Eu,

237,239

Np,

238–242

Pu,

241,243

Am and

242,244

Cm were released in a matrix of fuel particles

only. More than 90% of

89,90

Sr and

103,106

Ru activities

was also released in fuel particles. The release

fraction of

Sr,

154

Eu,

238

Pu,

239,240

Pu and

241

Am,

and, therefore, of the nuclear fuel itself, deposited

outside the Chernobyl nuclear power plant

industrial site has been recently estimated to be

only 1.5% ± 0.5% [3.9], which is half that of earlier

estimates [3.1].

The chemical and radionuclide composition of

fuel particles was close to that of irradiated nuclear

fuel, but with a lower fraction of volatile radionu-

clides, a higher oxidation state of uranium and the

presence of various admixtures, especially in the

surface layer. In contrast, the chemical and radionu-

clide composition of condensed particles varied

widely. The specific activity of the radionuclides in

these particles was determined by the duration of

the condensation process and the process temper-

ature, as well as by the particle characteristics. The

radionuclide content of some of the particles was

dominated by just one or two nuclides, for example

103,106

Ru or

140

Ba/

140

La [3.10].

The form of a radionuclide in the release

determined the distance of its atmospheric

transport. Even the smallest fuel particles consisting

of a single grain of nuclear fuel crystallite had a

relatively large size (up to 10 µm) and high density

(8–10 g/cm

). Owing to their size, they were

transported only a few tens of kilometres. Larger

aggregates of particles were found only within

distances of several kilometres from the power

plant. For this reason, the deposition of refractory

radionuclides strongly decreased with distance from

the damaged reactor, and only traces of refractory

elements could be found outside the industrial site

of the power plant. In contrast, significant

deposition of gaseous radionuclides and sub-

micrometre condensed particles took place

thousands of kilometres from Chernobyl.

Ruthenium particles, for example, were found

throughout Europe [3.11]. At distances of hundreds

of kilometres from Chernobyl the deposition of

137

Cs was as high as 1 MBq/m

[3.12, 3.13].

Another important characteristic of fallout is

related to its solubility in aqueous solutions. This

determines the mobility and bioavailability of

deposited radionuclides in soils and surface waters

during the initial period after deposition. In fallout

sampled at the Chernobyl meteorological station

from 26 April to 5 May 1986 with a 24 h sampling

period, the water soluble and exchangeable

(extractable with 1M CH

COONH

) forms of

137

varied from 5% to more than 30% [3.14]. The water

soluble and exchangeable forms of

Sr in deposits on

26 April accounted for only about 1% of the total;

this value increased to 5–10% in subsequent days.

The low solubility of deposited

137

Cs and

near the nuclear power plant indicates that fuel

particles were the major part of the fallout, even at

20 km from the source. At shorter distances the

portion of water soluble and exchangeable forms of

137

Cs and

Sr was, obviously, lower, due to the

presence of larger particles; at longer distances the

fraction of soluble condensed particles increased.

As one example, almost all the

137

Cs deposited in

1986 in the United Kingdom was water soluble and

exchangeable [3.15].

3.1.3. Meteorological conditions during the

course of the accident

At the time of the accident the weather in

most of Europe was dominated by a vast anti-

cyclone. At the 700–800 m and 1500 m altitudes, the

area of the Chernobyl nuclear power plant was at

the south-west periphery of a high atmospheric

pressure zone with air masses moving north-west

with a velocity of 5–10 m/s [3.12].

At daybreak, the altitude of the air mixing

layer was about 2500 m. This resulted in rapid

mixing of the airborne debris throughout the mixing

layer and dispersion of the cloud at different layers

of the mixing height. Further dissemination of the

particles originating from the time of the accident

within the 700–1500 m layer occurred as the air

mass moved towards the north-east, with a

subsequent turn to the north; this plume was

detected in Scandinavian countries.

Ground level air on 26 April was transported

to the west and north-west and reached Poland and

the Scandinavian countries by 27–29 April. In

southern and western Ukraine, the Republic of

Moldova, Romania, Slovakia and Poland the

weather was influenced by a low gradient pressure

field. In the following days the cyclone moved

slowly south-east and the low gradient pressure

field with several poorly defined pressure areas

dispersed over the major part of the European

sector of the former USSR. One of the pressure

areas was a small near surface cyclone located on

the morning of 27 April south of Gomel.

Later, the releases from the reactor were

carried predominantly in the south-western and

southern directions until 7–8 May. During the first

five days after the accident commenced, the wind

pattern had changed through all directions of the

compass [3.12].

Within a few days after the accident,

measurements of radiation levels in air over Europe,

Japan and the USA showed the presence of

radionuclides at altitudes of up to 7000 m. The force

of the explosion, rapid mixing of air layers due to

thunderstorms near the Chernobyl nuclear power

plant and the presence of warm frontal air masses

between the Chernobyl nuclear power plant and the

Baltic Sea all contributed to the transport of

radionuclides to such heights.

To understand the complex meteorological

situation better, Borzilov and Klepikova [3.16]

carried out calculations with assumed input pulses

of unit activity at various times of the accident. The

height of the source was selected to be 1000 m until

14:00 (GMT) on 28 April, and later 500 m. The

results of calculations are presented in Fig. 3.2 for

six time periods (GMT time) with differing long

range transport conditions as follows:

(1) From the start of the accident to 12:00 (GMT)

on 26 April: towards Belarus, Lithuania, the

Kaliningrad region (of the Russian Feder-

ation), Sweden and Finland.

(2) From 12:00 on 26 April to 12:00 on 27 April: to

Polessye, then Poland and then south-west.

(3) From 12:00 on 27 April to 29 April: to the

Gomel (Belarus) region, the Bryansk

(Russian Federation) region and then the east.

(4) 29 April to 30 April: to the Sumy and Poltava

regions (Ukraine) and towards Romania.

(5) 1–3 May: to southern Ukraine and across the

Black Sea to Turkey.

(6) 4–5 May: to western Ukraine and Romania,

and then to Belarus.

Atmospheric precipitation plays an important

role in determining whether an area might receive

heavy contamination, as the processes of rainout

(entrainment in a storm system) and washout (rain

falling through a contaminated air mass) are

important mechanisms in bringing released material

to the ground. In particular, significant heteroge-

neity in the deposition of radioactive material is

related to the presence or absence of precipitation

during passage of the cloud. Also, there are

differences in behaviour regarding how effectively

different radionuclides, or chemical forms of the

same radionuclide, are rained or washed out.

There were many precipitation events during

the course of the accident, and these events

produced some areas of high ground deposition at

distances far from the reactor. An example of the

complex precipitation situation during the accident

is shown in Fig. 3.3, which is a map of average daily

precipitation intensity on 29 April for the parts of

Belarus, the Russian Federation and Ukraine most

heavily affected by the accident.

In the case of dry deposition, the contami-

nation levels were lower, but the radionuclide

mixture intercepted by vegetation was substantially

enriched with radioiodine isotopes; in the case of

wet deposition, the radionuclide content in the

fallout was similar to that in the radioactive cloud.

As a result, both the levels and ratios of radio-

nuclides in areas with different deposition types

varied.

3.1.4. Concentration of radionuclides in air

The activity concentrations of radioactive

material in air were measured at many locations in

the former USSR and throughout the world.

Examples of such activity concentrations in air are

shown in Fig. 3.4 for two locations: Chernobyl and

Baryshevka, Ukraine. The location of the

Chernobyl sampler was the meteorological station

in the city of Chernobyl, which is more than 15 km

south-east of the Chernobyl nuclear power plant.

The initial concentrations of airborne material were

very high, but dropped in two phases. There was a

rapid fall over a few months, and a more gradual

decrease over several years. Over the long term, the

sampler at Chernobyl records consistently higher

activity concentrations than the sampler at

Baryshevka (about 150 km south-east of the

Chernobyl nuclear power plant), presumably due to

resuspension [3.17].

Even with the data smoothed by a rolling

average, there are some notable features in the data

WAR S AW

VILNIUS

MINSK

molensk

Bryansk

Orel

Tula

Kaluga

Mogilev

Gomel

Cherkassy

Vinnitsa

Rovno

Lvov

Sumy

Brest

Chernovtsy

Kirovograd

Kharkov

300 km250200100 150050

Lublin

KIEV

Chernobyl

Chernigov

Zhitomir

FIG. 3.2. Calculated plume formation according to the meteorological conditions for instantaneous releases on the followin

dates and times (GMT): (1) 26 April 1986, 00:00; (2) 27 April, 00:00; (3) 27 April, 12:00; (4) 29 April, 00:00; (5) 2 May, 00:00;

and (6) 4 May, 12:00 [3.16].

FIG. 3.3. Map of average precipitation intensity (mm/h) on

9 April 1986 in the area near the Chernobyl nuclear powe

lant [3.12].

collected over the long term. The clearly discernible

peak that occurred during the summer of 1992

(month 78) was due to widespread forest fires in

Belarus and Ukraine.

3.1.5. Deposition of radionuclides on soil

surfaces

As already mentioned, surveys with airborne

spectrometers over large areas were undertaken

soon after the accident to measure the deposition of

137

Cs (and other radionuclides) on the soil surface in

several countries. In the mapping of the deposition,

137

Cs was chosen because it is easy to measure and is

of radiological significance. Soil deposition of

137

equal to 37 kBq/m

(1 Ci/km

) was chosen as a

provisional minimum contamination level, because:

(a) this level was about ten times higher than the

137

Cs deposition in Europe from global fallout; and

(b) at this level the human dose during the first year

after the accident was about 1 mSv and was

considered to be radiologically important.

Knowledge of the extent and spatial variation of

deposition is critical in defining the magnitude of

the accident, predicting future levels of external and

internal dose, and determining what radiation

protection measures are necessary. In addition,

many soil samples were collected and analysed at

radiological laboratories.

Thus massive amounts of data were collected

and subsequently published in the form of an atlas

that covers essentially all of Europe [3.13]. Another

atlas produced in the Russian Federation [3.12]

covers the European part of the former USSR. An

example is shown in Fig. 3.5.

It is clear from Fig. 3.5 and Table 3.2 that the

three countries most heavily affected by the

accident were Belarus, the Russian Federation and

Ukraine. From the total

137

Cs activity of about

64 TBq (1.7 MCi) deposited on European territory

in 1986, Belarus received 23%, the Russian

Federation 30% and Ukraine 18%. However, due

FIG. 3.4. Rolling seven month mean atmospheric concen-

tration of

137

Cs at Baryshevka and Chernobyl (June 1986

–

ugust 1994) [3.17].

TABLE 3.2. AREAS IN EUROPE CONTAMINATED BY CHERNOBYL FALLOUT IN 1986

[3.6, 3.13]

Area with

137

Cs deposition density range (km

)

37–185 kBq/m

185–555 kBq/m

555–1480 kBq/m

>1480 kBq/m

Russian Federation 49 800 5 700 2100 300

Belarus 29 900 10 200 4200 2200

Ukraine 37 200 3 200 900 600

Sweden 12 000 — — —

Finland 11 500 — — —

Austria 8 600———

Norway 5 200 — — —

Bulgaria 4 800 — — —

Switzerland 1 300 — — —

Greece 1 200 — — —

Slovenia 300 — — —

Italy 300———

Republic of Moldova 60 — — —

to the wet deposition processes discussed above,

there were also major contaminated areas in

Austria, Finland, Germany, Norway, Romania and

Sweden. A more detailed view of the nearby heavily

contaminated areas is shown in Fig. 3.6 [3.4].

Water and wind erosion of soil may lead to

137

Cs transfer and redistribution on a local scale at

relatively short distances. Wind erosion may also

lead to

137

Cs transfer with soil particles on a regional

scale.

Soon after the accident, a 30 km radius

exclusion zone (the CEZ) was established around

the reactor. Further relocations of populations took

place in subsequent months and years in Belarus,

the Russian Federation and Ukraine; eventually,

116 000 persons were evacuated or relocated.

The total area with

137

Cs soil deposition of

0.6 MBq/m

(15 Ci/km

) and above in 1986 was

10 300 km

, including 6400 km

in Belarus, 2400 km

in the Russian Federation and 1500 km

in Ukraine.

In total, 640 settlements with about 230 000

inhabitants were located on these contaminated

territories. Areas with

137

Cs depositions of more

than 1 Ci/km

(37 kBq/m

) are classified as radioac-

tively contaminated according to the laws on social

protection in the three most affected countries. The

number of people who were living in such contami-

nated areas in 1995 is shown in Table 3.3.

Immediately after the accident, most concern

was focused on contamination of food with

131

I. The

broad pattern of the deposition of

131

I is shown in

Fig. 3.7. Unfortunately, due to the rapid decay of

131

after its deposition, there was not enough time to

collect a large number of samples for detailed

analysis. At first, it was assumed that a strong

correlation could be assumed between depositions

131

I and

137

Cs. However, this has not been found

to be consistently valid. More recently, soil samples

have been collected and analysed for

129

I, which has

a physical half-life of 16 × 10

years and can only be

measured at very low levels by means of accelerator

mass spectrometry. Straume et al. [3.19] have

reported the successful analysis of samples taken in

Belarus, from which they have established that, at

the time of the accident, there were 15 ± 3 atoms of

129

I for each atom of

131

I. This estimated ratio

enables better estimates of the deposition of

131

I for

the purpose of reconstructing radiation doses

received by people.

Similar maps can be drawn for the other radio-

nuclides of interest shown in Table 3.1. The

deposition of

Sr is shown in Fig. 3.8. In comparison

1.08

0.27

0.054

1480

185

kBq/m

Ci/km

Total caesium-137

(nuclear weapons test,

Chernobyl, ...) deposition

Data not available

National capital

Scale 1:11 250 000

Projection: Lambert Azimuthal

800

600

400

200

kilometres 0

100

200

500

400

300

200

100

0 miles

100

200

Roshydromet

(Russia)/Minchernobyl

(Ukraine)/Belhydromet (Belarus),

1998

FIG. 3.5. Surface ground deposition of

137

Cs throughout Europe as a result of the Chernobyl accident [3.13].

with

137

Cs, (a) there was less

Sr released from the

reactor and (b) strontium is less volatile than

caesium. Thus the spatial extent of

Sr deposition

was much more confined to areas close to the

Chernobyl nuclear power plant than that of

137

Cs.

The amounts of plutonium deposited on soil have

also been measured (see Fig. 3.9). Nearly all areas

with plutonium deposits above 3.7 kBq/m

(0.1

Ci/km

) are within the CEZ.

3.1.6. Isotopic composition of the deposition

The most extensive measurements of surface

activity concentrations have been performed for

137

Cs. Values for other radionuclides, especially

134

Cs,

136

Cs,

131

133

140

Ba/

140

La,

Zr/

Nb,

103

Ru,

106

Ru,

132

Te,

125

Sb and

144

Ce, have been expressed as

ratios to the reference radionuclide,

137

Cs. These

ratios depend on the location, because of (a) the

St dohko

Styr

’nyroG

anseD

T reteve

anseD

rpenD

aksoR

FIG. 3.6. Surface ground deposition of

137

Cs in areas of Belarus, the Russian Federation and Ukraine near the accident site

[3.4].

TABLE 3.3. DISTRIBUTION OF INHABITANTS LIVING IN AREAS CONSIDERED TO BE RADIO-

ACTIVELY CONTAMINATED IN BELARUS, THE RUSSIAN FEDERATION AND UKRAINE IN 1995

[3.6]

Caesium-137 deposition

density (kBq/m

)

Thousands of inhabitants

Belarus Russian Federation Ukraine Total

37–185 1543 1654 1189 4386

185–555 239 234 107 580

555–1480 98 95 0.3 193

Total 1880 1983 1296 5159

For social and economic reasons, some people living in areas of contamination of less than 37 kBq/m

are also included.

different deposition behaviour of fuel particles,

aerosols and gaseous radionuclides and (b) the

variation in radionuclide composition with time of

release. In fact, these ratios are not necessarily

constant with time. Depending on the time of

release and the corresponding release character-

istics (e.g. temperature of the core), significant

variations in the release ratios were observed after

the Chernobyl accident [3.2, 3.20].

The first plume, which moved to the west,

carried the release that occurred during the

explosive phase, when the exposed core was not as

hot as in the later phases. The second plume, which

moved north to north-east, carried releases from a

core that was becoming increasingly hot, while the

third plume, moving mainly south, was charac-

terized by releases from a core heated to tempera-

tures above 2000°C; at such temperatures the less

volatile radionuclides, such as molybdenum,

strontium, zirconium, ruthenium and barium, are

readily released. During this phase, releases of

iodine radioisotopes also increased.

Caesium hot spots occurred in the far zone of

Belarus and in the Kaluga, Tula and Orel regions of

FIG. 3.7. Surface ground deposition of

131

I [3.18] (Ci/km

on 15 May 1986).

Zhlobin

Rogachev

Chernin

Svetilovichi

Vetka

Dobrush

Terekhovka

Lyubech

Chernobyl

Kraslikovka

Terma khovka

Varovsk

Bazar

Bogdany

Morovsk

Oster

Dov lya dy

Kirovo

Dernovichi

Polesskoye

Dronyki

Mikulichi

Narovlya

Dobryny

Nov. Radcha

Ilyincy

Zimovisce

Pirki

Slavutich

Komarin

Gdeny

Kovpyta

Krasnoje

Ivanovka

Pripyat

Repki

Loev

Bragin

Malozhin

Chemerisy

Savichi

Zabolotje

Chechersk

Sidorovichi

Svetlogorsk

Rechitsa

Mozyr

Khoyniki

Kalinkovichi

Va s il e vich i

GOMEL

CHERNIGOV

Novo Shepelichi

Denisovichi

Stracholese

-2

74-111 kBq m

-2

37-74 kBq m

-2

>1 11 kBq m

FIG. 3.8. Surface ground deposition of

Sr [3.4].

Slovecna

Kiev

Pripyat

Sepelichi

Dovlyady

Savici

Pirki

Sla vut ich

Dronyki

Kovpyta

Komarin

Gdeny

Morovsk

Chernobyl

Polesskoye

Nov. Radcha

Kirovo

Dernovichi

Dobryny

Narovlya

Bragin

Mikulichi

Malozhin

Che merisy

Borscevka

Mikhalki

Yurovichi

Khoyniki

Volchkov

T ermakhovka

ornostojpol

Krasilovka

Ilyinc y

Novo-

Ivankov

Var ov s k

Kuchari

Zarudje

Bujan

Vo r z el y

Motyzhin

Dymer

Chernin

Bazar

Bojarka

Brovary

Irpeny

KIEV

Denisovichi

Zimovishche

FIG. 3.9. Areas (orange) where the surface ground deposi-

tion of

239,240

Pu exceeds 3.7 kBq/m

[3.4].

the Russian Federation. The composition of the

deposited radionuclides in each of these highly

contaminated areas was similar. The ratios of

different radionuclides to

137

Cs as observed in

ground deposits in the different release vectors are

shown in Table 3.4.

The activity ratios for the western and

northern plumes were similar and in many cases

identical, in contrast to the ratios for the southern

plume. All activity ratios show, with the exception

132

Te/

137

Cs, a decrease with increasing distance

from the nuclear power plant. The decrease is less

profound for

Zr and

144

Ce (about a factor of three)

than with

Mo and

140

Ba (two orders of magnitude)

Sr and

103

Ru (one order of magnitude). For the

ratio

131

137

Cs only a slight decrease, by about a

factor of four, was observed over a 1000 km

distance. Within the first 200 km virtually no

variation in the ratio was observed.

3.2. URBAN ENVIRONMENT

3.2.1. Deposition patterns

Radioactive fallout resulted in long term

contamination of thousands of settlements in the

USSR and some other European countries and in

the irradiation of their inhabitants due to both

external gamma radiation and internal exposure due

to consumption of contaminated food. Near to the

Chernobyl nuclear power plant, the towns of

Pripyat and Chernobyl and some other smaller

settlements were subjected to substantial contami-

nation from an ‘undiluted’ radioactive cloud under

dry meteorological conditions, whereas more distant

settlements were significantly affected because of

precipitation at the time of cloud passage.

When the radioactive fallout was deposited on

settlements, exposed surfaces such as lawns, parks,

streets, roofs and walls became contaminated with

radionuclides. Both the activity level and elemental

composition of the radioactive fallout was signifi-

cantly influenced by the type of deposition

mechanism, namely wet deposition with precipi-

tation or dry deposition influenced by atmospheric

mixing, diffusion and chemical adsorption. Surfaces

such as trees, bushes, lawns and roofs become

relatively more contaminated under dry conditions

than when there is precipitation. Under wet

conditions, horizontal surfaces, including soil plots

and lawns (see Fig. 3.10), receive the highest levels

of contamination. Particularly high

137

Cs activity

concentrations have been found around houses

where rain has transported radioactive material

from roofs to the ground.

TABLE 3.4. ESTIMATED RELATIVE SURFACE ACTIVITY CONCENTRATION OF DIFFERENT

RADIONUCLIDES AFTER RELEASE FROM THE CHERNOBYL NUCLEAR POWER PLANT

(26 APRIL 1986) [3.2]

Half-life

Activity per unit area relative to

137

Western plume

(near zone)

Northern plume

(near zone)

Southern plume

(near zone)

At caesium hot spots

(far zone)

Strontium-90 28.5 a 0.5 0.13 1.5 0.014

Zirconium-95 64.0 d 5 3 10 0.06

Molybdenum-99 66.0 h 8 3 25 0.11

Ruthenium-103 39.35 d 4 2.7 12 1.9

Tellurium-132 78.0 h 15 17 13 13

Iodine-131 8.02 d 18 17 30 10

Caesium-137 30.0 a 1.0 1.0 1.0 1.0

Barium-140 12.79 d 7 3 20 0.7

Cerium-144 284.8 d 3 2.3 6 0.07

Neptunium-239 2.355 d 25 7 140 0.6

Plutonium-239 24 400.a 0.0015 0.0015 — —

3.2.2. Migration of radionuclides in the urban

environment

Due to natural weathering processes such as

the effects of rainfall and snow melting, and to

human activities such as traffic movement and

street washing and cleaning, radionuclides became

detached from the surfaces on which they were

deposited and were transported within settlements.

Contaminated leaves and needles from trees and

bushes are removed from settlements after seasonal

defoliation and radionuclides deposited on asphalt

and concrete pavements are eroded or washed off

and removed via sewage systems. These natural

processes and human activities significantly reduced

dose rates in inhabited and recreational areas

during 1986 and in successive years [3.21].

In general, vertical surfaces of houses are not

subjected to the same degree of weathering through

rain as horizontal surfaces such as roofs. The loss of

contamination from walls after 14 years has been

typically 50–70% of the initial deposit. Roof

contamination levels in Denmark decreased after

14 years by 60–95% of those originally present, due

to natural processes (see Fig. 3.11) [3.22].

In contrast, the level of radiocaesium on

asphalt surfaces has decreased substantially, such

that, generally, less than 10% of the initially

retained radiocaesium is now left. Only a small

fraction of the radiocaesium contamination is

associated with the bitumen fraction of the asphalt;

most is associated with a thin layer of street dust,

which will eventually be weathered off.

Measurements made in 1993 in the city of

Pripyat near the Chernobyl nuclear power plant

showed high residual levels of radiocaesium on the

roads. However, this town was evacuated in the

early accident phase, and therefore traffic there has

been limited. Some 5–10% of the initially deposited

radiocaesium seems to be firmly fixed to concrete

paved surfaces, and no significant decrease has been

recorded over the past few years. The weathering on

horizontal hard surfaces was, as expected, generally

faster in the areas with more traffic.

One of the consequences of these processes

has been secondary contamination of sewage

systems and sludge storage areas, which has necessi-

tated special cleanup measures. Generally, radionu-

clides in soil have not been transferred to other

urban areas but have migrated down into the soil

column due to natural processes or due to mixing

during the digging of gardens, kitchen gardens and

parks.

1986

2000

Undisturbed

soil

Trees, bushes Roofs Walls Streets,

pavements

Distribution of caesium-137 (relative units)

(a)

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1986 2000

Undisturbed

soil

Trees, bushes Roofs Walls Streets,

pavements

Distribution of caesium-137 (relative units)

(b)

FIG. 3.10. Typical distribution of

137

Cs on differen

urfaces within settlements in 1986 and 14 years after depo-

ition of the Chernobyl fallout. (a) Dry deposition; (b) we

deposition [3.23].

Time after deposition (years)

0.0

0.2

0.4

0.6

0.8

1.0

0 2 4 6 8 10 12 14

Corrugated Eternit – 45°

Red clay tile – 45°

Silicon treated Eternit – 30°

roof

(t) /

soil

(0)

FIG. 3.11. Measured

137

Cs contamination levels (relative to

the initial soil contamination) on three types of roof a

Risø, Denmark [3.22].

3.2.3. Dynamics of the exposure rate in urban

environments

Gamma radiation from radionuclides

deposited in the urban environment has contributed

to human external exposure. Compared with the

dose rate in open fields (see Section 5.2.2), the dose

rate within a settlement is significantly lower,

because of photon absorption in building structures,

especially those made of brick and concrete. The

lowest dose rates have been observed inside

buildings, and especially on the upper floors of

multistorey buildings. Due to radioactive decay of

the initial radionuclide mixture, wash-off from solid

surfaces and soil migration, dose rates in air have

been gradually decreasing with time in typical urban

areas.

Another relevant parameter is the time

dependence of the ratio of the dose rate in air at an

urban location to that in an open field (the ‘location

factor’) due to radionuclide migration processes.

The dependence of urban location factors on time

after the Chernobyl accident, as derived from

measurements performed in the town of

Novozybkov in the Russian Federation, is shown in

Fig. 3.12 [3.24]. While for virgin sites such as parks

or grassy plots the location factors are relatively

constant, values for hard surfaces such as asphalt

decrease considerably with time. Similar time

dependences have been found in other countries

[3.25, 3.26].

At present, in most of the settlements

subjected to radioactive contamination after the

Chernobyl accident, the air dose rate above solid

surfaces has returned to the pre-accident

background level. Some elevated air dose rates can

be measured, mainly in areas of undisturbed soil.

The highest level of urban radioactive contami-

nation is found in Pripyat, which is 3 km from the

Chernobyl nuclear power plant; its inhabitants were

resettled to non-contaminated areas within 1.5 days

of the accident.

3.3. AGRICULTURAL ENVIRONMENT

3.3.1. Radionuclide transfer in the terrestrial

environment

Radioactive elements behave differently in

the environment; some, such as caesium, iodine and

strontium, are environmentally mobile and transfer

readily, under certain environmental conditions, to

foodstuffs. In contrast, radionuclides with low

solubility, such as the actinides, are relatively

immobile and largely remain in the soil. The main

routes for the cycling of radionuclides and the

possible pathways to humans are shown in Fig. 3.13.

Many factors influence the extent to which

radionuclides are transferred through terrestrial

pathways. If transfer is high in a particular

environment it is said to be radioecologically

sensitive, because such transfer can lead to

relatively high radiological exposure [3.28].

0.0

0.2

0.4

0.6

0.8

1.0

0 1 2 3 4 5 6 7 8 9 10

Time after the accident (years)

Location factor

Virgin land (inside town)

Dirt surfaces

Asphalt

FIG. 3.12. Ratio of the radiation dose rate above differen

urfaces to that in open fields in the town of Novozybkov,

Russian Federation, after the Chernobyl accident [3.24].

Groundwater and rocks of geological environment

Organic–mineral soil component

Soil solution

Surface washout

Sorbed radionuclides

Soil–vegetation litter

Initial radionuclide fallout:

- Condensation component

- Fuel particles

Dissolution of

fuel particles

Sorption–

desorption

Migration

FIG. 3.13. Main transfer pathways of radionuclides in the

terrestrial environment [3.27].

Of the radionuclides deposited after the

Chernobyl accident, during the short initial phase

(zero to two months) those of iodine were the most

important with regard to human exposure via

agricultural food chains. In the longer term, radio-

caesium has been the most important (and, to a

much lesser extent, radiostrontium).

Radioecological sensitivity to radiocaesium is

generally higher in seminatural ecosystems than in

agricultural ecosystems, sometimes by a few orders

of magnitude [3.29]. This difference is caused by a

number of factors, the more important being in

some natural ecosystems the different physico-

chemical behaviour in soils (lack of competition

between caesium and potassium, resulting in higher

transfer rates of radiocaesium in nutrient-poor

ecosystems) and the presence of specific food chain

pathways, leading to highly contaminated produce

from seminatural ecosystems. Also, forest soils are

fundamentally different from agricultural soils; they

have a clear multilayered vertical structure charac-

terized mainly by a clay-poor mineral layer, which

supports a layer rich in organic matter. In contrast,

agricultural soils generally contain less organic

matter and higher amounts of clay.

3.3.2. Food production systems affected by the

accident

The radioactive material released by the

Chernobyl accident contaminated large areas of the

terrestrial environment and had a major impact on

both agricultural and natural ecosystems not only

within the former USSR but also in many other

countries in Europe.

In the former USSR countries, the food

production system that existed at the time of the

accident can be divided into two types: large

collective farms and small private farms. Collective

farms routinely apply land rotation combined with

ploughing and fertilization to improve productivity.

Traditional small private farms, in contrast, seldom

apply artificial fertilizers and often use manure for

improving yield. Typically they have one or, at most,

a few cows, and produce milk mainly for private

consumption. The grazing regime of private farms

was initially limited to the utilization of marginal

land not used by the collective farms, but nowadays

includes some better quality pasture.

In western Europe, poor soils are used

extensively for agriculture, mainly for grazing of

ruminants (e.g. sheep, goats, reindeer and cattle).

Such areas include alpine meadows and upland

regions in western and northern Europe that have

organic soils.

3.3.3. Effects on agriculture in the early phase

At the time of the Chernobyl nuclear power

plant accident, vegetation in the affected areas was

at different stages of growth, depending on latitude

and elevation. Initially, interception on plant leaves

of dry deposition and atmospheric washout with

precipitation were the main mechanisms by which

vegetation became contaminated. In the medium

and long term, root uptake predominated. The

highest activity concentrations of radionuclides in

most foodstuffs occurred in 1986.

In the initial phase,

131

I was the radionuclide of

most concern and milk was the main contributor to

internal dose. This is because radioiodine was

released in large amounts and intercepted by plant

surfaces that were then grazed by dairy cows. The

ingested radioiodine was completely absorbed in

the gut of the cow [3.31] and then rapidly

transferred to the animal’s thyroid and milk (within

about one day). Thus peak values occurred rapidly

after deposition in late April or early May 1986,

depending on when deposition occurred in different

countries. During this period, in the former USSR

and some other European countries,

131

I activity

concentrations in milk exceeded the national and

regional (European Union (EU)) action levels of a

few hundred to a few thousand becquerels per litre

(see Section 4.1).

There are no time trend data available for

131

activity concentrations in milk in the first few days

after the accident in the heavily affected areas of the

USSR, for the obvious and understandable reason

that the authorities were dealing with other

immediate accident response priorities. Never-

theless, data are available for the period starting two

weeks after the accident from the Tula region of the

Russian Federation, and the data in Fig. 3.14(a)

show an exponential decline in

131

I activity concen-

tration in milk normalized to

137

Cs deposition,

which can be extrapolated back to the first days to

estimate the initial

131

I activity concentration in

milk. Furthermore, a direct comparison of

131

activity in milk in early May with

137

Cs deposition

shows the contribution of dry deposition to

131

I in

milk, because the linear relationship line shown

does not go through the zero deposition point

(Fig. 3.14(b)). In the early spring in northern

Europe, dairy cows and goats were not yet on

pasture, therefore there was very little milk

contamination. In contrast, in the southern regions

of the USSR, as well as in Germany, France and

southern Europe, dairy animals were already

grazing outdoors and some contamination of cow,

goat and sheep milk occurred. The

131

I activity

concentration in milk decreased with an effective

half-life of four to five days [3.32], due to its short

physical half-life and the reduction in iodine activity

concentrations on plants due to atmospheric

removal processes from leaf surfaces (Fig. 3.15).

This removal occurred with a mean weathering half-

life on grass of nine days for radioiodine and 11 days

for radiocaesium [3.33]. Leafy vegetables were also

contaminated on their surfaces and also made a

contribution to the radiation dose to humans via the

food chain (Fig. 3.15).

Both plants and animals were also contami-

nated with radiocaesium and, to a lesser extent,

radiostrontium. From June 1986 radiocaesium was

the dominant radionuclide in most environmental

samples (except in the CEZ) and in food products.

As shown in Fig. 3.16, the contamination of milk

with radiocaesium decreased during spring 1986

with an effective half-life of about two weeks, due to

(b)

Caesium-137 in soil (kBq/m

)

0 50 100 150 200 250 300 350 400 450

Iodine-131 in milk on 8 May 1986 (kBq/kg)

0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30 32

1000

100

Iodine-131 in milk normalized to

137

Cs soil deposition

(Bq

131

l/kg)/(kBq

137

Cs/m

)

(a)

Days after fallout

FIG. 3.14. Variation in

131

I activity concentration in milk

in the Tula region of the Russian Federation (a) with time

after deposition, standardized by

137

Cs soil deposition, and

(b) with

137

Cs soil deposition (time corrected to

8 May 1986) [3.30].

10 000

1000

100

ASTRAL assessment

60 000 Bq/m

20 000 Bq/m

Measurement results

Alsace

Ht Rhin

Moselle

Meurthe et M

Bas Rhin

Ardennes/Meuse

1 May 11 May 21 May 31 May 10 June 20 June 30 June

Iodine-131 activity in leafy vegetables (Bq/kg fresh)

(a)

Iodine-131 activity in milk (Bq/L)

Ardennes

Ht Marne/Aube

Meurthe et M

Meuse

Moselle

Ht Rhin/Ter. Bel

Vosges

10 000 Bq/m

40 000 Bq/m

ASTRAL assessment:

1000

100

28 Apr. 8 May 18 May 28 May 8 Jun. 18 Jun.

Measurement results:

(b)

FIG. 3.15. Changes with time of

131

I activity concentrations

in (a) leafy vegetables and (b) cow’s milk in differen

regions of France, May–June 1986 [3.34].

Measurements from eastern France

Measurements from central France

Measurements from western France

ASTRAL assessments:

5000 Bq/m

2000 Bq/m

400 Bq/m

May 1986 Jul. Sep. Nov. Jan. 1987 Mar. May Jul. 1987

100

Caesium-137 activity in milk (Bq/L)

FIG. 3.16. Changes with time of

137

Cs activity concentra-

tions in cow’s milk in France in 1986–1987 as observed and

imulated by the ASTRAL model [3.34].

weathering, biomass growth and other natural

processes. However, radiocaesium activity concen-

trations increased again during winter 1986/1987,

due to the feeding of cows with contaminated hay

harvested in spring/summer 1986. This phenomenon

was observed in the winter period in many countries

after the accident.

The transfer to milk of many of the other

radionuclides present in the terrestrial environment

during the early phase of the accident was low. This

was because of low inherent transfer in the gut of

those elements, compounded by low bioavailability

due to their association within the matrix of fuel

particles [3.35]. Nevertheless, some high transfers

occurred, notably that of

110m

Ag to the liver of

ruminants [3.36].

3.3.4. Effects on agriculture in the long term

Since 1987, the radionuclide content of both

plants and animals has been largely determined by

the interaction between radionuclides and different

soil components, as soil is the main reservoir of long

lived radionuclides deposited on terrestrial

ecosystems. This process controls radionuclide

availability [3.37, 3.38] for uptake into plants and

animals and also influences radionuclide migration

down the soil column.

3.3.4.1. Physicochemistry of radionuclides in the

soil–plant system

Plants take up nutrients and pollutants from

the soil solution. The activity concentration of

radionuclides in the soil solution is the result of

physicochemical interactions with the soil matrix, of

which competitive ion exchange is the dominant

mechanism. The concentration and composition of

the major and competitive elements present in the

soil are thus of prime importance for determining

the radionuclide distribution between the soil and

the soil solution. Many data obtained after the

Chernobyl accident demonstrate that the amount

and nature of clay minerals present in soil are key

factors in determining radioecological sensitivity

with regard to radiocaesium. These features are

crucially important for understanding radiocaesium

behaviour, especially in areas distant from the

Chernobyl nuclear power plant, where

137

Cs was

initially deposited mainly in condensed, water

soluble forms.

Close to the nuclear power plant, radio-

nuclides were deposited in a matrix of fuel particles

that have slowly dissolved with time; this process is

not complete today. The more significant factors

influencing the fuel particle dissolution rate in soil

are the acidity of the soil solution and the physico-

chemical properties of the particles (notably the

degree of oxidization) (see Fig. 3.17). In a low pH of

pH4, the time taken for 50% dissolution of particles

was about one year, whereas for a higher pH of pH7

up to 14 years were needed [3.39–3.41]. Thus in acid

soils most of the fuel particles have already

dissolved. In neutral soils, the amount of mobile

released from the fuel particles is now increasing,

and this will continue over the next 10–20 years.

In addition to soil minerals, microorganisms

can significantly influence the fate of radionuclides

in soils [3.42, 3.43]. They can interact with minerals

and organic matter and consequently affect the

bioavailability of radionuclides. In the specific case

of mycorrhizal fungi, soil microorganisms may even

act as a carrier, transporting radionuclides from the

soil solution to the associated plant.

Oxidation state +5 ± 0.5

Oxidation state +4 ± 0.5

(a)

3 3.5 4 4.5 5 5.5 6 6.5 7 7.5 8

South North West

FP (% activity in fuel particles)

(b)

FIG. 3.17. (a) Variation in the oxidation within a

Chernobyl fuel particle [3.40]; (b) fraction of

Sr presen

in fuel particles (FP) ten years after the Chernobyl acciden

as a function of soil acidity [3.39].

A traditional approach of characterizing the

mobility and bioavailability of a radioactive

contaminant in soil is by applying sequential

extraction techniques. A number of experimental

protocols have been developed that use a sequence

of progressively aggressive chemicals, each of which

is assumed to selectively leach a fraction of the

contaminant bound to a specific soil constituent. An

example of the results available from this procedure

is presented in Fig. 3.18, which shows that a much

higher proportion of radiocaesium was fixed in the

soil than of radiostrontium. The selectivity and

reproducibility of chemical extraction procedures

varies and therefore often should be considered to

give only qualitative estimates of bioavailability.

By use of sequential extraction techniques, the

fraction of exchangeable

137

Cs was found to

decrease by a factor of three to five within a decade

after 1986 [3.44, 3.45]. This time trend, which

resulted in a reduction of plant contamination, may

be due to progressive fixation of radiocaesium in

interlayer positions of clay minerals and to its slow

diffusion and binding to the frayed edge sites of clay

minerals. This process reduces the exchangeability

of radiocaesium so that is not then available to enter

the soil solution from which plants take up most of

the radiocaesium via the roots. For

Sr an increase

with time of the exchangeable fraction has been

observed, which is attributed to the leaching of the

fuel particles [3.39].

3.3.4.2. Migration of radionuclides in soil

The vertical migration of radionuclides down

the soil column can be caused by various transport

mechanisms, including convection, dispersion,

diffusion and biological mixing. Root uptake of

radionuclides into plants is correlated with vertical

migration. Typically, the rate of movement of radio-

nuclides varies with soil type and physicochemical

form. As an example, Fig. 3.19 shows the change

with time of the depth distributions of

Sr and

137

measured in the Gomel region of Belarus. Although

there has been a significant downward migration of

both radionuclides, much of the radionuclide

activity has remained within the rooting zone of

plants. At such sites, where contamination occurred

through atmospheric deposition, there is a low risk

of radionuclide migration to groundwater.

The rate of downward migration in different

types of soil varies for radiocaesium and radio-

strontium. Low rates of

Sr vertical migration are

observed in peat soils, whereas

137

Cs migrates at the

highest rate in these (highly organic) soils, but

moves much more slowly in soddy podzolic sandy

soils. In dry meadows, the migration of

137

Cs below

the root-containing zone (0–10 cm) was hardly

detectable in the ten years after the fallout

deposition. Thus the contribution of vertical

migration to the decrease of

137

Cs activity concen-

trations in the root-containing zone of mineral soils

86%

12%

37%

51%

Fixed Extractable Exchangeable

Caesium-137 Strontium-90

FIG. 3.18. Forms of radionuclides in soddy podzolic loam

and soil of the Gomel region of Belarus in 1998 [3.46].

0 20 40 60 80 100

0–5

6–10

11–15

16–20

21–25

26–30

31–35

1987

2000

0 20 40 60 80 100

0–5

6–10

11–15

16–20

21–25

26–30

31–35

Strontium-90 activity (%)

1987

2000

Depth (cm)

Caesium-137 activity (%)

FIG. 3.19. Depth distributions of

137

Cs and

Sr measured in 1987 and 2000 in a soddy gley sandy soil (in per cent of tota

activity) in the Gomel region of Belarus [3.46].

is negligible. In contrast, in wet meadows and in

peatland, downward migration can be an important

factor in reducing the availability of

137

Cs for plants

[3.48].

The higher rates of

Sr vertical migration are

observed in low humified sandy soil (Fig. 3.20),

soddy podzolic sandy soil and sandy loam soil with

an organic content of less than 1% [3.27]. Generally,

the highest rate of

Sr vertical migration occurs

where there are completely non-equilibrium soil

conditions. This occurs in the floodplains of rivers,

where the soil is not structurally formed (light

humified sands), in arable lands in a non-

equilibrium state and in soils in which the organic

layers have been removed, for example at sites of

forest fires and sites with deposited sand with a low

content of organic matter (<1%). In such conditions

there is a high rate of radiostrontium vertical

migration to groundwater with convective moisture

flow, and high activity levels can occur in localized

soil zones. Thus the spatial distribution of

Sr can

be particularly heterogeneous in soils in which there

have been changes in sorption properties.

Agricultural practices have a major impact on

radionuclide behaviour. Depending on the type of

soil tillage and on the tools used, a mechanical redis-

tribution of radionuclides in the soil may occur. In

arable soils, radionuclides are distributed fairly

uniformly down the whole depth of the tilled layer.

The lateral redistribution of radionuclides in

catchments, which can be caused by both water and

wind erosion, is significantly less than their vertical

migration into the soil and the underlying geological

layers [3.27]. The type and density of plant cover

may significantly affect erosion rates. Depending on

the intensity of erosive processes, the content of

radionuclides in the arable layer on flat land with

small slopes may vary by up to 75% [3.49].

3.3.4.3. Radionuclide transfer from soil to crops

The uptake of radionuclides, as well as of

other trace elements, by plant roots is a competitive

process [3.50]. For radiocaesium and radiostrontium

the main competing elements are potassium and

calcium, respectively. The major processes

influencing radionuclide transport processes within

the rooting zone are schematically represented in

Fig. 3.21, although the relative importance of each

component varies with the radionuclide and soil

type.

The fraction of deposited radionuclides taken

up by plant roots differs by orders of magnitude,

depending primarily on soil type. For radiocaesium

and radiostrontium, the radioecological sensitivity

of soils can be broadly divided into the categories

listed in Table 3.5. For all soils and plant species, the

root uptake of plutonium is negligible compared

with the direct contamination of leaves via rain

splash or resuspension.

Transfer from soil to plants is commonly

quantified using either the transfer factor (TF,

020406080100

0–2

2–4

4–6

6–8

8–10

10–14

14–18

18–22

22–26

26–30

30–34

34–38

38–42

42–46

46–50

50–54

54–60

60–65

65–100

Part of radionuclide activity (%)

Americium-241

Europium-154

Strontium-90

Caesium-137

Depth (cm)

FIG. 3.20. Depth distributions of radionuclides in low

humified sandy soil (in per cent of total activity) measured

in 1996 [3.47].

Soil organic matrix

PLANTS

Soil Solution

[RN]

ROOTS,

Mycorhiza

[RN]

min

[RN]

mic

Soil mineral matrix

Additional

non-radioactive p

ollutants

[RN]

org

Plants

Soil solution

Roots,

mycorrhiza

Soil microorganisms

Soil mineral matrix

Fertilizers

Soil organic matrix

FIG. 3.21. Radionuclide pathways from soil to plants with

consideration of biotic and abiotic processes [3.43].

dimensionless, equal to plant activity concentration,

Bq/kg, divided by soil activity concentration, Bq/kg)

or the aggregated transfer coefficient (T

, m

/kg,

equal to plant activity concentration, Bq/kg, divided

by activity deposition on soil, Bq/m

The highest

137

Cs uptake by roots from soil to

plants occurs in peaty, boggy soils, and is one to two

orders of magnitude higher than in sandy soils; this

uptake often exceeds that of plants grown on fertile

agricultural soils by more than three orders of

magnitude. The high radiocaesium uptake from

peaty soils became important after the Chernobyl

accident because in many European countries such

soils are vegetated by natural unmanaged grassland

used for the grazing of ruminants and the

production of hay.

The amount of radiocaesium in agricultural

products in the medium to long term depends not

only on the density of contamination but also on the

soil type, moisture regime, texture, agrochemical

properties and plant species. Agricultural activity

often reduces the transfer of radionuclides from soil

to plants by physical dilution (e.g. ploughing) or by

adding competitive elements (e.g. fertilizing). There

are also differences in radionuclide uptake between

plant species. Although among species variations in

uptake may exceed one or more orders of

magnitude for radiocaesium, the impact of differing

radioecological sensitivities of soils is often more

important in explaining the spatial variation in

transfer in agricultural systems.

The influence of other factors that have been

reported to influence plant root uptake of radionu-

clides (e.g. soil moisture) is less clear or may be

explained by the basic mechanisms discussed above;

for example, the accumulation of radiocaesium in

crops and pastures is related to soil texture. In sandy

soils the uptake of radiocaesium by plants is approx-

imately twice as high as in loam soils, but this effect

is mainly due to the lower concentrations of its main

competing element, potassium, in sand.

The main process controlling the root uptake

of radiocaesium into plants is the interaction

between the soil matrix and solution, which

depends primarily on the cation exchange capacity

of the soil. For mineral soils this is influenced by the

concentrations and types of clay minerals and the

concentrations of competitive major cations,

especially potassium and ammonium. Examples of

these relationships are shown in Fig. 3.22 for both

radiocaesium and radiostrontium. The modelling of

soil solution physicochemistry, which takes account

of these major factors, enables prediction of the root

uptake of both radionuclides [3.51, 3.52].

TABLE 3.5. CLASSIFICATION OF RADIOECOLOGICAL SENSITIVITY FOR SOIL–PLANT

TRANSFER OF RADIOCAESIUM AND RADIOSTRONTIUM

Sensitivity Characteristic Mechanism Example

Radiocaesium

High Low nutrient content

Absence of clay minerals

High organic content

Little competition with

potassium and ammonium in

root uptake

Peat soils

Medium Poor nutrient status, consisting of

minerals, including some clays

Limited competition with

potassium and ammonium in

root uptake

Podzol, other sandy soils

Low High nutrient status

Considerable fraction of clay

minerals

Radiocaesium strongly held to

soil matrix (clay minerals),

strong competition with

potassium and ammonium in

root uptake

Chernozem, clay and

loam soils (used for

intensive agriculture)

Radiostrontium

High Low nutrient status

Low organic matter content

Limited competition with

calcium in root uptake

Podzol sandy soils

Low High nutrient status

Medium to high organic matter

content

Strong competition with calcium

in root uptake

Umbric gley soils, peaty

soils

Thus differences in radioecological sensitiv-

ities of soils explain why in some areas of low

deposition high concentrations of radiocaesium are

found in plants and mushrooms harvested from

seminatural ecosystems and, conversely, why areas

of high deposition can show only low to moderate

concentrations of radiocaesium in plants. This is

illustrated in Fig. 3.23, in which the variability in

activity concentrations of radiocaesium and radio-

strontium in plants is shown for a normalized

concentration in soil.

3.3.4.4. Dynamics of radionuclide transfer to crops

In 1986 the

137

Cs content in plants, which was

at its maximum in that year, was primarily

determined by aerial contamination. During the

first post-accident year (1987), the

137

Cs content in

plants dropped by a factor of three to one hundred

(depending on soil type) as roots became the

dominant contamination route.

For meadow plants in the first years after

deposition,

137

Cs behaviour was considerably

influenced by the radionuclide distribution between

soil and mat. In this period,

137

Cs uptake from mat

significantly exceeded (up to eight times) that from

soil. Further, as a result of mat decomposition and

radionuclide transfer to soil, the contribution of mat

decreased rapidly, and in the fifth year after the

deposition it did not exceed 6% for automorphous

soils and 11% for hydromorphous soils [3.41].

In most soils the transfer rate of

137

Cs to plants

has continued to decrease since 1987, although the

rate of decrease has slowed, as can be seen from

Fig. 3.24 [3.55]. A decrease with time similar to that

shown in Fig. 3.24 has been observed in many

(a)

0.05

0.1

0.15

0.2

0.25

0.3

0.35

Less than 80 81–140 141–200 201–300 More 300

Potassium content (mg/kg)

Sands

Sandy loam

Clay loam

(

137

Cs) (10

–3

/kg)

(b)

0 2 4 6 8 10 12

Calcium content (mg-equ/100 g)

Tag (

Sr) (10

–3

/kg)

ag = 3.9Ca

–1

FIG. 3.22. (a) Transfer of

137

Cs into oat grain in soddy

odzolic soils of various textures with varying potassium

contents [3.61] and (b) transfer of

Sr into seeds of winte

rye with varying concentrations of exchangeable calcium in

different soils [3.53].

0.10

1.00

10.00

100.00

Strontium-90

Caesium-137

gk/qB

Wheat seedsNatural grass (hay)

FIG. 3.23. Variation in the concentrations of

137

Cs and

in two plant species with soil type; the data refer to a soi

deposition of 1 kBq/m

[3.54]. 1: peat soil; 2: soddy

odzolic soil; 3: chernozem soil.

1986 1990 1994 1998 20021988 1992 1996 2000 20

100

1000

Grain

Potato

Year

gk/qB

FIG. 3.24. Changes with time of

137

Cs concentrations in

rain and potato produced in contaminated districts of the

Bryansk region of the Russian Federation (Bq/kg) [3.55].

studies of plant root uptake in different crops, as can

be seen in Figs 3.25 and 3.26 for cereals and natural

grasses, respectively, growing in two different soil

types [3.56]. Two experimental points for

chernozem soil (18 and 20 years) were obtained

from the measurements made in 1980–1985 (i.e.

after

137

Cs global fallout and before the Chernobyl

accident) (Fig. 3.25). Values of

137

Cs TFs for cereals

as well as for potato and cow’s milk obtained about

20 years following global fallout do not differ signif-

icantly from those observed eight to nine years and

later after the Chernobyl fallout in remote areas

with dominant sandy, sandy loam and chernozem

soils [3.56, 3.57]. The difference between T

values

relevant to cereals grown on fertilized soil is much

lower than the difference for natural grasses.

For the transfer of radiocaesium from soil to

plants, a decrease with time is likely to reflect: (a)

physical radionuclide decay; (b) the downward

migration of the radionuclide out of the rooting

zone; and (c) physicochemical interactions with the

soil matrix that result in decreasing bioavailability.

In many soils, the ecological half-lives of the plant

root uptake of radiocaesium can be characterized

by two components: (a) a relatively fast decrease,

with a half-life between 0.7 and 1.8 years,

dominating for the first four to six years, leading to a

reduction of concentration in plants by about an

order of magnitude compared with 1987; and (b) a

slower decrease with a half-life of between seven

and 60 years [3.45, 3.55, 3.57, 3.58]. The dynamics of

the decrease of

137

Cs availability in the soil–plant

system are considerably influenced by soil

properties, and as a result the rates of decreasing

137

Cs uptake by plants can differ by a factor of three

to five [3.41].

Some caution should be exercised, however, in

generalizing these observations, because some data

indicate almost no decrease in the root uptake of

radiocaesium with time beyond the first four to six

years, which suggests that there is no reduction in

bioavailability in soil within the time period of

observation. Furthermore, the prediction of

ecological half-lives that exceed the period of

observation can be highly uncertain. The

application of countermeasures aimed at reducing

the concentration of radiocaesium in plants will also

modify the ecological half-life.

Compared with radiocaesium, the uptake of

Sr by plants has usually not shown such a marked

decrease with time. In the areas close to the

Chernobyl nuclear power plant, the gradual

dissolution of fuel particles has enhanced the

bioavailability of

Sr, and therefore there has been

an increase with time in

Sr uptake by plants

(Fig. 3.27 [3.39]).

In remote areas, where strontium radio-

nuclides were predominantly deposited in

condensed form and in lesser amounts as fine

dispersed fuel particles, the dynamics of long term

transfer of

Sr to plants were similar to those of

radiocaesium, but with different ecological half-

lives for plant root uptake. This difference is

associated with various mechanisms of soil transfer

for these two elements. The fixation of strontium by

soil components depends less on the clay content of

the soil than that of caesium (see Table 3.5). More

generally, the values of

Sr transfer parameters

from soil to plants depend less on the soil properties

than the transfer parameters for radiocaesium

[3.37]. An example of the time dependence of

uptake by plants is given in Fig. 3.28 [3.56].

1.000

0.100

0.010

0.001

0 2 4 6 8 10 12 14 16 18 20

Years after fallout

(1)

(2)

ag2

= 0.53 exp(–ln 2t/0.9) + 0.018

ag1

= 1.2 exp(–ln 2t/0.8) + 0.035

(Cs) (10

–3

/kg)

FIG. 3.25. Dynamics of the

137

Cs T

for cereals. 1: sandy

and sandy loam soil, Bryansk region, Russian Federation;

: chernozem soil, Tula and Orel regions, Russian Federa-

tion [3.56].

100.00

10.00

1.00

0.10

0.01

0 2 4 6 8 10 12 14 16 18

Years after fallout

(1)

(2)

ag2

= 6.0 exp(–ln 2t/0.9) + 0.10

ag1

= 200 exp(–ln 2t/1.0) + 2.2

(Cs) (10

–3

/kg)

FIG. 3.26. Dynamics of the

137

Cs T

(dry weight) fo

natural grasses. 1: sandy and sandy loam soil, Bryansk

region, Russian Federation; 2: chernozem soil, Tula and

Orel regions, Russian Federation [3.56].

3.3.4.5. Radionuclide transfer to animals

Animals take up radionuclides through

contaminated forage and direct soil ingestion. Milk

and meat were major contributors to the internal

radiation dose to humans after the Chernobyl

accident, both in the short term, due to

131

I, and in

the long term, due to radiocaesium. In intensively

managed agricultural ecosystems, high levels of

contamination of animal food products can be

expected only for a few weeks, or at most a few

months, after a pulse of fallout. In these circum-

stances the extent of interception and retention on

plant surfaces largely determines both the duration

and the level of contamination of animal derived

food products. An exception is found where very

high deposition occurs or where plant uptake is high

and sustained, both of which occurred in some areas

after the Chernobyl accident.

The levels of radiocaesium in animal food

products can be high and persist for a long time,

even though the original deposition may not have

been very high. This is because: (a) soils often allow

significant uptake of radiocaesium; (b) some plant

species accumulate relatively high levels of radio-

caesium, for example ericaceous species and fungi;

and (c) areas with poor soils are often grazed by

small ruminants, which accumulate higher caesium

activity concentrations than larger ruminants

[3.35].

The contamination of animal products by

radionuclides depends on their behaviour in the

plant–soil system, the absorption rate and metabolic

pathways in the animal and the rate of loss from the

animal (principally in urine, faeces and milk).

Although absorption can occur through the skin

and lungs, oral ingestion of radionuclides in feed,

and subsequent absorption through the gut, is the

major route of uptake of most radionuclides.

Absorption of most nutrients takes place in the

rumen or the small intestine at rates that vary from

almost negligible, in the case of actinides, to 100%

for radioiodine, and varying from 60% to 100% for

radiocaesium, depending on the form [3.31].

After absorption, radionuclides circulate in

the blood. Some accumulate in specific organs; for

example, radioiodine accumulates in the thyroid,

and many metal ions, including

144

Ce,

106

Ru and

110m

Ag, accumulate in the liver. Actinides and

especially radiostrontium tend to be deposited in

the bone, whereas radiocaesium is distributed

throughout the soft tissues [3.36, 3.37, 3.50, 3.59,

3.60].

The transfer of radionuclides to animal

products is often described by transfer coefficients

defined as the equilibrium ratio between the radio-

nuclide activity concentration in milk, meat or eggs

divided by the daily dietary radionuclide intake.

Transfer coefficients for radioiodine and radio-

caesium to milk, and for radiocaesium to meat, are

generally lower for large animals such as cattle than

for small animals such as sheep, goats and chickens.

The transfer of radiocaesium to meat is higher than

that to milk.

The long term time trend of radiocaesium

contamination levels in meat and milk, an example

of which is displayed in Fig. 3.29, follows that for

1986 1988 1990 1992 1994 1996 1998 2000 2002

Year

TF ((Bq/kg)/(kBq/m

))

FIG. 3.27. Dynamics of the

Sr TF into natural grass from

oddy podzolic soil in the CEZ [3.39].

100.00

10.00

1.00

0.10

0.01

1 3 5 7 9 11 13 15

Years after fallout

(1)

(2)

ag2

= 0.30 exp(–ln 2t/4.0) + 0.11

ag1

= 18 exp(–ln 2t/4.1) + 3.8

(Sr) (10

–3

/kg)

ag3

= 0.12 exp(–ln 2t/3.3) + 0.034

(3)

FIG. 3.28. Dynamics of the

Sr aggregated TF for natura

rasses (1: sandy and sandy loam soil, Bryansk region,

Russian Federation) and cow’s milk (2: sandy and sandy

loam soil, Bryansk region, Russian Federation; 3:

chernozem soil, Tula and Orel regions, Russian Federa-

tion) [3.56].

vegetation and can be divided into two phases [3.55,

3.57, 3.58]. For the first four to six years after the

deposition of the radiocaesium there was an initial

fast decrease with an ecological half-life of between

0.8 and 1.2 years. For later times, only a small

decrease has been observed [3.55, 3.56].

There are differing rates of

137

Cs transfer to

milk in areas with different soil types, as demon-

strated over nearly two decades after the accident

(Fig. 3.30) in milk from the Bryansk, Tula and Orel

regions of the Russian Federation, where few

countermeasures have been used. The transfer of

137

Cs to milk is illustrated using the T

, which

normalizes the data for different levels of soil

contamination; this makes comparison among soil

types easier. The transfer to milk declines in the

order peat bog > sandy and sandy loam >

chernozem and grey forest soils. Both the dynamics

137

Cs activity concentration in milk and its

dependence on soil type are similar to those in

natural grasses (see Fig. 3.26) sampled in areas

where cattle graze.

Similar long term data are available for

comparing the transfer of

137

Cs to beef in the

Russian Federation for different soil types. They

also show higher transfer in areas with sandy/sandy

loam soils compared with chernozem soils

(Fig. 3.31); there has been little decline in

137

transfer over the past decade.

The long term dynamics of

Sr in cow’s milk

sampled in Russian areas with dominant soddy

podzolic and chernozem soils (see Fig. 3.28) are

different from those of

137

Cs. The graphs for

Sr in

milk do not contain the initial decreasing portion

with an ecological half-life of about one year, as

shown in the graphs for

137

Cs, which are presumed

to reflect fixation of caesium in the soil matrix. In

contrast, the

Sr activity concentration in cow’s

milk gradually decreases with an ecological half-life

of three to four years; the second component (if

any) has not yet been identified. The physical and

chemical processes responsible for these time

dynamics obviously include diffusion and

convection with vertical transfer of

Sr into soil, as

well as its radioactive decay. However, the chemical

interactions with the soil components may differ

significantly from those known for caesium.

Meat

Milk

10 000

1000

100

1986 1988 1990 1992 1994 1996 1998 2000 2002 2004

Year

gk/qB

FIG. 3.29. Changes with time in mean

137

Cs activity

concentrations in meat and milk produced in contaminated

districts of the Bryansk region of the Russian Federation

(Bq/kg) [3.55].

10.00

1.00

0.10

0.01

0 2 4 6 8 10 12 14 16 18

r f

(1)

(2)

ag2

= 0.34 exp(–ln 2t/1.6) + 0.03

ag1

= 13 exp(–ln 2t/1.6) + 0.78

(Cs) (10

–

/kg)

(

)

10.0

1.0

0.1

0 2 4 6 8 10 12 14 16

Years after fallout

(1)

(2)

ag2

= 3 exp(–ln 2t/1.8) + 0.09

ag1

= 7 exp(–ln 2t/1.7) + 0.12

(Cs) (10

–3

/kg)

(b)

FIG. 3.30. (a) Dynamics of the

137

Cs aggregated TF for

cow’s milk. 1: peat bog soil, Bryansk region, Russian

Federation; 2: chernozem soil, Tula and Orel regions,

Russian Federation [3.56]. (b) Dynamics of

137

Cs aggre-

gated TF for cow’s milk (sandy and sandy loam soil,

Bryansk region, Russian Federation). 1:

137

Cs soil deposi-

tion <370 kBq/m

; 2:

137

Cs soil deposition >370 kBq/m

[3.56].

By combining information on radionuclide

transfer with spatially varying information in

geographic information systems, it is possible to

identify zones in which a specified average activity

concentration in milk is likely to be exceeded. An

example is shown in Fig. 3.32.

A significant amount of production in the

former USSR is confined to the grazing of privately

owned cows on poor, unimproved meadows. Owing

to the poor productivity of these areas, radiocaesium

uptake is relatively high compared with that on land

used by collective farms. As an example of the

difference between farming systems, changes in

137

activity concentrations in milk from private and

collective farms in the Rovno region of Ukraine are

shown in Fig. 3.33. The activity concentrations in

milk from private farms exceeded the action levels

until 1991, when countermeasures were

implemented that resulted in a radical improvement.

3.3.5. Current contamination of foodstuffs and

expected future trends

Table 3.6 shows summarized data of measured

current (2000–2003) activity concentrations of

radiocaesium in grain, potato, milk and meat

produced in highly and less highly contaminated

areas covering many different types of soil with

widely differing radioecological sensitivities in

Belarus, the Russian Federation and Ukraine.

Caesium-137 activity concentrations are consist-

ently higher in animal products than in plant

products.

Currently, due to natural processes and

agricultural countermeasures, radiocaesium activity

concentrations in agricultural food products

produced in areas affected by the Chernobyl fallout

are generally below national, regional (EU) and

international action levels [3.64, 3.65]. However, in

some limited areas with high radionuclide contami-

nation (parts of the Gomel and Mogilev regions in

Belarus and the Bryansk region in the Russian

Federation) or poor organic soils (the Zhytomyr

and Rovno regions in Ukraine), radiocaesium

activity concentrations in food products, especially

milk, still exceed the national action levels of about

100 Bq/kg. In these areas remediation may still be

warranted (see Section 4).

Contaminated milk from privately owned

cows with

137

Cs activity concentrations exceeding

100 Bq/L (the current permissible level for milk)

was being produced in more than 400, 200 and 100

Ukrainian, Belarusian and Russian settlements,

100.00

10.00

1.00

0.10

0.01

0 2 4 6 8 10 12 14

Years after fallout

(1)

(2)

ag2

= 0.77 exp(–ln 2t/0.9) + 0.12

ag1

= 17 exp(–ln 2t/0.9) + 1.0

(Cs) (10

–3

/kg)

FIG. 3.31. Dynamics of

137

Cs aggregated TF for beef. 1:

andy and sandy loam soil; 2: chernozem soil [3.56].

FIG. 3.32. Isolines of different levels of probability of

exceeding 100 Bq/L of

137

Cs in milk (for 1991) in the

Kaluga region, Russian Federation [3.54]. The red colour

indicates a high probability, the pink and green colours

indicate medium and low probabilities, respectively.

500

1000

1500

2000

1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000

Year

Bq/L

Private farms

Collective farms

TPL

FIG. 3.33. Typical dynamics of

137

Cs activity concentra-

tions in milk produced on private and collective farms in

the Rovno region of Ukraine with a comparison with the

temporary permissible level (TPL) [3.62].

respectively, 15 years after the accident. Levels of

milk contamination higher than 500 Bq/L occur in

six Ukrainian, five Belarusian and five Russian

settlements (in 2001).

The concentrations and transfer coefficients

shown in the above mentioned figures and tables

show that there has been only a slow decrease in

radiocaesium activity concentrations in most plant

and animal foodstuffs during the past decade. This

indicates that radionuclides must be close to

equilibrium within the agricultural ecosystems,

although continued reductions with time are

expected, due to continuing radionuclide migration

down the soil profile and to radioactive decay (even

if there was an equilibrium established between

137

Cs in the labile and non-labile pools of soil).

Given the slow current rates of decline, and the

difficulties in quantifying long term effective half-

lives from currently available data because of high

uncertainties, it is not possible to conclude that

there will be any further substantial decrease over

the next decades, except due to the radioactive

decay of both

137

Cs and

Sr, which have half-lives of

about 30 years.

Radionuclide activity concentrations in

foodstuffs can increase through fuel particle disso-

lution, changes in the water table as a consequence

of change of management of currently abandoned

land or cessation of the application of counter-

measures.

3.4. FOREST ENVIRONMENT

3.4.1. Radionuclides in European forests

Forest ecosystems were one of the major

seminatural ecosystems contaminated as a result of

fallout from the Chernobyl plume. The primary

concern from a radiological perspective is the long

term contamination of the forest environment and

its products with

137

Cs, owing to its 30 year half-life.

In the years immediately following contamination,

the shorter lived

134

Cs isotope was also significant.

In forests, other radionuclides such as

Sr and the

plutonium isotopes are of limited significance for

humans, except in relatively small areas in and

around the CEZ. As a result, most of the available

environmental data concern

137

Cs behaviour and

the associated radiation doses.

Forests provide economic, nutritional and

recreational resources in many countries.

TABLE 3.6. MEAN AND RANGE OF CURRENT CAESIUM-137 ACTIVITY CONCENTRATIONS IN

AGRICULTURAL PRODUCTS ACROSS CONTAMINATED AREAS OF BELARUS [3.49], THE

RUSSIAN FEDERATION [3.55] AND UKRAINE [3.63]

(data are in Bq/kg fresh weight for grain, potato and meat and in Bq/L for milk)

Caesium-137 soil deposition range Grain Potato Milk Meat

Belarus

>185 kBq/m

(contaminated districts of

the Gomel region)

30 (8–80) 10 (6–20) 80 (40–220) 220 (80–550)

37–185 kBq/m

(contaminated districts

of the Mogilev region)

10 (4–30) 6 (3–12) 30 (10–110) 100 (40–300)

Russian Federation

>185 kBq/m

(contaminated districts

of the Bryansk region)

26 (11–45) 13 (9–19) 110 (70–150) 240 (110–300)

37–185 kBq/m

(contaminated districts

of the Kaluga, Tula and Orel regions)

12 (8–19) 9 (5–14) 20 (4–40) 42 (12–78)

Ukraine

>185 kBq/m

(contaminated districts

of the Zhytomyr and Rovno regions)

32 (12–75) 14 (10–28) 160 (45–350) 400 (100–700)

37–185 kBq/m

(contaminated districts

of the Zhytomyr and Rovno regions)

14 (9–24) 8 (4–18) 90 (15–240) 200 (40–500)

Figure 3.34 shows the wide distribution of forests

across the European continent. Following the

Chernobyl accident, substantial radioactive contam-

ination of forests occurred in Belarus, the Russian

Federation and Ukraine, and in countries beyond

the borders of the former USSR, notably Finland,

Sweden and Austria (see Fig. 3.5). The degree of

forest contamination with

137

Cs in these countries

ranged from >10 MBq/m

in some locations to

between 10 and 50 kBq/m

, the latter range being

typical of

137

Cs deposition in several countries of

western Europe.

Since the Chernobyl accident it has become

apparent that the natural decontamination of

forests is proceeding extremely slowly. The net

export of

137

Cs from forest ecosystems was less than

1%/a [3.66, 3.67], so it is likely that, without artificial

intervention, it is the physical decay rate of

137

that will largely influence the duration over which

forests continue to be affected by the Chernobyl

fallout. Despite the fact that the absolute natural

losses of

137

Cs from forests are small, recycling of

radiocaesium within forests is a dynamic process in

which reciprocal transfers occur on a seasonal, or

longer term, basis between biotic and abiotic

components of the ecosystem. To facilitate

appropriate long term management of forests, a

reliable understanding of these exchange processes

is required. Much information on such processes

has been obtained from experiments and field

measurements, and many of these data have been

used to develop predictive mathematical models

[3.68].

3.4.2. Dynamics of contamination during the

early phase

Forests in the USSR located along the

trajectory of the first radioactive plume were

contaminated primarily as a result of dry

deposition, while further away, in countries such as

Austria and Sweden, wet deposition occurred and

resulted in significant hot spots of contamination.

Other areas in the USSR, such as the Mogilev

region in Belarus and Bryansk and some other

regions in the Russian Federation, were also

contaminated by deposition with rain.

Tree canopies, particularly at forest edges, are

efficient filters of atmospheric pollutants of all

kinds. The primary mechanism of tree contami-

nation after the Chernobyl accident was direct

interception of radiocaesium by the tree canopy,

which intercepted between 60% and 90% of the

initial deposition [3.66]. Within a 7 km radius of the

reactor this led to very high levels of contamination

on the canopies of pine trees, which, as a conse-

quence, received lethal doses of radiation from the

complex mixture of short and long lived radionu-

clides released in the accident. Gamma dose rates in

the days and weeks immediately following the

accident were in excess of 5 mGy/h in the area close

FIG. 3.34. Forest map of Europe. The darkest colour,

reen, indicates a proportion of 88% forest in the area,

while yellow indicates less than 10% [3.69].

AoL (litter layer)

AoF (organic)

AoH (organic)

A (mineral)

B (mineral)

Deposition

(wet/dry)

Canopy

interception

Biological

uptake

Stem flow

Through flow

Leaf/needle fall

Soil migration

FIG. 3.35. Major storages and fluxes in radionuclides of

contaminated forest ecosystems [3.70].

to the reactor. The calculated absorbed gamma dose

amounted to 80–100 Gy in the needles of pine trees.

This small area of forest became known as the Red

Forest, as the trees died and became a reddish

brown colour, which was the most readily

observable effect of radiation damage on organisms

in the area (see Section 6).

The contamination of tree canopies was

reduced rapidly over a period of weeks to months

due to wash-off by rainwater and the natural

process of leaf/needle fall (Fig. 3.35). Absorption of

radiocaesium by leaf surfaces also occurred,

although this was difficult to measure directly. By

the end of the summer of 1986, approximately 15%

of the initial radiocaesium burden in the tree

canopies remained, and by the summer of 1987 this

had been further reduced to approximately 5%.

Within this roughly one year period, therefore, the

bulk of radiocaesium was transferred from the tree

canopy to the underlying soil.

During the summer of 1986 radiocaesium

contamination of forest products such as

mushrooms and berries increased, which led to

increased contamination of forest animals such as

deer and moose. In Sweden activity concentrations

137

Cs in moose exceeded 2 kBq/kg fresh weight,

while those in roe deer were even higher [3.71].

3.4.3. Long term dynamics of radiocaesium in

forests

Within approximately one year after the initial

deposition, the soil became the major repository of

radiocaesium contamination within forests. Subse-

quently, trees and understorey plants became

contaminated due to root uptake, which has

continued as radiocaesium has migrated into the

soil profile. Just as for potassium, the nutrient

analogue rate of radiocaesium cycling within forests

is rapid and a quasi-equilibrium is reached a few

years after atmospheric fallout [3.72]. The upper,

organic rich, soil layers act as a long term sink but

also as a general source of radiocaesium for contam-

ination of forest vegetation, although individual

plant species differ greatly in their ability to

accumulate radiocaesium from this organic soil

(Fig. 3.36).

Release of radiocaesium from the system via

drainage water is generally limited due to its

fixation on micaceous clay minerals [3.67]. An

important role of forest vegetation in the recycling

of radiocaesium is the partial and transient storage

of radiocaesium, particularly in perennial woody

components such as tree trunks and branches, which

can have a large biomass. A portion of radiocaesium

taken up by vegetation from the soil, however, is

recycled annually through leaching and needle/leaf

fall, resulting in the long lasting biological availa-

bility of radiocaesium in surface soil. The stored

amount of radiocaesium in the standing biomass of

forests is approximately 5% of the total activity in a

temperate forest ecosystem, with the bulk of this

activity residing in trees.

Due to biological recycling and storage of

radiocaesium, migration within forest soils is limited

and the bulk of contamination in the long term

resides in the upper organic horizons (Fig. 3.37).

Slow downward migration of radiocaesium

1.E–04

1.E–03

1.E–02

1.E–01

1.E+00

1.E+01

1.E+02

90th percentile

Median

10th percentile

Soil

Tree

Understorey

Fungi

Game

FIG. 3.36. Calculated percentage distributions of radiocae-

ium in specified components of coniferous forest ecosys-

tems [3.73].

0.0

5.0

10.0

15.0

20.0

25.0

0 500 1000 1500 2000 2500 3000

1992

1993

1995

1996

1997

Soil depth (cm)

Caesium-137 (Bq/kg)

FIG. 3.37. Soil profiles of radiocaesium in a Scots pine

orest near Gomel in Belarus, 1992–1997 [3.74]. The hori-

zontal line indicates the boundary between organic and

mineral soil layers.

continues to take place, however, although the rate

of migration varies considerably with soil type and

climate.

The hydrological regime of forest soils is an

important factor governing radionuclide transfer in

forest ecosystems [3.75]. Depending on the hydro-

logical regime, the radiocaesium T

for trees,

mushrooms, berries and shrubs can vary over a

range of more than three orders of magnitude. The

minimum T

values were found for automorphic

(dry) forests and soils developed on even slopes

under free surface runoff conditions. The maximum

values are related to hydromorphic forests

developed under prolonged stagnation of surface

waters. Among other factors influencing radionu-

clide transfer in forests, the distribution of root

systems (mycelia) in the soil profile and the capacity

of different plants for radiocaesium accumulation

are of importance [3.76].

The vertical distribution of radiocaesium

within soil has an important influence on the

dynamics of uptake by herbaceous plants, trees and

mushrooms. It also influences the change in

external gamma dose rate with time. The upper soil

layers provide increasing shielding from radiation as

the peak of the contamination migrates downwards

(Fig. 3.38). The most rapid downward vertical

transfer was observed for hydromorphic forests

[3.75].

Once forests become contaminated with

radiocaesium, any further redistribution is limited.

Processes of small scale redistribution include resus-

pension [3.78], fire [3.79] and erosion/runoff, but

none of these processes are likely to result in any

significant migration of radiocaesium beyond the

location of initial deposition.

3.4.4. Uptake into edible products

Edible products obtained from forests include

mushrooms, fruits and game animals. In forests

affected by the Chernobyl deposition, each of these

products became contaminated. The highest levels

of contamination with radiocaesium have been

observed in mushrooms, due to their great capacity

to accumulate some mineral nutrients as well as

radiocaesium. Mushrooms provide a common and

significant food source in many of the affected

countries, particularly in the countries of the former

USSR. Changes with time in the contamination of

mushrooms reflect the bioavailability of

137

Cs in the

various relevant nutrient sources utilized by the

different mushroom species.

Some mushroom species exploit specific soil

layers for their nutrition, and the dynamics of

contamination of such species have been related to

the contamination levels of these layers [3.80]. The

high levels of contamination in mushroom species

are reflected in generally high soil to mushroom

transfer coefficients. However, these transfer coeffi-

cients (T

) are also subject to considerable

variability and can range from 0.003 to 7 m

/kg (i.e.

by a factor of approximately 2000 [3.81]). There are

significant differences in the accumulation of radio-

caesium in different species of mushroom (see

Fig. 3.39) [3.82]. In general, the saprotrophs and

wood degrading fungi, such as the honey fungus

7 MBq/m

4 MBq/m

0.7 MBq/m

1988 1989 1990 1991 1992 1993 1994 1995 1996

Dose rate (PGy/h)

Year

FIG. 3.38. Gamma dose rates in air at three forest locations

with different

137

Cs soil depositions in the Bryansk region

of the Russian Federation, 150 km north-east of Chernoby

[3.77].

M P S

Nutritional type

–5

–10

Ln(T

)

FIG. 3.39. Variation in the logarithm of T

with differen

nutritional types of mushroom [3.82]. M: mycorrhizal;

P: parasitic; S: saprotrophic nutritional type.

(Armillaria mellea), have a low level of contami-

nation, while those fungi forming symbioses with

tree roots (mycorrhizal fungi such as Xerocomus

and Lactarius) have a high uptake. The degree of

variability of mushroom contamination is illustrated

in Fig. 3.40, which also indicates the tendency for a

slow decrease in contamination during the 1990s.

Contamination of mushrooms in forests is

often much higher than that of forest fruits such as

bilberries. This is reflected in the T

for forest

berries, which ranges from 0.02 to 0.2 m

/kg [3.81].

Due to the generally lower radiocaesium levels and

the relative masses consumed, forest berries pose a

smaller radiological hazard to humans than do

mushrooms. However, both products contribute

significantly to the diet of grazing animals and

therefore provide a second route of exposure of

humans via game. Animals grazing in forests and

other seminatural ecosystems often produce meat

with high radiocaesium levels. Such animals include

wild boar, roe deer, moose and reindeer, but also

domestic animals such as cattle and sheep, which

may graze marginal areas of forests.

Most data on the contamination of game

animals such as deer and moose have been obtained

from western European countries in which the

hunting and eating of game is commonplace.

Significant seasonal variations occur in the body

content of radiocaesium in these animals due to the

seasonal availability of foods such as mushrooms

and lichens, the latter being particularly important

as a component of the diet of reindeer. Good time

series measurements have been obtained from the

Nordic countries and Germany. Figure 3.41 shows a

complete time series of annual average radio-

caesium activities for moose from 1986 to 2003 for

one hunting area in Sweden, and Fig. 3.42 shows

individual measurements of

137

Cs activity concen-

trations in the muscle of roe deer in southern

Germany. A major factor for the contamination of

game, and roe deer in particular, is the high concen-

tration of radiocaesium in mushrooms. The T

for

moose ranges from 0.006 to 0.03 m

/kg [3.81]. The

mean T

for moose in Sweden has been falling since

the period of high initial contamination, indicating

that the ecological half-life of radiocaesium in

moose is less than 30 years (i.e. less than the

physical half-life of

137

Cs).

3.4.5. Contamination of wood

Most forests in Europe and the former USSR

affected by the Chernobyl accident are planted and

managed for the production of timber. The export

of contaminated timber, and its subsequent

processing and use, could give rise to radiation

doses to people who would not normally be exposed

in the forest itself. Uptake of radiocaesium from

forest soils into wood is rather low; aggregated TFs

range from 0.0003 to 0.003 m

/kg. Hence wood used

for making furniture or the walls and floors of

houses is unlikely to give rise to significant radiation

exposure of people using these products [3.85].

However, the manufacture of consumer goods such

500 000

1 000 000

1 500 000

2 000 000

2 500 000

3 000 000

3 500 000

4 000 000

4 500 000

1990 1991 1992 1993 1994 1995 1996 1997 1998 1999

Year

Xerocomus

badius

Russula

paludosa

Suillus

luteus

Cantharellus

cibarius

Boletus

edulis

Bq/kg dry weight

FIG. 3.40. Caesium-137 activity concentrations (Bq/kg dry weight) in selected

mushroom species harvested in a pine forest in the Zhytomyr region of

Ukraine, approximately 130 km south-west of Chernobyl. The soil deposition

137

Cs at this site in 1986 was 555 kBq/m

. From Ref. [3.68].

as paper involves the production of both liquid and

solid waste that can become significantly contami-

nated with radiocaesium. The handling of this waste

by workers in paper pulp factories can give rise to

radiation doses within the industry [3.86].

Use of other parts of trees such as needles,

bark and branches for combustion may involve the

problem of disposal of radioactive wood ash. This

practice has increased in recent years due to the

upsurge in biofuel technology in the Nordic

countries, and the problem of radiocaesium in wood

ash has become significant because the radio-

caesium activity concentration in ash is a factor of

50–100 times higher than in the original wood. For

domestic users of firewood in contaminated regions,

a buildup of ash in the home and/or garden may also

give rise to external exposure to gamma radiation

from radiocaesium [3.85].

3.4.6. Expected future trends

Much effort has been put into developing

mathematical models that make use of the large

array of measurements of radiocaesium contami-

nation in forests since 1986 [3.68]. These models are

useful in helping to improve our understanding of

the way the Chernobyl contamination behaves in

forest ecosystems. Furthermore, they can also be

used to provide forecasts of future trends of

contamination, which can assist when making

decisions about the future management of contami-

nated regions.

Predictive models of radiocaesium behaviour

in forests are intended to quantify the fluxes and

distributions in the ecosystem over time. Forecasts

can be made for specific ecological compartments

such as the wood of trees and edible products such

as mushrooms. Figures 3.43 and 3.44 show examples

of such forecasts obtained using a variety of models.

Figure 3.43 shows predictions of the evolution of

radiocaesium activity in wood for two distinct types

of forest ecosystem with two age classes of trees.

This illustrates the importance of both soil

conditions and the stage of tree development at the

time of deposition in controlling the contamination

of harvestable wood. Figure 3.44 shows a summary

of 50 year forecasts for a pine forest in the

Zhytomyr region of Ukraine, approximately 130 km

south-west of Chernobyl. The figure shows the

degree of variability among the predictions made by

11 different models and also the inherent variability

within data collected from a single forest site. The

uncertainty in both monitoring data and among

models makes the task of forecasting future trends

of forest contamination rather difficult.

3.4.7. Radiation exposure pathways associated

with forests and forest products

Contaminated forests can give rise to

radiation exposures of workers in the forest and in

associated industries, as well as of members of the

general public. Forest workers receive direct

radiation exposure during their working hours, due

to the retention of radiocaesium in the tree canopy

and the upper soil layers. Similarly, members of the

public can receive external exposures from wood

products, for example furniture or wooden floors,

but, in addition, they may be exposed as a result of

the consumption of game, wild mushrooms and

berries containing radiocaesium. Forest margins

may also be used to graze domestic animals such as

891

8891

9891

0991

1991

2991

399

4991

991

6991

7991

8991

999

0002

2002

002

Year

Caesium-137 (Bq/kg)

900

800

700

600

500

400

300

200

100

FIG. 3.41. The average concentration of

137

Cs in moose in

one hunting area in Sweden, based on approximately 100

animals per year [3.83].

500

1000

1500

2000

2500

3000

Date

1 May 1986

25 Jan. 1989

22 Oct. 1991

18 Jul. 1994

13 Apr. 1997

8 Jan. 2000

4 Oct. 2002

Bq/kg dry weight

FIG. 3.42. Caesium-137 activity concentrations in the

muscle of roe deer (Capreolus capreolus) harvested in a

orest close to Bad Waldsee, southern Germany. The tota

deposition of

137

Cs at this site in 1986 was 27 kBq/m

[3.84].

cattle and sheep. This can lead to the milk of these

animals becoming contaminated and to human

exposure as a result of the consumption of dairy

products and meat. A further exposure pathway

results from the collection and use of firewood for

domestic purposes. This can give rise to exposures

both in the home and in the garden if wood ash is

used as a domestic fertilizer. Also, the industrial use

of forest products for energy production can give

rise to exposure both of workers and of members of

the public. Quantitative information on human

radiation doses associated with forests and forest

products is given in Ref. [3.85] and in Section 6 of

this report.

Another set of important exposure pathways

results from the harvesting, processing and use of

timber and wood products from contaminated

forest areas. Timber and wood products become

sources of potential exposure once they are

exported from the forest, often over considerable

distances and sometimes across national borders.

The relative importance of these exposure pathways

has been evaluated and quantified [3.85].

3.5. RADIONUCLIDES IN AQUATIC

SYSTEMS

3.5.1. Introduction

Radioactive material from Chernobyl affected

surface water systems in many parts of Europe. The

majority of the radioactive fallout, however, was

deposited in the catchment of the Pripyat River,

which forms an important component of the

Dnieper River–reservoir system, one of the larger

surface water systems in Europe [3.13]. After the

accident, therefore, there was particular concern

over contamination of the water supply for the area

along the Dnieper cascade of reservoirs covering a

distance of approximately 1000 km to the Black Sea

(see Figs 3.6–3.9). Other large river systems in

Europe, such as the Rhine and Danube, were also

affected by fallout, although the contamination

levels in those rivers were not radiologically

significant [3.5, 3.6].

Initial radionuclide concentrations in river

water in parts of Belarus, the Russian Federation

and Ukraine were relatively high compared both

with other European rivers and with the safety

standards for radionuclides in drinking water. The

contamination was due to direct fallout on to river

surfaces and runoff of contamination from

catchment areas. During the first few weeks after

the accident, activity concentrations in river waters

declined rapidly, because of the physical decay of

short lived isotopes and absorption of radionuclides

on to catchment soils and bottom sediments. In the

longer term, the long lived

137

Cs and

Sr comprised

the major component of contamination of aquatic

ecosystems. Although the levels of these

radionuclides in rivers were low after the initial

0 10 20 30 40 50

Time after deposition (years)

0.0

2.0

4.0

6.0

8.0

10.0

Caesium-137 in wood (Bq/kg)

FIG. 3.43. Predicted

137

Cs activity concentration in wood

or different types of forest soil and ages of trees calculated

using a computer model, FORESTLAND, for a deposition

of 1 kBq/m

[3.87]. 1, 2: automorphic soil, 3, 4: semi-

hydromorphic soil; 1, 3: initial age 20 years; 2, 4: initial age

80 years.

500

1000

1500

2000

2500

3000

3500

1986 1988 1990 1992 1994 1996 1998 2000

Year

Max.

Median

Min.

Bq/kg dry weight

FIG. 3.44. Summary of predictions of pine wood contami-

nation with

137

Cs in the Zhytomyr region of Ukraine made

by use of 11 models within the IAEA’s BIOMASS

rogram. Caesium-137 soil deposition was abou

55 kBq/m

. Max., Median and Min. indicate maximum,

median and minimum values of pooled model predictions,

respectively. The points show means of measured values,

and the broken lines indicate the maximum and minimum

values of measurements [3.88].

peak, temporary increases in activity concentrations

during flooding of the Pripyat River caused serious

concern in areas using water from the Dnieper

cascade.

Lakes and reservoirs were contaminated by

fallout on the water surface and by transfers of

radionuclides from the surrounding catchment

areas. Radionuclide concentrations in water

declined rapidly in reservoirs and in those lakes

with significant inflow and outflow of water (open

lake systems). In some cases, however, activity

concentrations of radiocaesium in lakes remained

relatively high due to runoff from organic soils in

the catchment. In addition, internal cycling of radio-

caesium in closed lake systems (i.e. lakes with little

inflow and outflow of water) led to much higher

activity concentrations in their water and aquatic

biota than were typically seen in open lakes and

rivers.

Bioaccumulation of radionuclides (particu-

larly radiocaesium) in fish resulted in activity

concentrations (both in the most affected regions

and in western Europe) that were in some cases

significantly above the permissible levels for

consumption [3.89–3.94]. In some lakes in Belarus,

the Russian Federation and Ukraine these

problems have continued to the present day and

may continue for the foreseeable future. Freshwater

fish provide an important source of food for many

inhabitants of the contaminated regions. In the

Dnieper cascade in Ukraine, commercial fisheries

catch more than 20 000 t of fish per year. In some

parts of western Europe, particularly parts of

Scandinavia, radiocaesium activity concentrations

in fish are still relatively high [3.95].

The marine systems closest to Chernobyl are

the Black Sea and the Baltic Sea — both several

hundred kilometres from the site. Radioactivity in

the water and fish of these seas has been intensively

studied since the Chernobyl accident. Since the

average deposition to these seas was relatively low,

and owing to the large dilution in marine systems,

radionuclide concentrations were much lower than

in freshwater systems [3.96].

3.5.2. Radionuclides in surface waters

3.5.2.1. Distribution of radionuclides between

dissolved and particulate phases

Retention of radionuclide fallout by

catchment soils and river and lake sediments plays

an important role in determining subsequent

transport in aquatic systems. The fraction of a radio-

nuclide that is adsorbed to suspended particles

(which varies considerably in surface waters)

strongly influences both its transport and its bioac-

cumulation. Most

Sr is present in the dissolved

phase (0.05–5% in the solid phase), but in the near

zone a significant proportion of strontium fallout

was in the form of fuel particles. The soils of the

CEZ are heavily contaminated with

Sr (see

Fig. 7.7), and some of it is washed off during flood

events when the low lying areas become inundated.

In the Pripyat River, during the first decade

after the accident approximately 40–60% of radio-

caesium was found in the particulate phase [3.97],

but estimates in other systems [3.98] vary from 4%

to 80%, depending on the composition and concen-

tration of the suspended particles and on the water

chemistry. Fine clay and silt particles absorb radio-

caesium more effectively than larger, less reactive

sand particles. Sandy river beds, even close to the

reactor, were relatively uncontaminated, but fine

particles transported radiocaesium over relatively

large distances. The settling of fine particles in the

deep parts of the Kiev reservoir led to high levels of

contamination of bed sediments [3.99].

Measurements of the distribution of radio-

nuclides between the dissolved and particulate

phases in Pripyat River water showed that the

strength of adsorption to suspended particles

increases in the following order:

Sr,

137

Cs,

transuranic elements (

239,240

Pu,

241

Am) [3.100].

There is a possibility that natural organic colloids

may determine the stability of transuranic elements

in surface water and in their transport from contam-

inated soil; such colloids have less effect on

Sr and

137

Cs [3.101].

Generally in marine systems, lower particle

sorption capacities and higher concentrations of

competing ions (i.e. higher salinity) tend to make

particle sorption of radionuclides less significant

than in freshwaters. In the Baltic Sea after the

Chernobyl accident less than 10% of

137

Cs was

bound to particles and the average particulate

sorbed fraction was approximately 1% [3.102,

3.103]. In the Black Sea the particulate bound

fraction of

137

Cs was less than 3% [3.96].

3.5.2.2. Radioactivity in rivers

Initial radioactivity concentrations in rivers

close to Chernobyl (the Pripyat, Teterev, Irpen and

Dnieper Rivers) were largely due to direct

deposition of radioactivity on to the river surface.

The highest concentrations of radionuclides were

observed in the Pripyat River at Chernobyl, where

the

131

I activity concentration was up to 4440 Bq/L

(Table 3.7). In all water bodies the radioactivity

levels declined rapidly during the first few weeks,

due to decay of short lived isotopes and absorption

of nuclides to catchment soils and river bed

sediments.

Over longer time periods after the fallout

occurred, relatively long lived

Sr and

137

Cs retained

in catchment soils are slowly transferred to river

water by erosion of soil particles and by desorption

from soils. The rates of transfer are influenced by the

extent of soil erosion, the strength of radionuclide

binding to catchment soils and migration down the

soil profile. An example time series of

Sr and

137

activity concentrations in the water of the Pripyat

River near Chernobyl is shown in Fig. 3.45.

After the Chernobyl accident, water

monitoring stations were established within the

exclusion zone and along the major rivers to

determine the concentrations of radionuclides and

their total fluxes. Measurements from these stations

allow estimates to be made of radionuclide fluxes of

Sr and

137

Cs into and out of the CEZ. The

migration of

137

Cs has decreased markedly with time

and shows relatively little change from upstream to

downstream of the CEZ (see Fig. 3.46(a)).

In contrast, the transboundary migration of

Sr has fluctuated yearly depending on the extent

of annual flooding along the banks of the Pripyat

River (Fig. 3.46(b)). There is also a significant flux

from the CEZ — fluxes downstream of the zone are

much higher than those upstream. Note, however,

that the extent of washout of radionuclides by the

river system is only a very small percentage of the

total inventory contained in the catchment area.

The decline in

Sr and

137

Cs activity concen-

trations occurred at a similar rate for different rivers

in the vicinity of Chernobyl and in rivers in western

Europe [3.108]. Measurements of

137

Cs activity

concentration in different European rivers

(Fig. 3.47) show a range of approximately a factor of

TABLE 3.7. MAXIMUM RADIONUCLIDE

ACTIVITY CONCENTRATIONS (DISSOLVED

PHASE) MEASURED IN THE PRIPYAT RIVER

AT CHERNOBYL [3.91, 3.104, 3.105]

Maximum concentration

(Bq/L)

Caesium-137 1591

Caesium-134 827

Iodine-131 4440

Strontium-90 30

Barium-140 1400

Molybdenum-99 670

Ruthenium-103 814

Ruthenium-106 271

Cerium-144 380

Cerium-141 400

Zirconium-95 1554

Niobium-95 420

Plutonium-241 33

Plutonium-239, 240 0.4

0.01

0.10

1.00

10.00

1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999

Bq/L

1 Strontium-90

2 Caesium-137, dissolved

Year

IG. 3.45. Monthly averaged

Sr and

137

Cs activity concen-

trations in the Pripyat River near Chernobyl [3.106].

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

8.0

9.0

Áåëàÿ Ñîðîêà

×åðíîáûëü

Belaya Soroka

Chernobyl

0.0

2.0

4.0

6.0

8.0

10.0

12.0

14.0

16.0

18.0

20.0

1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000

Yea r

Year

Caesium-137 (TBq)Strontium-90 (TBq)

(a)

(b)

IG. 3.46. Annual fluxes of

137

Cs (a) and

Sr (b) in the

ripyat River at Belaya Soroka, near the Belarus–Ukraine

order (inlet to the CEZ), and downstream of Chernobyl

(outlet of the CEZ) [3.107].

30, even when differences in fallout have been

accounted for. In small catchments [3.67, 3.112,

3.113], highly organic soils (particularly saturated

peat soils) released up to an order of magnitude

more radiocaesium to surface waters than some

mineral soils. Thus rivers in Finland with large areas

of wet organic soils in the catchment have higher

radiocaesium concentrations (per unit of

radioactive fallout) than rivers with predominantly

mineral catchments [3.109, 3.111].

3.5.2.3. Radioactivity in lakes and reservoirs

In the affected areas of Belarus, the Russian

Federation and Ukraine many lakes were signifi-

cantly contaminated by radionuclides. In most lakes

radionuclides were well mixed throughout the lake

water during the first days to weeks after the fallout

occurred. In deep lakes such as Lake Zurich (mean

depth 143 m), however, it took several months for

full vertical mixing to take place [3.114]. In some

areas of northern Europe lakes were covered with

ice at the time of the accident and maximum activity

concentrations in lake waters were only observed

after the ice melted.

Radionuclides deposited to a lake or reservoir

are removed through water outflow and by transfer

to bed sediments. As in rivers, radiocaesium activity

concentrations in lakes declined relatively rapidly

during the first weeks to months after the fallout.

This was followed by a slower decline over a period

of years as radiocaesium became more strongly

absorbed to catchment soils and lake sediments and

migrated to deeper layers in the soil and sediments.

Figure 3.48 illustrates the temporal change in

137

activity concentration in lakes using measurements

from Lake Vorsee, a small shallow lake in

Germany.

Inputs to lakes also result from transport of

radionuclides from contaminated catchment soils.

In the longer term (secondary phase),

137

Cs activity

concentrations in Lake Vorsee remained much

higher than in most other lakes due to inputs of

137

Cs from organic soils in the catchment and

remobilization from bed sediments (Fig. 3.48

[3.93]). In Devoke Water (UK), radiocaesium

flowing from organic catchment soils maintained

activity concentrations in the water that were

approximately an order of magnitude higher than in

nearby lakes with mineral catchments [3.112]. In

some cases, lakes in western Europe with organic

catchments had activity concentrations in water and

fish similar to those in the more highly contaminated

areas of Belarus and Ukraine.

Long term contamination can also be caused

by remobilization of radionuclides from bed

sediments [3.115]. In some shallow lakes where

there is no significant surface inflow and outflow of

water, the bed sediments play a major role in

controlling radionuclide activity concentration in

the water. Such lakes have been termed ‘closed’

lakes [3.105, 3.116]. The more highly contaminated

water bodies in the Chernobyl affected areas are the

closed lakes of the Pripyat floodplain within the

CEZ. During 1991

137

Cs activity concentrations in

these lakes were up to 74 Bq/L (Lake Glubokoye)

and

Sr activity concentrations were between 100

and 370 Bq/L in six of the 17 water bodies studied

[3.105]. Seventeen years after the accident there

were still relatively high activity concentrations in

the closed lakes in the CEZ [3.117] and at quite

large distances from the reactor; for example,

during 1996 Lakes Kozhanovskoe and Svyatoe in

0 5 10 15

Time since accident (years)

Kymijoki

Kokemaenjoki

Oulujoki

Kemijoki

Tornionjoki

Dora Baltea

Dnieper

Sozh

Iput

Besed

Pripyat (Mozyr)

Danube

Pripyat (Chernobyl)

Caesium-137 in water per Bq/m

of fallout (m

–1

)

0.1

0.01

0.001

0.0001

0.00001

FIG. 3.47. Caesium-137 activity concentration in differen

rivers per unit of deposition [3.109–3.111].

0.01

0.1

100

1000

10 000

0 5 10 15 20

Time (years)

Water

Pike (predatory)

Small Cyprinidae (non-predatory)

Caesium-137 in water (Bq/L) and fish (Bq/kg)

FIG. 3.48. Time series of

137

Cs in Lake Vorsee, Germany

[3.93].

the Bryansk region of the Russian Federation

(approximately 200 km from Chernobyl) contained

0.6–1.5 Bq/L of

Sr and 10–20 Bq/L of

137

(Fig. 3.49). Activity concentrations in water were

higher than in many lakes close to Chernobyl,

because of remobilization from sediments in these

closed lakes [3.116]. The Russian intervention level

for drinking water of 11 Bq/L for

137

Cs [3.119] is

shown for comparison.

(a) Chernobyl cooling pond

The Chernobyl cooling pond covers an area of

approximately 23 km

and contains approximately

149 × 10

of water. It is located between the

former Chernobyl nuclear power plant and the

Pripyat River. The total inventory of radionuclides

in the pond is in excess of 200 TBq (about 80% is

137

Cs, 10%

Sr, 10%

241

Pu and less than 0.5% each

is of

238

Pu,

239

Pu,

240

Pu and

241

Am), with the deep

sediments containing most of the radioactivity. The

Sr annual flux to the Pripyat River from the

cooling pond via groundwater was estimated in a

recent study to be 0.37 TBq [3.120]. This is a factor

of 10–30 less than the total annual

Sr fluxes in the

Pripyat River in recent years. Thus the cooling pond

is not a significant source of

Sr contamination of

the Pripyat River. Radionuclide activity concentra-

tions in the cooling pond water (Fig. 3.50) are

currently low, at 1–2 Bq/L. Seasonal variations of

137

Cs concentration are caused by changes in algae

and phytoplankton biomass [3.121].

(b) Reservoirs of the Dnieper cascade

The Dnieper cascade reservoirs were signifi-

cantly affected, due to both atmospheric fallout and

riverine inputs from the contaminated zones (see

Fig. 3.6). The different affinities of

137

Cs and

Sr for

suspended matter influenced their transport

through the Dnieper system. Caesium-137 tends to

become fixed on to clay sediments, which are

deposited in the deeper sections of the reservoirs,

particularly in the Kiev reservoir (Fig. 3.51). Owing

to this process, very little

137

Cs flows through the

cascade of reservoirs, and consequently the present

concentration entering the Black Sea is indistin-

guishable from the background level.

However, although the

Sr activity concen-

tration decreases with distance from the source

1992 1994 1996 1998 2000 2002

Year

Svyatoe

Kozhanovskoe

IL, DIL

Caesium-137 (Bq/L)

FIG. 3.49. Dynamics of

137

Cs activity concentration in the

water of Lakes Svyatoe and Kozhanovskoe (Russian

Federation), approximately 200 km from Chernoby

[3.118].

0.1

1.0

10.0

100.0

1000.0

Date

Caesium-137

Strontium-90

Concentration (Bq/L)

5 Jan. 1986 5 Jan. 1988 4 Jan. 1990 4 Jan. 1992 3 Jan. 1994 3 Jan. 1996 2 Jan. 1998 2 Jan. 2000 1 Jan. 2002

FIG. 3.50. Monthly averaged

137

Cs and

Sr activity concentrations in the water of the Chernobyl cooling pond

[3.121].

(mainly due to dilution), about 40–60% passes

through the cascade and reaches the Black Sea.

Figure 3.52 shows the trend in average annual

activity concentration in the Dnieper reservoirs

since the accident. As

137

Cs is trapped by sediments

in the reservoir system, activity concentrations in

the lower part (Novaya Kakhovka) of the system

are orders of magnitude lower than in the Kiev

reservoir (Vishgorod). In contrast,

Sr is not

strongly bound by sediments, so activity concentra-

tions in the lower part of the river–reservoir system

are similar to those measured in the Kiev reservoir.

The peaks in

Sr activity concentration in the

reservoirs of the Dnieper cascade (Fig. 3.52) were

caused by flooding of the most contaminated

floodplains in the CEZ; for example, flooding of the

Pripyat River, caused by blockages of the river by

ice in the winter of 1990–1991, led to temporary

significant increases in

Sr in this system, but did

not significantly affect

137

Cs levels. Activity concen-

trations of

Sr in the river water increased from

about 1 to 8 Bq/L for a five to ten day period [3.105].

Similar events took place during the winter flood of

1994, during summer rainfall in July 1993 and

during the high spring flood in 1999 [3.122].

Small amounts of radionuclides are eroded

from soils and transferred to rivers, lakes and

eventually the marine system. Such transfers can

take place through erosion of surface soil particles

and by runoff in the dissolved phase. Studies of

nuclear weapon tests and of Chernobyl

Sr in rivers

[3.109, 3.110, 3.123, 3.124] suggest long term loss

rates of about 1–2%/a or less from the terrestrial

environment to rivers. Thus, in the long term, runoff

of radionuclides does not significantly reduce the

amount of radioactivity in the terrestrial system,

although it does result in continuing (low level)

contamination of river and lake systems.

3.5.2.4. Radionuclides in freshwater sediments

Bed sediments are an important long term

sink for radionuclides. Radionuclides can become

attached to suspended particles in lakes, which then

fall and settle on bed sediments. Radionuclides in

lake water can also diffuse into bed sediments.

These processes of radionuclide removal from lake

water have been termed ‘self-cleaning’ of the lake

or reservoir [3.114].

In the Chernobyl cooling pond approximately

one month after the accident, most of the radioac-

tivity was found in bed sediments [3.91, 3.97]. In the

long term, approximately 99% of the radiocaesium

in a lake is typically found in the bed sediment. In

measurements in Lake Svyatoe (Kostiukovichy

region, Belarus) during 1997, approximately 3 × 10

Bq of

137

Cs was in the water and 2.5 × 10

Bq was in

sediments [3.125]. In Lake Kozhanovskoe in the

Russian Federation, approximately 90% of the

radiostrontium was found in the bed sediments

during 1993–1994 [3.126].

In the rapidly accumulating sediments of the

Kiev reservoir, the layer of maximum radiocaesium

Dnieper River

Contaminated areas (kBq/m

)

Less than 37

37–74

74–185

185–370

370–555

555–925

Pripyat River

Teterev River

Irpen River

Vishgorod

Kiev hydroelectric power plant

>925

FIG. 3.51. Caesium-137 in the bottom sediments of the Kiev reservoir [3.97].

concentration is now buried several tens of

centimetres below the sediment surface (Fig. 3.53).

In more slowly accumulating sediments, however,

the peak in radiocaesium activity concentration

remains near the sediment surface. Peaks in the

sediment layer contamination in 1988 and 1993

reflect the consequences of high summer rainfall

floods and soil erosion events.

Close to Chernobyl, a high proportion of the

deposited radioactive material was in the form of

fuel particles (see Section 3.1). Radionuclides

deposited as fuel particles are generally less mobile

than those deposited in a dissolved form. In the

sediments of Lake Glubokoye in 1993, most fuel

particles remained in the surface 5 cm of sediment

[3.126]. Fuel particle breakdown was at a much

lower rate in lake sediments than in soils. Studies in

the cooling pond have shown that the half-life of

fuel particles in sediments is approximately 30

years, so by 2056 (70 years after the Chernobyl

accident) one quarter of the radioactive material

deposited as fuel particles in the cooling pond will

still remain in fuel particle form [3.39].

3.5.3. Uptake of radionuclides to freshwater fish

Consumption of freshwater fish is an

important part of the aquatic pathway for the

transfer of radionuclides to humans. Although the

transfer of radionuclides to fish has been studied in

many countries, most attention here is focused on

Belarus, the Russian Federation and Ukraine,

because of the higher contamination of water

bodies in these areas.

3.5.3.1. Iodine-131 in freshwater fish

There are limited data on

131

I in fish.

Iodine-131 was rapidly absorbed by fish in the Kiev

reservoir, with maximum concentrations in fish

being observed in early May 1986 [3.91]. Activity

concentrations in fish muscle declined from around

6000 Bq/kg fresh weight on 1 May 1986 to 50 Bq/kg

fresh weight on 20 June 1986. This represents a rate

of decline similar to that of the physical decay of

131

I. Owing to the rapid physical decay,

131

I activity

concentrations in fish became insignificant a few

months after the accident.

3.5.3.2. Caesium-137 in freshwater fish and other

aquatic biota

During the years following the Chernobyl

accident there have been many studies of the levels

of radiocaesium contamination of freshwater fish.

As a result of high radiocaesium bioaccumulation

factors, fish have remained contaminated in some

areas, despite low radiocaesium levels in water.

Uptake of radiocaesium into small fish was

relatively rapid, the maximum concentration being

observed during the first weeks after the accident

[3.93, 3.95]. Due to the slow uptake rates of radio-

caesium in large predatory fish (pike, eel),

100

1000

1987 1989 1991 1993 1995 1997 1999 2001

Vishgorod Novaya Kakhovka

100

200

300

400

500

600

1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002

Year

Vishgorod Novaya Kakhovka

Bq/m

Year

(a)

(b)

FIG. 3.52. Annually averaged

137

Cs (a) and

Sr (b)

activity concentrations in the water of the Kiev reservoir

(Vishgorod) near the dam and in the Kakhovka reservoir

of the Dnieper cascade [3.107].

0 5000 10 000 15 000 20 000 25 000 30 000

0–2

8–10

16–18

24–26

32–34

40–42

48–50

56–58

64–66

72–74

Z (cm)

Cs ,

bottom

1988 hi

h summer rain

1993 high summer rain

fld

137

Bq/kg

Cs ,

bottom

1986

1988 high summer rain

1993 high summer rain

137

FIG. 3.53. Activity concentrations of

137

Cs in the deep sil

deposits in the upper part of the Kiev reservoir, 1998

[3.106].

maximum activity concentrations were not

observed until six to 12 months after the fallout

event [3.93, 3.127] (Fig. 3.48).

In the Chernobyl cooling pond,

137

Cs activity

concentrations in carp, silver bream, perch and pike

were about 100 kBq/kg fresh weight in 1986,

declining to a few tens of kBq/kg in 1990 [3.89, 3.91]

and 2–6 kBq/kg in 2001. In some closed lakes in the

vicinity of the Chernobyl nuclear power plant

[3.121] the

137

Cs activity concentration in predatory

fish 15 years after the accident was 10–27 kBq/kg

fresh weight. Typical changes with time in

137

Cs in

two fish species over 16 years after the accident are

illustrated in Fig. 3.54.

In the Kiev reservoir,

137

Cs activity concentra-

tions in fish were 0.6–1.6 kBq/kg fresh weight (in

1987) and 0.2–0.8 kBq/kg fresh weight (for 1990–

1995), and declined to 0.2 kBq/kg or less for adult

non-predatory fish in 2002. Values for predatory

fish species were 1–7 kBq/kg during 1987 and

0.2

1.2 kBq/kg from 1990 to 1995 [3.106].

In the lakes of the Bryansk region of the

Russian Federation, approximately 200 km from

Chernobyl,

137

Cs activity concentrations in a

number of fish species varied within the range of

0.2–19 kBq/kg fresh weight during the period 1990–

1992 [3.126, 3.150]. In shallow closed lakes such as

Lake Kozhanovskoe (Bryansk region of the Russian

Federation) and Lake Svyatoe (Kostiukovichy

region in Belarus),

137

Cs activity concentrations in

fish have declined slowly in comparison with fish in

rivers and open lake systems, due to the slow

decline in

137

Cs activity concentrations in the water

of the lakes [3.92, 3.116].

In western Europe, lakes in some parts of

Finland, Norway and Sweden were particularly

heavily contaminated. About 14 000 lakes in

Sweden had fish with

137

Cs activity concentrations

above 1500 Bq/kg fresh weight (the Swedish

guideline value) in 1987 [3.90]. In some alpine lakes

in Germany,

137

Cs activity concentrations in pike

were up to 5000 Bq/kg fresh weight shortly after the

Chernobyl accident [3.93]. In Devoke Water in the

UK Lake District, perch and brown trout contained

around 1000 Bq/kg fresh weight in 1988, declining

slowly to a few hundreds of Bq/kg in 1993 [3.129].

Bioaccumulation of radiocaesium in fish is

dependent on a number of factors. The presence of

potassium in a lake or river influences the rate of

accumulation of radiocaesium in fish because of its

chemical similarity to caesium [3.130]. Strong

inverse relationships were observed between the

lake water potassium concentration and the

137

activity concentration in fish following nuclear

weapon testing [3.128, 3.130] and the Chernobyl

accident [3.94]. In the long term, activity concentra-

tions in predatory fish were significantly higher than

in non-predatory fish, and large fish tended to have

higher activity concentrations than small fish. The

higher activity concentration in large fish is termed

the ‘size effect’ [3.127, 3.131] and is due to metabolic

and dietary differences. In addition, older, larger

fish were exposed to higher levels of

137

Cs in the

water than younger, smaller fish.

The differences in the bioaccumulation of

radiocaesium in different fish species can be signif-

icant; for example, in Lake Svyatoe in Belarus, the

levels in large pike and perch (predatory fish) were

five to ten times higher than in non-predatory fish

such as roach. Similarly, bioaccumulation factors in

lakes with a low potassium concentration can be

one order of magnitude higher than in lakes with a

high potassium concentration. Thus it was observed

[3.94] that fish from lakes in the agricultural areas of

Belarus (where runoff of potassium fertilizer is

significant) had lower bioaccumulation factors than

fish from lakes in seminatural areas.

3.5.3.3. Strontium-90 in freshwater fish

Strontium behaves, chemically and biologi-

cally, in a similar way to calcium. Strontium is most

100

200

300

400

500

600

700

800

900

1000

5891

689

789

8891

989

099

1991

3991

4991

991

6991

991

9991

0002

200

400

600

800

1000

1200

1400

1600

1800

6891

7891

8891

9891

1991

2991

4991

5991

6991

7991

9991

0002

Year

Bq/kg fresh weight Bq/kg fresh weight

(a)

(b)

FIG. 3.54. Averaged

137

Cs activity concentrations in non-

redatory (bream (a)) and predatory (pike (b)) fish; the

ish are from the Kiev reservoir [3.106].

strongly bioaccumulated in low calcium (‘soft’)

waters. The relatively low fish–water bioaccumu-

lation factors for

Sr (of the order of 10

L/kg) and

the lower fallout of this isotope meant that

activity concentrations in fish were typically much

lower than those of

137

Cs. In the Chernobyl cooling

pond,

Sr activity concentrations were around

2 kBq/kg (whole fish) in fish during 1986, compared

with around 100 kBq/kg for

137

Cs in 1993 [3.91]. In

2000, for the most contaminated lakes around

Chernobyl, the maximum level of

Sr concentration

in the muscles of predatory and non-predatory fish

varied between 2 and 15 Bq/kg fresh weight. In

2002–2003,

Sr in fish in reservoirs of the Dnieper

cascade was only 1–2 Bq/kg, which is close to the

pre-Chernobyl level. Freshwater molluscs showed

significantly higher bioaccumulation of

Sr than

fish. In the Dnieper River, molluscs had approxi-

mately ten times more

Sr in their tissues than in

fish muscle [3.132]. Similarly, the bioaccumulation

Sr in the bones and skin of fish is approximately

a factor of ten times higher than in muscle [3.130].

3.5.4. Radioactivity in marine ecosystems

Marine ecosystems were not seriously affected

by fallout from Chernobyl, the nearest seas to the

reactor being the Black Sea (around 520 km) and

the Baltic Sea (about 750 km). The primary route of

contamination of these seas was atmospheric

fallout, with smaller inputs from riverine transport

occurring over the years following the accident.

Surface deposition of

137

Cs was approximately

2.8 PBq over the Black Sea [3.96, 3.133] and 3.0 PBq

over the Baltic Sea [3.105].

3.5.4.1. Distribution of radionuclides in the sea

Radioactive fallout on to the surface of the

Black Sea was not uniform and mainly occurred

during 1 and 3 May 1986 [3.105, 3.133]. In the Black

Sea surface water concentrations of

137

Cs ranged

from 15 to 500 Bq/m

in June–July 1986, although by

1989 horizontal mixing of surface waters had

resulted in relatively uniform concentrations in the

range of 41–78 Bq/m

[3.105], which by 2000 had

declined to between 20 and 35 Bq/m

[3.96].

In addition to the caesium isotopes, short lived

radionuclides such as

144

Ce and

106

Ru were

observed. The inventory of

137

Cs in the water of the

Black Sea due to the Chernobyl deposition doubled

the existing inventory of

137

Cs from global fallout

from atmospheric nuclear weapon testing to

approximately 3100 TBq. The amount of

increased by 19% in comparison with the pre-

Chernobyl period and was estimated to be about

1760 TBq [3.96, 3.105]. Vertical mixing of surface

deposited radioactivity also reduced the maximum

concentrations observed in water over the months

to years after the fallout. Removal of radioactivity

to deeper waters steadily reduced

137

Cs activity

concentrations in the surface (0–50 m) layer of the

Black Sea. The present situation with regard to the

Black Sea marine environment is shown in Table 3.8

[3.96].

A significant proportion of the

137

Cs,

Sr and

239,240

Pu in the Black Sea originated from nuclear

weapon testing rather than from the Chernobyl

accident. The riverine radionuclide input to the

Black Sea was much less significant than direct

atmospheric fallout to the sea surface. Over the

period 1986–2000, riverine input of

137

Cs was only

4–5% of the atmospheric deposition, although

riverine inputs were more significant, being approx-

imately 25% of the total inputs from atmospheric

deposition [3.96, 3.134]. For the Baltic Sea, riverine

inputs were at a similar level as for the Black Sea,

being approximately 4% and 35% of atmospheric

fallout for

137

Cs and

Sr, respectively [3.135]. The

greater relative riverine input of

Sr is due to its

weaker adsorption to catchment soils and lake and

TABLE 3.8. RADIONUCLIDES IN VARIOUS SAMPLES TAKEN FROM THE BLACK SEA COAST

DURING 1998–2001 [3.96]

Environmental sample Caesium-137 Strontium-90

Plutonium-239,

240

Sea water (Bq/m

) 14–29 12–28 (2.4–28) × 10

–3

Beach sand and shells (Bq/kg) 0.9–8.0 0.5–60 (shell) (80–140) × 10

–3

Seaweeds, Cystoseira barbata (Bq/kg fresh weight) 0.2–2.3 0.4–0.9 (9.0–14) × 10

–3

Mussels, Mytilus galloprovincialis (tissue, Bq/kg fresh weight) 0.3–1.7 0.02–3.2 (1.5–2.5) × 10

–3

Fish, Sprattus sprattus, Trashurus (Bq/kg fresh weight) 0.2–6.0 0.02–0.7 (0.3–0.5) × 10

–3

river sediments and to lower

Sr atmospheric

fallout (compared with

137

Cs) at large distances

from the Chernobyl reactor site. Sedimentation

processes in the marine environment, as in the

freshwater environment, are an important factor in

the ‘self-purification’ of the aquatic ecosystem.

However, the sedimentation rate for the Black Sea

is relatively low [3.96].

Data presented in Fig. 3.55 demonstrate that,

in the central deep basin of the Black Sea, the

Chernobyl deposition is covered by a layer of less

than 1 cm of sediment formed since the accident

[3.96].

Due to dilution and sedimentation, the

concentration of

137

Cs quickly declined, reducing

the seawater contamination at the end of 1987 to

two to four times lower than that observed in the

summer of 1986. The average

137

Cs activity concen-

tration in the Baltic Sea estimated in Ref. [3.136] for

the initial period after deposition was approxi-

mately 50 Bq/m

, with maximum values two to four

times greater being observed in some areas of the

sea.

3.5.4.2. Transfers of radionuclides to marine biota

Bioaccumulation of radiocaesium and radio-

strontium in marine systems is generally lower than

in freshwater, because of the much higher content

of competing ions in saline water. The lower bioac-

cumulation of

137

Cs and

Sr in marine systems, and

the large dilution in these systems, meant that

activity concentrations in marine biota after the

Chernobyl accident were relatively low. Table 3.8

gives examples of

137

Cs,

Sr and

239,240

Pu in water

and marine biota of the Black Sea during the period

1998–2001 [3.96]. Detailed data on Baltic Sea fish

contamination during the post-Chernobyl decades

are available in Ref. [3.136], which shows that most

species of fish had a relatively low level of radio-

caesium contamination, in most cases in the range

of 30–100 Bq/kg or less during the period up to

1995.

3.5.5. Radionuclides in groundwater

3.5.5.1. Radionuclides in groundwater: Chernobyl

exclusion zone

Sampling of groundwater in the affected areas

showed that radionuclides can be transferred from

surface soil to groundwater. However, the level of

the groundwater contamination in most areas

(excluding locations of radioactive waste storage

and the Chernobyl shelter industrial site) is low.

Furthermore, the rates of migration from the soil

surface to groundwater are also very low. Some

areas in the CEZ with relatively fast radionuclide

migration to the aquifers were found in areas with

morphological depressions [3.137]. Horizontal

fluxes of radionuclides in groundwaters are also

very low because of the slow flow velocity of

groundwaters and high retardation of radionuclides

[3.138].

Short lived radionuclides are not expected to

affect groundwater supplies, because groundwater

residence times are much longer than the physical

decay time of short lived nuclides. The only

significant transfer of radionuclides to groundwater

has occurred within the CEZ. In some wells during

the past ten years the

137

Cs activity concentration

has declined, while that of

Sr has continued to

increase in shallow groundwaters (Fig. 3.56).

Transfer of radionuclides to groundwater has

occurred from radioactive waste disposal sites in the

0 500 1000 1500 2000

0–0.15

0.30–0.50

0.70–0.90

1.00–1.20

1.40–1.60

1.80–2.00

2.25–2.50

2.75–3.00

Caesium-137 activity (Bq/kg)

Slice (cm)

FIG. 3.55. Caesium-137 profile in the bottom sediment of

the Black Sea (core BS-23/2000), taken during the IAEA

Black Sea expedition in 2000 [3.96].

8 Dec. 1988

8 Dec. 1989

8 Dec. 1990

8 Dec. 1991

7 Dec. 1992

7 Dec. 1993

7 Dec. 1994

7 Dec. 1995

6 Dec. 1996

6 Dec. 1997

6 Dec. 1998

6 Dec. 1999

5 Dec. 2000

Activity concentration (Bq/L)

100.00

10.00

1.00

0.10

0.01

Caesium-137

Strontium-90

FIG. 3.56. Caesium-137 and

Sr in shallow groundwater in

the Red Forest area near the Chernobyl industrial site

[3.139].

CEZ. After the accident, FCM and radioactive

debris were temporarily stored at industrial sites at

the power station and in areas near the floodplain of

the Pripyat River. In addition, trees from the Red

Forest were buried in shallow unlined trenches. At

these waste disposal sites,

Sr activity concentra-

tions in groundwaters are, in some cases, of the

order of 1000 Bq/L [3.140]. The health risks from

groundwater consumption by hypothetical residents

returning to these areas, however, were low in

comparison with external radiation and radiation

doses from the intake of foodstuffs [3.138].

Although there is a potential for off-site

transfer of radionuclides from the disposal sites,

Bugai et al. [3.138] concluded that this will not be

significant in comparison with washout of surface

deposited radioactivity. Studies have shown that

groundwater fluxes of radionuclides are in the

direction of the Pripyat River, but the rate of radio-

nuclide migration is very low and does not present a

significant risk to the Dnieper reservoir system. Off-

site transport of groundwater contamination around

the shelter is also expected to be insignificant,

because radioactivity in the shelter is separated

from groundwaters by an unsaturated zone of 5–6 m

thickness, and groundwater velocities are low

[3.138]. It is predicted that the maximum subsurface

Sr transport rate from waste disposal sites to

surface water bodies will occur from 33 to 145 years

after the accident. The maximum cumulative

transport from all the sources described above is

estimated to be 130 GBq over approximately 100

years, or 0.02%/a of the total inventory within the

contaminated catchments. Integrated radionuclide

transport for a 300 year period is estimated by

Bugai et al. [3.138] to be 15 TBq, or 3% of the total

initial inventory of radioactive material within the

catchments (Fig. 3.57).

The water level of the Chernobyl cooling pond

significantly influences groundwater flows around

the Chernobyl site. Currently, the water level of the

cooling pond is kept artificially high, at 6–7 m above

the average water level in the Pripyat River.

However, this will change when the cooling systems

at the Chernobyl nuclear power plant are finally

shut down and the pumping of water into the pond

is terminated. As the pond dries out, the sediments

will be partly exposed and subject to dispersal.

Recent studies suggest that the best strategy for

remediation of the cooling pond is to allow the

water level to decline naturally, with some limited

action to prevent secondary wind resuspension

using phytoremediation techniques [3.141].

When the water level in the cooling pond

declines to that of the river water surface level, this

will lead to reduction of groundwater fluxes from

the Chernobyl industrial site towards the river. This

will also reduce radionuclide fluxes from the main

radioactive waste disposal sites and the shelter to

the Dnieper cascade. The groundwater fluxes of

from the Chernobyl shelter to the Pripyat River

have been modelled within the framework of

environmental impact assessment studies for the

NSC to be erected above the shelter [3.142] (see

Fig. 3.58). It has been predicted that it would take

approximately 800 years for

Sr to reach the

Pripyat River. With its half-life of 29.1 years, the

activity of

Sr would reduce to an insignificant level

during this time. Thus infiltration of

Sr from the

shelter will not cause harmful impacts on the

Pripyat River. Caesium-137 moves much more

slowly than

Sr, and even after 2000 years its plume

is predicted to be only 200 m from the shelter.

Owing to its high adsorption to the soil matrix,

239

Pu migrates at a much slower rate than

Sr or

137

Cs; however, its half-life is much longer (24 000

100

120

140

0 50 100 150 200 250 300 350

Time (years)

Total groundwater transport of

Sr to the Pripyat River

Strontium-90 (GBq/a)

1. Total transport

2. Catchment of Lake Azbuchin

3. Catchment of Pripyat bay

4. Budbaza temporary waste

disposal/storage site

5. Chernobyl nuclear power plant

industrial site

FIG. 3.57. Prediction of

Sr transport via the groundwate

athway to the Pripyat River in the Chernobyl nuclea

ower plant near zone [3.138].

800 900 1000 1100 1200 1300 1400 1500 1600 1700

110

100

Height above sea level (m)

1x 10

Distance (m)

800 900 1000 1100 1200 1300 1400 1500 1600 1700

Bq/m

FIG. 3.58. Predicted

Sr groundwater concentration

(Bq/m

) in the vicinity of the Chernobyl shelter without the

NSC after 100 years [3.142].

years). The maximum groundwater

239

Pu influx

from the shelter into the Pripyat River is predicted

to be 2 Bq/s. When this influx is fully mixed with the

average Pripyat River discharge of 400 m

/s, the

resulting

239

Pu concentration in the river would be

only 0.005 Bq/m

, as compared with the current

239

Pu level of 0.25 Bq/m

[3.142]. In Ukraine the

regulatory limit for

239

Pu in water is 1 Bq/m

. Thus

infiltration of

239

Pu from the shelter, even without

the NSC, will not cause any significant impact on

the Pripyat River.

3.5.5.2. Radionuclides in groundwater: outside the

Chernobyl exclusion zone

The most detailed current studies of

groundwater contamination in the far zone (beyond

the CEZ) [3.137, 3.143] have concluded that ten

years after the initial surface ground pollution, the

levels of

137

Cs and

Sr in groundwater of the upper

horizons of the aquifer were 40–50 mBq/L around

Kiev and 20–50 mBq/L in the Bryansk region of the

Russian Federation and the majority of the contam-

inated areas in Belarus. In these areas, far from the

Chernobyl reactor (in Belarus and the Russian

Federation), the

137

Cs activity concentration of

water in the saturated zone of soils had a significant

correlation with the

137

Cs soil deposition. In most of

the studied area the activity concentration in

groundwater (per unit of

137

Cs soil deposition) was

significantly lower than in most river and lake

systems. All studies reported that the radionuclide

concentration in the contaminated area outside the

CEZ never exceeded the safety level for

consumption of water and was usually several

orders of magnitude below it.

After fallout from nuclear weapon testing, it

was observed that

Sr in Danish groundwater was

approximately ten times lower than in surface

streams [3.144]. Reference [3.144] also shows that

after the Chernobyl accident, although there were

measurable quantities of

137

Cs in streams, activity

concentrations in groundwater were below

detection limits.

3.5.5.3. Irrigation water

In the Dnieper River basin there is more than

1.8 × 10

ha of irrigated agricultural land. Almost

72% of this territory is irrigated with water from the

Kakhovka reservoir and other Dnieper reservoirs.

Accumulation of radionuclides in plants on

irrigated fields can take place via root uptake of

radionuclides introduced with irrigation water and

due to direct incorporation of radionuclides through

leaves due to sprinkling. However, in the case of the

irrigated lands of southern Ukraine, radionuclides

in irrigation water did not add significant radioac-

tivity to crops in comparison with that which had

been initially deposited in atmospheric fallout and

subsequently taken up in situ from the soil [3.145].

3.5.6. Future trends

3.5.6.1. Freshwater ecosystems

For the rivers and reservoirs of the Dnieper

system, the intensity of runoff of radionuclides will

gradually reduce. In the worst case scenario, hydro-

logical runoff during the next 50 years [3.146] would

cause average concentrations of

137

Cs and

approaching pre-accident levels. Contamination

levels of the water and the main consumer fish

species in the reservoirs of the middle and lower

Dnieper River will approach background levels

(Fig. 3.59). At the same time, in the isolated (closed)

100

200

300

400

500

600

700

800

900

1000

1996 2004 2012 2020 2028 2036 2044 2052

Year

Kiev reservoir

Kakhovka reservoir

Bq/m

(a)

100

200

300

400

500

600

700

800

900

1000

1996 2004 2012 2020 2028 2036 2044 2052

Year

Kiev reservoir

Kakhovka reservoir

Bq/m

(b)

FIG. 3.59. Predicted concentrations of

Sr in water of the upper and downstream reservoirs for the worst (a) and best (b)

probabilistic hydrological scenarios expected for the Pripyat River basin [3.146].

water bodies of the contaminated territories,

increased contents of

137

Cs, both in water and

aquatic biota, will be maintained for several

decades.

Recent data [3.95, 3.147] show that, at present,

137

Cs activity concentrations in surface water and

fish are declining quite slowly. The effective

ecological half-life in water and young fish has

increased from one to four years during the first five

years after the accident to six to 30 years in recent

years. Future contamination levels can be estimated

with the use of an estimated long term decline of

radiocaesium activity concentrations in water and

fish with an effective ecological half-life (T

eff

) of

approximately 20 years, although there is wide

variation in the rates of decline [3.125].

Activity concentrations of radiocaesium in

water are, at present, relatively low (1 Bq/L at

most), except in the shallow closed lakes in the CEZ

and in other highly contaminated areas. Activity

concentrations are expected to continue to decline

slowly during the coming decades. In some lakes,

however,

137

Cs activity concentrations in both water

and fish are expected to remain relatively high for

some decades, as illustrated in Tables 3.9 and 3.10.

Activity concentrations of

Sr in water were also

estimated using a predicted T

eff

of 20 years. This

may, again, be slightly conservative, as long term

TABLE 3.9. CAESIUM-137 ACTIVITY CONCENTRATIONS IN WATER IN VARIOUS CHERNOBYL

AFFECTED LAKES AROUND EUROPE AND PREDICTIONS FOR 30, 50 AND 70 YEARS AFTER

THE ACCIDENT [3.125]

Measured

137

Cs (Bq/L)

(year of measurement)

Predicted

137

Cs (Bq/L)

2016 2036 2056

Lake Kozhanovskoe, Russian Federation 7.0 (2001) 4.2 2.1 1.0

Kiev reservoir, Ukraine 0.028 (1998) 0.015 0.007 0.004

Chernobyl cooling pond 2.5 (2001) 1.5 0.8 0.4

Lake Svyatoe, Belarus

4.7 (1997) 2.4 1.2 0.6

Lake Vorsee, Germany 0.055 (2000) 0.032 0.016 0.008

Devoke Water, UK 0.012 (1998) 0.006 0.003 0.002

This lake had a countermeasure applied in 1998. The prediction is for levels in the absence of countermeasures.

TABLE 3.10. CAESIUM-137 ACTIVITY CONCENTRATIONS (PER UNIT FRESH WEIGHT) IN FISH

IN VARIOUS CHERNOBYL AFFECTED LAKES AROUND EUROPE AND PREDICTIONS FOR 30, 50

AND 70 YEARS AFTER THE ACCIDENT [3.125]

Fish species

Measured

137

(Bq/kg) (year of

measurement)

Predicted

137

Cs (Bq/kg fresh weight)

2016 2036 2056

Lake Kozhanovskoe, Russian

Federation

Goldfish 10 000 (1997) 5 200 2 600 1 300

Kiev reservoir, Ukraine Perch 300 (1997) 160 80 40

Chernobyl cooling pond Perch 18 000 (2001) 11 000 5 400 2 700

Lake Svyatoe, Belarus

Perch 104 000 (1997) 54 000 27 000 14 000

Lake Vorsee, Germany Pike 174 (2000) 100 50 25

Lake Høysjøen, Norway Trout 390 (1998) 210 100 50

Devoke Water, UK Trout 370 (1996–1998) 200 100 50

This lake had a countermeasure applied in 1998. The prediction is for levels in the absence of countermeasures.

rates of decline of

Sr from weapons testing had a

eff

of about ten years [3.148]. Similar to

137

Cs,

activity concentrations of

Sr in water are expected

to decline from their current low levels during the

coming decades (Table 3.11).

Fuel particle breakdown took place at a much

slower rate in lake sediments than in soils [3.149].

The half-life of fuel particles in sediments in the

cooling pond is approximately 30 years [3.39], and

hence radionuclides in fuel particles will remain in

their original form for many years.

3.5.6.2. Marine ecosystems

At present, radionuclides (mainly radio-

caesium) in marine systems are at much lower

concentrations than those observed in freshwater

systems. Activity concentrations in sea water and

marine biota in the Black Sea are expected to

continue to decline (see Table 3.8). This is mainly

due to physical decay, but continued transfers to

seabed sediments and further dilution will also

contribute to the decline.

3.6. CONCLUSIONS

The highest radionuclide deposition occurred

in Belarus, the Russian Federation and Ukraine, but

high depositions also occurred in a number of other

European countries.

Most of the strontium and plutonium radionu-

clides were deposited close to the reactor and were

associated with fuel particles. The environmental

mobility of these radionuclides was lower than that

of the fallout associated with condensed particles,

which predominated in other areas, although the

bioavailability of

Sr has increased with time as the

fuel particles have partially dissolved.

Most of the originally released radionuclides

have disappeared, due to radioactive decay;

137

Cs is

currently of most concern. In the long term future

(more than 100 years) only plutonium isotopes and

241

Am will remain.

The deposition in urban areas in the nearest

city of Pripyat and surrounding settlements could

have initially given rise to a substantial external

dose to the inhabitants, which was averted by the

evacuation measures. The deposition of radioactive

material in other urban areas has provided

substantial contributions to the dose to humans

during the years after the accident and up to the

present.

During the first weeks to months after the

accident, the transfer of short lived radioiodine

isotopes to milk was rapid and high, leading to

substantial doses to humans in the former USSR.

Due to the emergency situation and the short half-

life of

131

I, there are few reliable data on the spatial

distribution of deposited radioiodine. Current

measurements of

129

I may assist in estimating

131

deposition better and thereby improve thyroid dose

reconstruction.

The high concentrations of radioactive

substances in surface water directly after the accident

reduced rapidly, and drinking water and water used

for irrigation have very low concentrations of

radionuclides today.

Due to radioactive decay, rain and wind,

human activities, and countermeasures, surface

contamination in urban areas by radioactive

material has been substantially reduced. External

doses in urban areas, compared with open areas, are

reduced by shielding effects.

TABLE 3.11. STRONTIUM-90 ACTIVITY CONCENTRATIONS IN WATER IN VARIOUS

CHERNOBYL AFFECTED LAKES AND RIVERS AND PREDICTIONS FOR 30, 50 AND 70 YEARS

AFTER THE ACCIDENT [3.125]

Measured

Sr (Bq/L)

(year of measurement)

Predicted

Sr (Bq/L)

2016 2036 2056

Pripyat River 0.28 (1998) 0.15 0.08 0.04

Kiev reservoir 0.16 (1998) 0.09 0.04 0.02

Chernobyl cooling pond 2.0 (2001) 1.2 0.6 0.3

Lake Glubokoye, CEZ 120 (2004) 80–90 40–60 20–30

At present, in most of the settlements

subjected to radioactive contamination, the

radiation dose rate above solid surfaces has

returned to the pre-accident background level.

Elevated dose rates remain mainly in areas of

undisturbed soil.

From the summer of 1986 onwards,

137

Cs was

the dominant radionuclide of concern in agricul-

tural products (milk and meat). During the first few

years, substantial amounts of food were discarded

from human consumption. The highest activity

concentrations of

137

Cs have been found in food

products from forest areas, especially in

mushrooms, berries, game and reindeer. High

137

activity concentrations in fish occur in lakes with

slow or no turnover of water, particularly if the lake

is also shallow and mineral nutrient poor.

The importance of

Sr in food products is

lower than that of

137

Cs because of its lower

deposition and because milk is the only major

animal food product to which it is transferred.

Strontium accumulation in the bones of agricultural

animals and fish occurs but does not typically lead

to radiation doses to humans.

There have been large long term variations in

137

Cs activity concentrations in food products, due

not only to deposition levels but also to differences

in soil types and management practices. In many

areas there are still food products, particularly from

extensive agricultural production systems and

forests, with

137

Cs activity concentrations exceeding

the intervention limits. Large land areas in the

former USSR are still excluded from agricultural

production for radiological reasons.

The major and persistent problems in the

affected areas occur in extensive agricultural

systems with soils with a high organic content and

where animals graze on unimproved pastures. This

particularly affects rural residents in the former

USSR, who are commonly subsistence farmers with

privately owned dairy cows.

In general, there was an initial substantial

reduction in the transfer of

137

Cs to vegetation and

animals, as would be expected, due to weathering,

physical decay, migration of radionuclides down the

soil column and reductions in bioavailability.

However, in the past decade there has been little

further obvious decline, and long term effective

half-lives have been difficult to quantify.

There has been a particularly slow decrease

since deposition in

137

Cs activity concentrations in

some products from forests, and some species of

mushrooms are expected to have high

137

Cs activity

concentrations for decades to come. Under certain

weather and ecological conditions, the biomass of

mushrooms in autumn can be much higher than

normal, leading to relatively high seasonal increases

137

Cs activity concentrations in game. Thus it

must not always be assumed that

137

Cs activity

concentrations in animals will remain as they are

now or decline each year.

Radiocaesium in timber is of minor

importance, although doses in the timber industry

have to be considered. Wood ash can contain higher

amounts of

137

Cs. Forest fires have increased the air

activity concentrations in local areas, but not to a

high extent.

Due to dilution, there were never high

concentrations of

137

Cs in marine fish in the Black

Sea or the Baltic Sea.

3.7. FURTHER MONITORING AND

RESEARCH NEEDED

Updated mapping of

137

Cs deposition in

Albania, Bulgaria and Georgia should be

performed in order to complete the study of post-

Chernobyl contamination of Europe.

Improved mapping of

131

I deposition, based

both on historical environmental measurements

carried out in 1986 and on recent measurements of

129

I in soil samples in areas where elevated thyroid

cancer incidence has been detected after the

Chernobyl accident, would reduce the uncertainty

of the thyroid dose reconstruction needed for deter-

mination of radiation risks.

Long term monitoring of

137

Cs and

activity concentrations in agricultural vegetable and

animal products produced in areas with various soil

and climate conditions and different agricultural

practices should be performed for decades to come

within focused research programmes at selected

sites.

The study of the distribution of

137

Cs and

plutonium radionuclides in the urban environment

(Pripyat, Chernobyl and some other contaminated

towns) in the future would improve the modelling

of human external exposure and inhalation of radio-

nuclides for possible application to any future

nuclear or radiological accident or malicious action.

The continued long term monitoring of

specific forest products such as mushrooms and

game needs to be carried out in those areas in which

forests were significantly contaminated. The results

from such monitoring are being used by the relevant

authorities in affected countries to provide advice to

the general public on the continued use of forests

for recreation and the gathering of wild foods.

In addition to the general monitoring of forest

products, required for radiation protection, more

detailed, scientifically based, long term monitoring

of specific forest sites is required to provide an

ongoing and improved understanding of the long

term dynamics and persistence of radiocaesium

contamination and its variability. Such monitoring is

also necessary to improve the existing predictive

models. Monitoring programmes are being carried

out in several of the more severely affected

countries, such as Belarus and the Russian

Federation, and it is important that these continue

for the foreseeable future if current uncertainties in

long term forecasts are to be reduced.

Aquatic systems have been intensively studied

and monitored during the years after the Chernobyl

accident, and transfers and bioaccumulation of the

most important long term contaminants,

Sr and

137

Cs, are now well understood. There is therefore

little urgent need for major new research

programmes on radionuclides in aquatic systems.

There is, however, a requirement for continued (but

perhaps more limited) monitoring of the aquatic

environment and for further research in some

specific areas, as detailed below.

Predictions of the future contamination of

aquatic systems with

Sr and

137

Cs would be

improved by continued monitoring of radioactivity

in key systems (the Pripyat–Dnieper system, the

seas, and selected rivers and lakes in the more

affected areas and western Europe). This would

continue the time series measurements of activity

concentrations in water, sediments and fish and

enable the refinement of predictive models for

these radionuclides.

Although they are currently of minor radio-

logical importance in comparison with

Sr and

137

Cs, further studies of transuranic elements in the

Chernobyl accident area would improve predictions

of environmental contamination in the very long

term (hundreds to thousands of years). Further

empirical studies of transuranic elements and

are unlikely to have direct implications for radio-

logical protection in the Chernobyl affected areas,

but would further add to our knowledge of the

environmental behaviour of these very long lived

radionuclides.

Future plans to reduce the water level of the

Chernobyl cooling pond will have significant impli-

cations for its ecology and the behaviour of radionu-

clides/fuel particles in newly exposed sediments.

Specific studies on the cooling pond should

therefore continue. In particular, further study of

fuel particle dissolution rates in aquatic systems

such as the cooling pond would improve knowledge

of these processes.

REFERENCES TO SECTION 3

[3.1] INTERNATIONAL NUCLEAR SAFETY

ADVISORY GROUP, Summary Report on the

Post-accident Review Meeting on the Chernobyl

Accident, Safety Series No. 75-INSAG-1, IAEA,

Vienna (1986).

[3.2] IZRAEL, Y.A., et al., Chernobyl: Radioactive

Contamination of the Environment, Gidrometeo-

izdat, Leningrad (1990).

[3.3] IZRAEL, Y.A., Radioactive Fallout After

Nuclear Explosions and Accidents, Elsevier,

Amsterdam (2002).

[3.4] INTERNATIONAL ADVISORY COMMIT-

TEE, The International Chernobyl Project:

Technical Report, IAEA, Vienna (1991).

[3.5] UNITED NATIONS, Sources, Effects and Risks

of Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(1988) 309–374.

[3.6] UNITED NATIONS, Sources and Effects of

Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), Vol. II, UN, New

York (2000) 451–566.

[3.7] DREICER, M., et al., “Consequences of the

Chernobyl accident for the natural and human

environments”, One Decade after Chernobyl:

Summing up the Consequences of the Accident

(Proc. Int. Conf. Vienna, 1996), IAEA, Vienna

(1996) 319–361.

[3.8] MÜCK, K., et al., A consistent radionuclide vector

after the Chernobyl accident, Health Phys. 82

(2002) 141–156.

[3.9] KASHPAROV, V.A., et al., Territory contamina-

tion with the radionuclides representing the fuel

component of Chernobyl fallout, Sci. Total

Environ. 317 (2003) 105–119.

[3.10] SANDALLS, F.J., SEGAL, M.G., VICTOROVA,

N.V., Hot particles from Chernobyl: A review, J.

Environ. Radioact. 18 (1993) 5–22.

[3.11] POLLANEN, R., VALKAMA, I., TOIVONEN,

H., Transport of radioactive particles from the

Chernobyl accident, Atmos. Environ. 31 (1997)

3575–3590.

[3.12] IZRAEL, Y. (Ed.), Atlas of Radioactive Contami-

nation of European Russia, Belarus and Ukraine,

Federal Service for Geodesy and Cartography of

Russia, Moscow (1998).

[3.13] DE CORT, M., Atlas of Caesium Deposition on

Europe after the Chernobyl Accident, Rep. 16733,

Office for Official Publications of the European

Communities, Luxembourg (1998).

[3.14] BOBOVNIKOVA, T.I., MAKHONKO, K.P.,

SIVERINA, A.A., RABOTNOVA, F.A.,

VOLOKITIN, A.A., Physical chemical forms of

radionuclides in atmospheric fallout after the

Chernobyl accident and their transformation in

soil, Atomnaya Energiya, No. 5 (1991) 449–454.

[3.15] HILTON, J., CAMBRAY, R.S., GREEN, N., Frac-

tionation of radioactive caesium in airborne

particles containing bomb fallout, Chernobyl

fallout, and atmospheric material from the Sella-

field site, J. Environ. Radioact. 15 (1992) 103–108.

[3.16] BORZILOV, V.A., KLEPIKOVA, N.V., “Effect of

meteorological conditions and release composi-

tion on radionuclide deposition after the

Chernobyl accident”, The Chernobyl Papers

(MERWIN, S.E., BALONOV, M.I., Eds),

Research Enterprises, Richland, WA (1993)

68.

[3.17] HOLLÄNDER, W., GARGER, E. (Eds),

Contamination of Surfaces by Resuspended

Material, ECP-1, Final Report, Rep. EUR 16527,

Office for Official Publications of the European

Communities, Luxembourg (1996).

[3.18] MAKHONKO, K.P., KOZLOVA, E.G., VOLOK-

ITIN, A.A., Radioiodine accumulation on soil and

reconstruction of doses from iodine exposure on

the territory contaminated after the Chernobyl

accident, Radiat. Risk, No. 7 (1996) 90–142.

[3.19] STRAUME, T., et al., The feasibility of using

129

to reconstruct

131

I deposition from the Chernobyl

reactor accident, Health Phys. 71 (1996) 733–740.

[3.20] BUZULUKOV, Y.P., DOBRYNIN, Y.L.,

“Release of radionuclides during the Chernobyl

accident”, The Chernobyl Papers (MERWIN, S.E.,

BALONOV, M.I., Eds), Research Enterprises,

Richland, WA (1993) 3–21.

[3.21] LOS, I., LIKHTAREV, I., The peculiarities of

urban environmental contamination and assess-

ment of actions aimed at reduction of public

exposure, Int. J. Radiat. Hyg. 1 (1993) 51–59.

[3.22] ANDERSSON, K.G., ROED, J., FOGH, C.L.,

Weathering of radiocaesium contamination on

urban streets, walls and roofs, J. Environ.

Radioact. 62 (2002) 49–60.

[3.23] ROED, J., ANDERSSON, K., personal communi-

cation, 2002.

[3.24] GOLIKOV, V.Y., BALONOV, M.I., JACOB, P.,

External exposure of the population living in areas

of Russia contaminated due to the Chernobyl

accident, Radiatiat. Environ. Biophys. 41 10

(2002) 185–193.

[3.25] ANDERSSON, K.G., ROED, J., JACOB, P.,

MECKBACH, R., “Weathering of Cs-137 on

various surfaces in inhabitated areas and calcu-

lated locations factors”, Deposition of Radio-

nuclides, their Subsequent Relocation in the Envi-

ronment and Resulting Implications, Rep. EUR

16604 EN, Office for Official Publications of the

European Communities, Luxembourg (1995)

56.

[3.26] JACOB, P., MECKBACH, R., “Measurements

after the Chernobyl accident regarding the

exposure of an urban population”, Restoration of

Environments Affected by Residues from Radio-

logical Accidents: Approaches to Decision

Making, IAEA-TECDOC-1131, IAEA, Vienna

(2000) 34–41.

[3.27] SHESTO PA L OV, V. M. , KASHPAROV, V. A .,

IVANOV, Y.A., Radionuclide migration into the

geological environment and biota after the

Chernobyl accident, Environ. Sci. Pollut. Res.

Special issue 1 (2003) 39–47.

[3.28] HOWARD, B.J., The concept of radioecological

sensitivity, Radiat. Prot. Dosim. 92 (2000) 29–34.

[3.29] HOWARD, B.J., et al., Estimation of radioecolog-

ical sensitivity, Radioprotection-Colloques 37 C1

(2002) 1167–1173.

[3.30] BALONOV, M.I., et al., Methodology of internal

dose reconstruction for Russian population after

the Chernobyl accident, Radiat. Prot. Dosim. 92

(2000) 247–253.

[3.31] BERESFORD, N.A., et al., The importance of

source dependent bioavailability in determining

the transfer of ingested radionuclides to ruminant

derived food products, Environ. Sci. Technol. 34

(2000) 4455–4462.

[3.32] PRÖHL, G., HOFFMAN, F.O., “Radionuclide

interception and loss processes in vegetation”,

Modelling of Radionuclide Interception and Loss

Processes in Vegetation and of Transfer in Semi-

natural Ecosystems, IAEA-TECDOC-857, IAEA,

Vienna (1996) 9–48.

[3.33] KIRCHNER, G., Transport of iodine and caesium

via the grass–cow–milk pathway after the

Chernobyl accident, Health Phys. 66 (1994)

653

665.

[3.34] RENAUD, P., et al., Les Retombées en France de

l’accident de Tchernobyl, Conséquences

Radioécologiques et Dosimétriques, EDP

Sciences, Les Ulis (1999).

[3.35] HOWARD, B.J., BERESFORD, N.A., HOVE, K.,

Transfer of radiocaesium to ruminants in unim-

proved natural and semi-natural ecosystems and

appropriate countermeasures, Health Phys. 61

(1991) 715–725.

[3.36] BERESFORD, N.A., The transfer of Ag-110m to

sheep tissues, Sci. Total Environ. 85 (1989) 81–90.

[3.37] ALEXAKHIN, R.A., KORNEEV, N.A. (Eds),

Agricultural Radioecology, Ecology, Moscow

(1991) (in Russian).

[3.38] DESMET, G.M., VAN LOON, L.R., HOWARD,

B.J., Chemical speciation and bioavailability of

elements in the environment and their relevance

to radioecology, Sci. Total Environ. 100 (1991)

105–124.

[3.39] KASHPAROV, V.A., et al., Kinetics of dissolution

of Chernobyl fuel particles in soil in natural condi-

tions, J. Environ. Radioact. 72 (2004) 335–353.

[3.40] SALBU, B., et al., High energy X-ray microscopy

for characterization of fuel particles, Nucl.

Instrum. Methods Phys. Res. A 467–468 (2001)

1249–1252.

[3.41] FESENKO, S.V., SANZHAROVA, N.I., SPIRI-

DONOV, S.I., ALEXAKHIN, R.M., Dynamics of

137

Cs bioavailability in a soil–plant system in areas

of the Chernobyl nuclear power plant accident

zone with a different physico-chemical composi-

tion of radioactive fallout, J. Environ. Radioact. 34

(1996) 287–313.

[3.42] GADD, G.M., Role of microorganisms in the

environmental fate of radionuclides, Endeavour

20 (1996) 150–156.

[3.43] TAMPONNET, C., PLASSARD, C., PAREKH,

N., SANCHEZ, S., “Impact of microorganisms on

the fate of radionuclides in rhizospheric soils”,

Radioactive Pollutants — Impact on the Environ-

ment (BRÉCHIGNAC, F., HOWARD, B., Eds),

EDP Sciences, Les Ulis (2001) 175–185.

[3.44] SANZHAROVA, N.I., et al., Changes in the

forms of

137

Cs and its availability for plants as

dependent on properties of fallout after the

Chernobyl nuclear power plant accident, Sci. Total

Environ. 154 (1994) 9–22.

[3.45] FESENKO, S.V., COLGAN, P.A., LISSIANSKI,

K.B., VAZQUEZ, C., GUARDANS, R., The

dynamics of the transfer of caesium-137 to animal

fodder in areas of Russia affected by the

Chernobyl accident and doses resulting from the

consumption of milk and milk products, Radiat.

Prot. Dosim. 69 (1997) 289–299.

[3.46] SHEVCHOUK, V.E., GOURACHEVSKIY, V.L.

(Eds), 15 Years After the Chernobyl Catastrophe:

Consequences in the Republic of Belarus and

their Overcoming, National Report, Committee

on the Problems of the Consequences of the

Accident at the Chernobyl NPP, Minsk (2001).

[3.47] KASHPAROV, V.A., OUGHTON, D.H.,

ZVARICH, S.I., PROTSAK, V.P., LEVCHUK,

S.E., Kinetics of fuel particle weathering and

mobility in the Chernobyl 30 km exclusion zone,

Health Phys. 76 (1999) 251–259.

[3.48] SANZHAROVA, N.I., FESENKO, S.V., KOTIK,

V.A., SPIRIDONOV, S.I., Behaviour of radionu-

clides in meadows and efficiency of countermeas-

ures, Radiat. Prot. Dosim. 64 (1996) 43–48.

[3.49] BOGDEVITCH, I.M. (Ed.), Guide for Agricul-

tural Practice on Lands Contaminated by Radio-

nuclides in the Republic of Belarus for 2002–2005,

Ministry of Agriculture and Food, Minsk (2002)

(in Russian).

[3.50] EHLKEN, S., KIRCHNER, G., Environmental

processes affecting plant root uptake of radioac-

tive trace elements and variability of transfer

factor data: A review, J. Environ. Radioact. 58

(2002) 97–112.

[3.51] SAURAS YERA, T., et al.,

137

Cs and

Sr root

uptake predictions under close-to-real controlled

conditions, J. Environ. Radioact. 45 (1999)

191

217.

[3.52] KONOPLEV, A.V., et al., Quantitative assess-

ment of radiocaesium bioavailability in forest soils,

Radiochim. Acta 88 (2000) 789–792.

[3.53] KASHPAROV, V. A . , et al., Complex monitoring

of agricultural products contamination with

Sr,

Bull. Agrarian Sci. Special issue (2001) 38–43 (in

Ukrainian).

[3.54] DEVILLE-CAVELIN, G., et al., Synthesis Report

of the Radioecology Project of the French–

German Initiative for Chernobyl, FGI Report 04-

01, Institute for Radiological Protection and

Nuclear Safety, Paris (2004).

[3.55] FESENKO, S.V., et al., Regularities of change in

137

Cs activity concentrations in the long term after

the accident at the Chernobyl NPP, Radiat. Biol.

Radioecol. 44 (2004) 35–49.

[3.56] SHUTOV, V., et al., personal communication,

2004.

[3.57] BRUK, G.Y., SHUTOV, V.N., BALONOV, M.I.,

BASALEYAVA, L.N., KISLOV, M.V., Dynamics

137

Cs content in agricultural food products

produced in regions of Russia contaminated after

the Chernobyl accident, Radiat. Prot. Dosim. 76

(1998) 169–178.

[3.58] PRISTER, B.S., et al., Experimental substantia-

tion in parametrization of the model describing

137

Cs and

Sr behavior in a soil–plant system,

Environ. Sci. Pollut. Res. Special issue 1 (2003)

126–136.

[3.59] BERESFORD, N.A., CROUT, N.M.J., MAYES,

R.W., HOWARD, B.J., LAMB, C.S., Dynamic

distribution of radioisotopes of cerium, ruthenium

and silver in sheep tissues, J. Environ. Radioact. 38

(1998) 317–338.

[3.60] CROUT, N.M.J., BERESFORD, N.A., DAWSON,

J., SOAR, J., MAYES, R.W., The transfer of

As,

109

Cd and

203

Hg to the milk and tissues of dairy

cattle, J. Agr. Sci. Camb. 142 (2004) 203–212.

[3.61] BOGDEVITCH, I.M., “Soil conditions of Belarus

and efficiency of potassium fertilizers”, Proc. 16th

World Congr. Soil Science, Montpellier, 1998,

International Potash Institute, Basel (1999) 21–26.

[3.62] PRISTER, B. (Ed.), Recommendations on Agri-

culture Management on Contaminated Territories,

Ukrainian Institute of Agricultural Radiology,

Kiev (1998) (in Ukrainian).

[3.63] BEBESHKO, V.G., et al., General Dosimetric

Passportization of Settlements in Ukraine Radio-

actively Contaminated after the Chernobyl

Accident: Summarized Data for 1998–2000, Issue

9, Ministry of Emergencies, Kiev (2001) (in

Russian).

[3.64] CODEX ALIMENTARIUS COMMISSION,

Guideline Levels for Radionuclides in Foods

Following Accidental Nuclear Contamination for

Use in International Trade, Rep. CAC/GL 5-1989,

Codex Alimentarius Commission, Rome (1989).

[3.65] FOOD AND AGRICULTURE ORGANIZA-

TION OF THE UNITED NATIONS, INTER-

NATIONAL ATOMIC ENERGY AGENCY,

INTERNATIONAL LABOUR ORGANISA-

TION, OECD NUCLEAR ENERGY AGENCY,

PAN AMERICAN HEALTH ORGANIZA-

TION, WORLD HEALTH ORGANIZATION,

International Basic Safety Standards for Protec-

tion against Ionizing Radiation and for the Safety

of Radiation Sources, Safety Series No. 115,

IAEA, Vienna (1996).

[3.66] TIKHOMIROV, F.A., SHCHEGLOV, A.I., Main

investigation results on the forest radioecology in

the Kyshtym and Chernobyl accident zones, Sci.

Total Environ. 157 (1994) 45–57.

[3.67] NYLÉN, T., Uptake, Turnover and Transport of

Radiocaesium in Boreal Forest Ecosystems,

Thesis, Swedish Univ. Agricultural Sciences,

Uppsala (1996).

[3.68] INTERNATIONAL ATOMIC ENERGY

AGENCY, Modelling the Migration and Accumu-

lation of Radionuclides in Forest Ecosystems,

IAEA-BIOMASS-1, IAEA, Vienna (2002).

[3.69] SCHUCK, A., et al., Compilation of a European

forest map from Portugal to the Ural mountains

based on earth observation data and forest statis-

tics, Forest Policy Econ. 5 (2003) 187–202.

[3.70] SHAW, G., AVILA, R., FESENKO, S.,

DVORNIK, A., ZHUCHENKO, T., “Modelling

the behaviour of radiocaesium in forest ecosys-

tems”, Modelling Radioactivity in the Environ-

ment (SCOTT, E.M., Ed.), Elsevier, Amsterdam

(2002).

[3.71] JOHANSSON, K., “Radiocaesium in game

animals in the Nordic countries”, Nordic Radio-

ecology: The Transfer of Radionuclides through

Nordic Ecosystems to Man (DAHLGAARD, H.,

Ed.), Elsevier, Amsterdam (1994) 287–301.

[3.72] SHCHEGLOV, A.I., TSVETNOVA, O.B.,

KLYASHTORIN, A.L., Biogeochemical

Migration of Technogenic Radionuclides in Forest

Ecosystems, Nauka, Moscow (2001).

[3.73] SHAW, G., personal communication, 2004.

[3.74] DVORNIK, A.M., ZHUCHENKO, T., personal

communication, 2004.

[3.75] BELLI, M., TIKHOMIROV, F. (Eds), Behaviour

of Radionuclides in Natural and Semi-natural

environments: Experimental Collaboration

Project No. 5, Final Report, Rep. EUR-16531 EN,

Office for Official Publications of the European

Communities, Luxembourg (1996).

[3.76] FESENKO, S.V., et al., Identification of processes

governing long-term accumulation of

137

Cs by

forest trees following the Chernobyl accident,

Radiat. Environ. Biophys. 40 (2001) 105–113.

[3.77] PANFILOV, A., “Countermeasures for radioac-

tively contaminated forests in the Russian Federa-

tion”, Contaminated Forests: Recent

Developments in Risk Identification and Future

Perspectives (LINKOV, I., SCHELL, W.R., Eds),

NATO Science Series, Vol. 58, Kluwer, Dordrecht

(1999) 271–279.

[3.78] OULD-DADA, Z., BAGHINI, N.M., Resuspen-

sion of small particles from tree surfaces, Atmos.

Environ. 35 (2001) 3799–3809.

[3.79] AMIRO, B.D., DVORNIK, A., “Fire and radioac-

tivity in contaminated forests”, Contaminated

Forests: Recent Developments in Risk Identifica-

tion and Future Perspectives (LINKOV, I.,

SCHELL, W.R., Eds), NATO Science Series,

Vol. 58, Kluwer, Dordrecht (1999) 311–324.

[3.80] RUHM, W., KAMMERER, L., HIERSCHE, L.,

WIRTH, E., The

137

Cs/

134

Cs ratio in fungi as an

indicator of the major mycelium location in forest

soil, J. Environ. Radioact. 35 (1997) 129–148.

[3.81] INTERNATIONAL ATOMIC ENERGY

AGENCY, Handbook of Parameter Values for the

Prediction of Radionuclide Transfer in Temperate

Environments, Technical Reports Series No. 364,

IAEA, Vienna (1994).

[3.82] BARNETT, C.L., et al., Radiocaesium intake in

Great Britain as a consequence of the consump-

tion of wild fungi, Mycologist 15 (2001) 98–104.

[3.83] JOHANSSON, K., personal communication, 2003.

[3.84] ZIBOLD, G., personal communication, 2004.

[3.85] INTERNATIONAL ATOMIC ENERGY

AGENCY, Assessing Radiation Doses to the

Public from Radionuclides in Timber and Wood

Products, IAEA-TECDOC-1376, IAEA, Vienna

(2003).

[3.86] RAVILA, A., HOLM, E., Radioactive elements in

the forest industry, Sci. Total Environ. 32 (1994)

339–356.

[3.87] AVILA, R., et al., “Conceptual overview of

FORESTLAND — A model to interpret and

predict temporal and spatial patterns of radioac-

tively contaminated forest landscapes”, Contami-

nated Forests: Recent Developments in Risk

Identification and Future Perspectives (LINKOV,

I., SCHELL, W.R., Eds), NATO Science Series,

Vol. 58, Kluwer, Dordrecht (1999) 173–183.

[3.88] SHAW, G., et al., Radionuclide migration in forest

ecosystems — Results of a model validation study,

J. Environ. Radioact. 84 (2005) 285–296.

[3.89] KRYSHEV, I.I., RYABOV, I.N., “About the effi-

ciency of trophic levels in the accumulation of

Cs-137 in fish of the Chernobyl NPP cooling

pond”, Biological and Radioecological Aspects of

the Consequences of the Chernobyl Accident

(RYABOV, I.N., RYABTSEV, I.A., Eds), USSR

Academy of Sciences, Moscow (1990) 116–121.

[3.90] HÅKANSON, L., ANDERSSON, T., NILSSON,

Ý., Radioactive caesium in fish from Swedish

lakes in 1986–1988 — General pattern related to

fallout and lake characteristics, J. Environ.

Radioact. 15 (1992) 207–229.

[3.91] KRYSHEV, I.I., Radioactive contamination of

aquatic ecosystems following the Chernobyl

accident, J. Environ. Radioact. 27 (1995) 207–219.

[3.92] RYABOV, I., et al., “Radiological phenomena of

the Kojanovskoe Lake”, The Radiological Conse-

quences of the Chernobyl Accident (Proc. Int.

Conf. Minsk, 1996), Rep. EUR 16544, Office for

Official Publications of the European Communi-

ties, Luxembourg (1996) 213–216.

[3.93] ZIBOLD, G., KAMINSKI, S., KLEMT, E.,

SMITH, J.T., Time-dependency of the

137

activity concentration in freshwater lakes, meas-

urement and prediction, Radioprotection-

Colloques 37 (2002) 75–80.

[3.94] SMITH, J.T., KUDELSKY, A.V., RYABOV, I.N.,

HADDERINGH, R.H., Radiocaesium concentra-

tion factors of Chernobyl-contaminated fish: A

study of the influence of potassium, and “blind”

testing of a previously developed model, J.

Environ. Radioact. 48 (2000) 359–369.

[3.95] JONSSON, B., FORSETH, T., UGEDAL, O.,

Chernobyl radioactivity persists in fish, Nature 400

(1999) 417.

[3.96] INTERNATIONAL ATOMIC ENERGY

AGENCY, Marine Environment Assessment of

the Black Sea: Final Report, Technical Coopera-

tion Project RER/2/003, IAEA, Vienna (2003).

[3.97] VOITSEKHOVITCH, O.V., KANIVETS, V.V.,

Maps of Cs-137 in the Bottom Sediments of the

Dnieper Reservoirs, Ukrainian Hydrometeorolog-

ical Institute, Kiev (1997).

[3.98] COMANS, R.N.J., et al., Modelling Fluxes and

Bioavailability of Radiocaesium and Radiostron-

tium in Freshwaters in Support of a Theoretical

Basis for Chemical/Hydrological Countermeas-

ures, Final Report to the European Commission,

Netherlands Energy Research, Petten (1999).

[3.99] VOITSEKHOVITCH, O.V., BORZILOV, V.A.,

KONOPLEV, A.V., “Hydrological aspects of

radionuclide migration in water bodies following

the Chernobyl accident”, Comparative Assess-

ment of the Environmental Impact of Radio-

nuclides Released during Three Major Nuclear

Accidents: Kyshtym, Windscale, Chernobyl, Rep.

EUR 13574, Office for Official Publications of the

European Communities, Luxembourg (1991) 528–

548.

[3.100] MATSUNAGA, T., et al., Characteristics of

Chernobyl-derived radionuclides in particulate

form in surface waters in the exclusion zone

around the Chernobyl nuclear power plants, J.

Contam. Hydrol. 35 (1998) 101–113.

[3.101] MATSUNAGA, T., et al., Association of

dissolved radionuclides released by the Chernobyl

accident with colloidal materials in surface water,

App. Geochem. 19 (2004) 1581–1599.

[3.102] CARLSON, L., HOLM, E., Radioactivity in

Fucus vesiculosus L. from the Baltic Sea following

the Chernobyl accident, J. Environ. Radioact. 15

(1992) 231–248.

[3.103] KNAPINSKA-SKIBA, D., BOJANOWSKI, R.,

RADECKI, Z., MILLWARD, G.E., Activity

concentrations and fluxes of radiocaesium in the

southern Baltic Sea, Estuar. Coast. Shelf Sci. 53

(2001) 779–786.

[3.104] VAKULOVSKY, S.M., et al., “Radioactive

contamination of water systems in the area

affected by releases from the Chernobyl nuclear

power plant accident”, Environmental Contami-

nation Following a Major Nuclear Accident (Proc.

Symp. Vienna, 1989), Vol. 1, IAEA, Vienna (1990)

231–246.

[3.105] VAKULOVSKY, S.M., et al., Cs-137 and Sr-90

contamination of water bodies in the areas

affected by releases from the Chernobyl nuclear

power plant accident: An overview, J. Environ.

Radioact. 23 (1994) 103–122.

[3.106] UKRAINIAN HYDROMETEOROLOGICAL

INSTITUTE, Database of Radiation Measure-

ments of Aquatic Samples, Ukrainian Hydromete-

orological Institute, Kiev (2004).

[3.107] INTERNATIONAL ATOMIC ENERGY

AGENCY, Radiation Conditions in the Dnieper

River Basin, IAEA, Vienna (2006).

[3.108] MONTE, L., Evaluation of radionuclide transfer

functions from drainage basins of freshwater

systems, J. Environ. Radioact. 26 (1995) 71–82.

[3.109] SAXÈN, R., ILUS, E., Discharge of

137

Cs and

by Finnish rivers to the Baltic Sea in 1986–1996, J.

Environ. Radioact. 54 (2001) 275–291.

[3.110] KUDELSKY, A.V., SMITH, J.T., ZHUKOVA,

O.M., MATVEYENKO, I.I., PINCHUK, T.M.,

Contribution of river runoff to the natural remedi-

ation of contaminated territories (Belarus), Proc.

Acad. Sci. Belarus 42 (1998) 90–94 (in Russian).

[3.111] SMITH, J.T., et al., Global analysis of the riverine

transport of

Sr and

137

Cs, Environ. Sci. Technol.

38 (2004) 850–857.

[3.112] HILTON, J., LIVENS, F.R., SPEZZANO, P.,

LEONARD, D.R.P., Retention of radioactive

caesium by different soils in the catchment of a

small lake, Sci. Total Environ. 129 (1993) 253–266.

[3.113] KUDELSKY, A.V., SMITH, J.T., OVSIAN-

NIKOVA, S.V., HILTON, J., The mobility of

Chernobyl-derived

137

Cs in a peatbog system

within the catchment of the Pripyat River, Belarus,

Sci. Total Environ. 188 (1996) 101–113.

[3.114] SANTSCHI, P.H., BOLLHALDER, S., ZINGY,

S., LUCK, A., FARRENKOTHER, K., The self

cleaning capacity of surface waters after radionu-

clide fallout: Evidence from European lakes after

Chernobyl 1986–88, Environ. Sci. Technol. 24

(1990) 519–527.

[3.115] COMANS, R.N.J., et al., Mobilization of radiocae-

sium in pore water of lake sediments, Nature 339

(1989) 367–369.

[3.116] BULGAKOV, A.A., et al., Modelling the long-

term dynamics of radiocaesium in a closed lake

system, J. Environ. Radioact. 61 (2002) 41–53.

[3.117] KUZMENKO, M.I., et al., Impact of Radionu-

clide Contamination on Hydrobionts of the

Chernobyl Exclusion Zone, Chernobylinterin-

form, Kiev (2001) (in Russian).

[3.118] KONOPLEV, A.V., et al., Study of the behavior of

Cs-137 and Sr-90 in Svyatoe and Kozhanovskoe

lakes in the Bryansk region, Meteorologiyai

Gidrologiya 11 (1998) 78–87 (in Russian).

[3.119] MINISTRY OF HEALTH, Radiation Safety

Standards (RSS-99), Sanitary Rules SR 2.6.1.758-

99, Ministry of Health of the Russian Federation,

Moscow (1999).

[3.120] BUCKLEY, M.J., et al., Drawing Up and Evalu-

ating Remediation Strategies for the Chernobyl

Cooling Pond, Final Report, Rep.

C6476/TR/001/2002, NNC, Knutsford, UK (2002).

[3.121] NASVIT, O.I., “Radioecological situation in the

cooling pond of Chornobyl NPP”, Recent

Research Activity about Chernobyl NPP Accident

in Belarus, Ukraine and Russia, Rep. KURAI-

KR-79, Kyoto University (2002) 74–85.

[3.122] VOITSEKHOVITCH, O.V., NASVIT, O., LOS,

I., BERKOVSKI, V., “Present thoughts on aquatic

countermeasures applied to regions of the

Dnieper river catchment contaminated by the

1986 Chernobyl accident”, Freshwater and

Estuarine Radioecology (DESMET, G., et al.,

Eds), Elsevier, Amsterdam (1997) 75–86.

[3.123] KONOPLEV, A.V., et al., Validation of models of

radionuclide wash-off from contaminated water-

sheds using Chernobyl data, J. Environ. Res. 42

(1999) 131–141.

[3.124] SMITH, J.T., et al., Temporal change in fallout

137

Cs in terrestrial and aquatic systems: A whole

ecosystem approach, Environ. Sci. Technol. 33

(1999) 49–54.

[3.125] SMITH, J.T., VOITSEKHOVITCH, O.V.,

HÅKANSON, L., HILTON, J., A critical review

of measures to reduce radioactive doses from

drinking water and consumption of freshwater

foodstuffs, J. Environ. Radioact. 56 (2001) 11–32.

[3.126] SANSONE, U., VOITSEKHOVITCH, O.V.,

Modelling and Study of the Mechanisms of the

Transfer of Radionuclides from the Terrestrial

Ecosystem to and in Water Bodies Around Cher-

nobyl, Rep. EUR 16529 EN, Office for Official

Publications of the European Communities,

Luxembourg (1996).

[3.127] ELLIOTT, J.M., et al., Sources of variation in

post-Chernobyl radiocaesium in fish from two

Cumbrian lakes (north-west England), J. Appl.

Ecol. 29 (1992) 108–119.

[3.128] FLEISHMAN, D.G., “Radioecology of marine

plants and animals”, Radioecology (KLECHKOV-

SKY, V.M., POLIKARPOV, G.G., ALEKSA-

KHIN, R.M., Eds), Wiley, New York (1973) 347–

370.

[3.129] CAMPLIN, W.C., LEONARD, D.R.P., TIPPLE,

J.R., DUCKETT, L., Radioactivity in Freshwater

Systems in Cumbria (UK) Following the

Chernobyl Accident, MAFF Fisheries Research

Data Report No. 18, Ministry of Agriculture,

Fisheries and Food, London (1989).

[3.130] BLAYLOCK, B.G., Radionuclide data bases

available for bioaccumulation factors for fresh-

water biota, Nucl. Saf. 23 (1982) 427–438.

[3.131] HADDERINGH, R.H., VAN AERSSEN,

G.H.F.M., RYABOV, I.N., KOULIKOV, O.A.,

BELOVA, N., “Contamination of fish with

137

in Kiev reservoir and old river bed of Pripyat near

Chernobyl”, Freshwater and Estuarine Radio-

ecology (DESMET, G., et al., Eds), Elsevier,

Amsterdam (1997) 339–351.

[3.132] KRYSHEV, I.I., SAZYKINA, T.G., Accumula-

tion factors and biogeochemical aspects of

migration of radionuclides in aquatic ecosystems

in the areas impacted by the Chernobyl accident,

Radiochim. Acta 66–67 (1994) 381–384.

[3.133] EREMEEV, V.N., IVANOV, L.M., KIRWAN,

A.D., MARGOLINA, T.M., Amount of

137

Cs and

134

Cs radionuclides in the Black Sea produced by

the Chernobyl accident, J. Environ. Radioact. 27

(1995) 49–63.

[3.134] KANIVETS, V.V., VOITSEKHOVITCH, O.V.,

SIMOV, V.G., GOLUBEVA, Z.A., The post-

Chernobyl budget of

137

Cs and

Sr in the Black

Sea, J. Environ. Radioact. 43 (1999) 121–135.

[3.135] NIELSEN, S.P., et al., The radiological exposure

of man from radioactivity in the Baltic Sea, Sci.

Total Environ. 237–238 (1999) 133–141.

[3.136] EUROPEAN UNION, The Radiological

Exposure of the Population of the European

Community to Radioactivity in the Baltic Sea,

Rep. EUR 19200 EN, Office for Official Publica-

tions of the European Communities, Luxembourg

(2000).

[3.137] SHESTOPALOV, V.M. (Ed.), Chernobyl

Disaster and Groundwater, Balkema, Leiden

(2002).

[3.138] BUGAI, D.A., WATERS, R.D., DZHEPO, S.P.,

SKALSKY, A.S., Risks from radionuclide

migration to groundwater in the Chernobyl 30-km

zone, Health Phys. 71 (1996) 9–18.

[3.139] VOITSEKOVITCH, O.V., SHETOPALOV, V.M.,

SKALSKY, A.S., KAIVETS, V.V., Monitoring of

Radioactive Contamination of Surface Ground

Water Following the Chernobyl Accident,

Ukrainian Hydrometeorological Institute, Kiev

(2001) (in Russian).

[3.140] VOITSEKHOVITCH, O.V., SANSONE, U.,

ZHELEZNYAK, M., BUGAI, D., “Water quality

management of contaminated areas and its effects

on doses from aquatic pathways”, The Radiolog-

ical Consequences of the Chernobyl Accident

(Proc. Int. Conf. Minsk, 1996), Rep. EUR 16544

EN, Office for Official Publications of the

European Communities, Luxembourg (1996) 401–

410.

[3.141] VANDENHOVE, H. (Ed.), PHYTOR, Evalua-

tion of Willow Plantation for the Phytorehabilita-

tion of Contaminated Arable Lands and Flood

Plane Areas, Final Report, Rep. SCK•CEN ERB

IC15-CT98 0213, SCK•CEN, Mol (2002).

[3.142] BECHTEL INTERNATIONAL SYSTEMS,

ELECTRICITÉ DE FRANCE, BATTELLE

MEMORIAL INSTITUTE, Environmental

Impact Assessment: New Safe Confinement

Conceptual Design: Chernobyl Nuclear Power

Plant Unit 4, Bechtel, EDF, Battelle, San

Francisco, CA (2003).

[3.143] UNITED NATIONS DEVELOPMENT

PROGRAMME, Water Evaluation and Predic-

tion in the Area Affected by the Chernobyl

Accident (the Bryansk Region), Final Report of

Project RUS/95/004, UNDP, Moscow (2001).

[3.144] HANSEN, H.J.M., AARKROG, A., A different

surface geology in Denmark, the Faroe Islands

and Greenland influences the radiological

contamination of drinking water, Water Res. 24

(1990) 1137–1141.

[3.145] FOOD AND AGRICULTURE ORGANIZA-

TION OF THE UNITED NATIONS, INTER-

NATIONAL ATOMIC ENERGY AGENCY,

Guidelines for Agricultural Countermeasures

Following an Accidental Release of Radio-

nuclides, Technical Reports Series No. 363, IAEA,

Vienna (1994).

[3.146] ZHELEZNYAK, M., SHEPELEVA, V., SIZO-

NENKO, V., MEZHUEVA, I., Simulation of

countermeasures to diminish radionuclide fluxes

from the Chernobyl zone via aquatic pathways,

Radiat. Prot. Dosim. 73 (1997) 181–186.

[3.147] SMITH, J.T., et al., Chernobyl’s legacy in food and

water, Nature 405 (2000) 141.

[3.148] CROSS, M.A., SMITH, J.T., SAXÉN, R., TIMMS,

D.N., An analysis of the time dependent environ-

mental mobility of radiostrontium in Finnish river

catchments, J. Environ. Radioact. 60 (2002) 149–

163.

[3.149] KONOPLEV, A.V., et al., “Physico-chemical and

hydraulic mechanisms of radionuclide mobiliza-

tion in aquatic systems”, The Radiological Conse-

quences of the Chernobyl Accident (Proc. Int.

Conf. Minsk, 1996), Rep. EUR 16544 EN, Office

for Official Publications of the European Commu-

nities, Luxembourg (1996) 121–135.

[3.150] FLEISHMAN, D.G., NIKIFOROV, V.A.,

SAULUS, A.A., KOMOV, V.T.,

137

Cs in fish of

some lakes and rivers of the Bryansk region and

north-west Russia in 1990–1992, J. Environ.

Radioact. 24 (1994) 145–158.

4. ENVIRONMENTAL COUNTERMEASURES

AND REMEDIATION

The need for the application of urgent

protective actions became evident very soon after

the Chernobyl accident occurred. A wide range of

countermeasures was applied for protecting the

public from radiation, from urgent evacuation in

1986 of the inhabitants from the area of highest

radioactive contamination to long term monitoring

of radionuclides in foodstuffs in many European

countries. The whole spectrum of the applied

countermeasures and their effectiveness have been

considered in a number of international reports

[4.1–4.7].

The main subject of this section is the counter-

measures that have been applied to the

environment in order to reduce the radiological

impact on humans. At the time of the Chernobyl

accident the philosophy of radiation protection of

non-human species had not been sufficiently

developed to be practically applied for the purposes

of justifying appropriate countermeasures. Such

policies are currently still under development [4.8].

This section does not consider specifically the

emergency and mitigatory actions applied to the

damaged reactor aimed at reducing radioactive

releases to the environment; these aspects have

been covered elsewhere [4.2].

Environmental countermeasures have been

applied since 1986 to urban, agricultural, forest and

aquatic ecosystems. Most of these countermeasures

were driven by relevant international and national

radiological criteria.

4.1. RADIOLOGICAL CRITERIA

Countermeasures, termed protective actions

at the emergency stage and remedial actions at the

post-emergency stage, are actions taken to reduce

the level of exposure as much as is reasonably

achievable. A fundamental aspect of radiation

protection philosophy is to optimize the dose

averted against the costs of applying the counter-

measure. However, the costs and benefits of

countermeasures are not always quantifiable in

purely monetary terms. The advantages of counter-

measures often include reassurance and a decrease

in anxiety in the affected population. However,

countermeasures may also have negative conse-

quences, either directly to ecosystems (e.g.

disruption of nutrient cycles) or to sectors of the

population either economically or due to disruption

of normal life.

4.1.1. International radiological criteria and

standards

At the time of the Chernobyl accident in 1986,

the relevant international radiation protection

standards for protection of the public and workers

were contained in International Commission on

Radiological Protection (ICRP) Publication 26

[4.9]. Specific recommendations on the protection

of the public in the event of a major radiation

accident were given in ICRP Publication 40 [4.10].

The corresponding IAEA Basic Safety Standards,

based on ICRP recommendations, were issued in

1982 [4.11]. The basic principles of modern

radiation protection — justification, optimization

and dose limitation — and the clear distinction

between protection in normal and intervention

situations were contained in these documents. At

that time, the annual limit for occupational

exposure was equal to 50 mSv and that for public

exposure was 5 mSv. The latter value was perceived

as a safe level of human exposure.

Special limits for public radiation protection

in the event of nuclear or radiological emergencies

were not specifically established in these

documents, and instead it was recommended:

(a) By almost all means to reduce human

accidental exposure below doses that may

result in deterministic health effects (acute

radiation syndrome, radiation damage to

particular organs or tissues);

(b) To intervene (i.e. to apply and subsequently

withdraw countermeasures aimed at reducing

stochastic health effects (cancer, genetic

anomalies)) based on an optimization

assessment taking into account both the

collective dose reduction achieved by the

application of the countermeasures and the

associated economic and social intervention

costs.

The most relevant ICRP guidance [4.10]

recommended some generic two level criteria for

intervention in the early accident phase — for

sheltering, 5–50 mSv of whole body dose or 50–

500 mSv to particular organs; for administration of

stable iodine aimed at thyroid protection against

intake of radioiodines, 50–500 mSv to the thyroid;

for evacuation, 50–500 mSv of whole body dose or

500–5000 mSv to particular organs. For the interme-

diate accident phase, the generic criteria of 5–

50 mSv of whole body dose or 50–500 mSv to

particular organs were recommended for control of

foodstuff contamination with radionuclides, and 50–

500 mSv of whole body dose for relocation.

Afterwards, in connection with public

concerns over the radiological consequences of the

Chernobyl accident, new additional international

regulations were developed. Thus in 1989 the Codex

Alimentarius Commission approved guidance

levels for radionuclides in food moving in interna-

tional trade for the first year after a major nuclear

accident (see Table 4.1) [4.12].

New international basic radiation protection

standards for the protection of the public and

workers were developed by the ICRP in 1990 after

research data had shown that radiation risk coeffi-

cients for stochastic human health effects were

substantially higher than previously thought. The

annual limits of exposure were substantially (by a

factor of 2.5–5) reduced and established equal to

20 mSv for workers and 1 mSv for members of the

general public [4.13]. The latter value is currently

perceived as a safe level of human exposure.

Special limits for public protection in the

event of nuclear or radiological emergencies were

not established in these documents. Appropriate

specific recommendations were developed later on

intervention for the protection of the public in a

radiological emergency [4.14]. In this guidance the

optimization concept was confirmed as the basic

one applicable in the event of an emergency and

further elaborated with regard to dose averted as

the consequence of intervention (see Fig. 4.1). The

ICRP discarded the previous two level intervention

criteria and recommended instead some inter-

vention levels (in terms of averted effective dose) —

50 mSv for sheltering, 500 mSv (thyroid dose) for

administration of stable iodine, 500 mSv for

evacuation, 1000 mSv (lifetime dose) for relocation

and 10 mSv/a for the control of foodstuffs.

A more recent ICRP publication (Publication

82) [4.15] considered public radiation protection in

conditions of prolonged radiation exposure, such as

in areas contaminated due to the Chernobyl

accident. In this document, the ICRP generally

recommends retaining the optimization principle,

but also suggests generic radiological criteria for

making decisions on countermeasure application. In

particular, it proposes the value of the ‘existing

annual dose’, including external and internal doses

from natural and human-made radionuclides, of

10 mSv as the generic dose below which inter-

vention is not usually expedient. This does not

exclude intervention at lower doses if site specific

optimization analysis proves it to be expedient.

Inter alia, the ICRP recommended a generic inter-

vention exemption level for radionuclides in

commodities dominating human exposure equal to

1 mSv/a. This criterion could be applied for justifi-

cation of the reference levels for radionuclides in

food.

TABLE 4.1. GUIDELINE LEVELS FOR RADIO-

NUCLIDES IN FOOD FOLLOWING ACCI-

DENTAL NUCLEAR CONTAMINATION, FOR

USE IN INTERNATIONAL TRADE [4.12]

Food for general

consumption

(Bq/g)

Milk and

infant food

(Bq/g)

Caesium-134, 137 1 1

Iodine-131 1 0.1

Strontium-90 0.1 0.1

Plutonium-239,

americium-241

0.01 0.001

Time after start of accident

Avertable

dose ('E)

Dose per unit time (E(t))

FIG. 4.1. Avertable dose and effective dose accumulated

er unit time as a function of time when the protective

measure is introduced at time t

and lifted again at time t

4.1.2. National radiological criteria and standards

Limitations on human exposure caused by the

Chernobyl accident, including standards for radio-

nuclides in food, drinking water, timber, etc., were

introduced soon after the accident, first by the

USSR but also by many other European countries

(i.e. Nordic countries, EU countries and eastern

European countries [4.1]).

In accordance with the Standards of Radiation

Safety [4.16] in force in 1986, the USSR Ministry of

Health introduced a temporary limit of average

equivalent whole body dose of 100 mSv for the first

year after the Chernobyl accident (from 26 April

1986 until 26 April 1987), then 30 mSv for the

second year and 25 mSv in each of 1988 and 1989

[4.3]. In all, until 1 January 1990, a dose to the

general public not exceeding 173 mSv was allowed

from the radioactive fallout of the Chernobyl

accident.

In order to limit the internal exposure of

members of the population, temporary permissible

levels (TPLs) of radionuclide content in food

products and drinking water were introduced in the

USSR. Table 4.2 presents the TPLs for the main

food products [4.3, 4.17]. The first TPL set approved

TABLE 4.2. TEMPORARY PERMISSIBLE LEVELS (Bq/kg) OF RADIONUCLIDE CONTENT IN

FOOD PRODUCTS AND DRINKING WATER ESTABLISHED IN THE USSR (1986–1991) AFTER THE

CHERNOBYL ACCIDENT [4.3, 4.17]

TPL

4104–88 129–252 TPL-88 TPL-91

Date of adoption 6 May 1986 30 May 1986 15 December 1987 22 January 1991

Radionuclide Iodine-131 Beta emitters Caesium-134 and

caesium-137

Caesium-134 and

caesium-137

Strontium-90

Drinking water 3700 370 18.5 18.5 3.7

Milk 370–3700 370–3700 370 370 37

Dairy products 18 500–74 000 3700–18 500 370–1850 370–1850 37–185

Meat and meat products — 3700 1850–3000 740 —

Fish 37 000 3700 1850 740 —

Eggs — 37 000 1850 740 —

Vegetables, fruit, potato,

root crops

— 3700 740 600 37

Bread, flour, cereals — 370 370 370 37

TABLE 4.3. ACTION LEVELS (Bq/kg) FOR CAESIUM RADIONUCLIDES IN FOOD PRODUCTS

ESTABLISHED AFTER THE CHERNOBYL ACCIDENT [4.3, 4.5]

Action level

Codex Alimentarius

Commission

EU Belarus

Russian

Federation

Ukraine

Year of adoption 1989 1986 1999 2001 1997

Milk 1000 370 100 100 100

Infant food 1000 370 37 40–60 40

Dairy products 1000 600 50–200 100–500 100

Meat and meat products 1000 600 180–500 160 200

Fish 1000 600 150 130 150

Eggs 1000 600 — 80 6 Bq/egg

Vegetables, fruit, potato, root crops 1000 600 40–100 40–120 40–70

Bread, flour, cereals 1000 600 40 40–60 20

by the USSR Ministry of Health on 6 May 1986

concerned

131

I activity concentrations in foodstuffs

and was aimed at limiting the thyroid dose to

children to 300 mGy. The next TPL set, adopted on

30 May 1986, concerned the content of all beta

emitters in food products caused by surface contam-

ination, with particular attention given to ecologi-

cally mobile and long lived caesium radionuclides.

Later, TPLs introduced in 1988 (TPL-88) and 1991

(TPL-91) concerned the sum of

134

Cs and

137

activities. The TPL-91 for caesium radionuclides

was supplemented by TPLs for

Sr.

Annual consumption by rural inhabitants of

the usual food ration, if all components contained

caesium radionuclides at the level of TPL-86, would

cause an internal dose of less than 50 mSv (at

TPL-88 it would be less than 8 mSv and at TPL-91 it

would be less than 5 mSv).

Action levels for

131

I in food established in

some European countries in May 1986 varied within

the range of 500–5000 Bq/kg. Later, the authorities

of the EU established two values for caesium radio-

nuclides in imported food, one for milk and infant

food and another for all other food products (see

Table 4.3) [4.3, 4.5]. Similar values were introduced

in Nordic countries, with an exception for wild foods

(reindeer meat, game, freshwater fish, forest

berries, fungi and nuts), which are important

products for some local populations and especially

for indigenous people. Thus in the first month

Sweden imposed action levels of 5 kBq/kg for

131

and 10 kBq/kg for

137

Cs in imported food; for

domestic foods the respective values were 2 and

1 kBq/kg. In the middle of May, action levels of

300 Bq/kg for

137

Cs in all food and 2 kBq/kg for

131

in milk and dairy products were introduced. For

wild foods produced or consumed in the Nordic

countries, the action levels varied between 1500 and

6000 Bq/kg in different countries and time periods.

Along with the standards for food products,

standards were introduced by the USSR for agricul-

tural raw material, wood (see Section 4.3) and

herbs, and for beta contamination of different

surfaces [4.3].

The general policy of the USSR, and later of

the authorities in the separate republics, was to

reduce both the radiological criteria and the TPLs

along with the natural improvement of radiological

conditions due to radionuclide decay and

penetration/fixation in soil. Gradual TPL reduction

has been used as an instrument to force producers

to apply technologies that decrease radionuclide

content in products in order to limit associated

human exposure. The TPLs were established by

experts balancing the desire to reduce internal dose

in populations with the need to maintain profitable

agricultural production and forestry in the

controlled areas. Different reference levels for

numerous groups of food products were established

with the aim of not restricting consumption of any

foods unless the dose criterion might be exceeded.

By the end of 1991 the USSR had split into

separate countries, among them Belarus, the

Russian Federation and Ukraine, which had been

strongly affected by the Chernobyl accident.

Afterwards each country implemented its own

policy of radiation protection of the public. Owing

to the acceptance by the ICRP in 1990 of the annual

effective dose limit for the public in regulated

situations (practices) equal to 1 mSv, this level was

considered by the authorities of the three countries

as safe also in post-emergency conditions. Therefore

it is still used in national legislations as an inter-

vention level of annual dose caused by Chernobyl

fallout for the introduction of countermeasures,

including long term remediation measures.

Current national TPLs for food products,

drinking water and wood in the three countries are

comparable with each other (see Table 4.3), and all

of them are substantially lower than both the EU

maximum permissible levels for import [4.5] and the

Codex Alimentarius Commission’s guidance levels

for radionuclides in food moving in international

trade [4.12].

The State authorities in the three countries

have struggled to meet the established TPLs for

products and the dose criteria by implementation of

environmental countermeasures as described below

and by inspection of foods throughout each country.

4.2. URBAN DECONTAMINATION

Decontamination of settlements was one of

the main countermeasures applied to reduce

external exposure of the public and cleanup

workers during the initial stage of response to the

Chernobyl accident. The immediate purpose of

settlement decontamination was the removal of

radiation sources distributed in urban environments

inhabited by humans.

Analysis of the sources of external exposure in

different population groups living in contaminated

areas revealed that a significant fraction of dose is

received by people from sources located in soil, on

coated surfaces such as asphalt and concrete and to

a small extent on building walls and roofs. This is

why most effective decontamination technologies

involved removal of the upper soil layer.

The decontamination efficiency can be charac-

terized by means of the following parameters: the

dose rate reduction factor (DRRF), which is the

relative reduction of dose rate above a surface

following decontamination, and the dose reduction

factor (DRF), which is the reduction of the

effective external dose to an individual from

gamma emitting radionuclides deposited in the

environment.

4.2.1. Decontamination research

In order to ensure high decontamination

effectiveness and to keep the associated costs low,

several research projects have been implemented

aimed at determining the values of the DRRF and

DRF for particular decontamination technologies

applied to different surfaces and artefacts in the

human environment [4.18–4.20]. Reports from

these experimental and theoretical studies contain

validated models of urban decontamination and sets

of model parameters and practical recommenda-

tions for cleanup in different time periods after

urban radioactive contamination. A preliminary

remediation assessment based on well developed

cost–benefit techniques is recommended in order to

justify decontamination and to optimize its imple-

mentation.

According to these and other studies, the

contributions of different urban surfaces to the

human external dose and the associated opportu-

nities for dose reduction are determined by

settlement and house design, construction material,

the habits of the populations, the mode of radionu-

clide deposition (dry or wet), the radionuclide and

physicochemical composition of the fallout, and

time (see Section 3.2).

Following dry deposition, street cleaning,

removal of trees and shrubs and ploughing of

gardens are efficient and inexpensive means of

achieving very significant reductions in dose and

would rate highly in a list of short term remediation

priorities. Roofs are important contributors to dose,

but the cost of cleaning roofs is high and this

countermeasure would not rank highly in a list of

priorities. Walls contribute little to dose, are

expensive and difficult to decontaminate and would

therefore carry a very low rating.

In the case of wet deposition, gardens and

lawns, both in the short term and the long term,

would be given first priority, because a considerable

reduction in dose (~60%) can be achieved at

relatively low cost. Street cleaning would also be of

benefit.

While planning decontamination for the long

term, it is important to take into account the contri-

bution of external dose to the total (external and

internal) dose. In areas dominated by clay soils,

transfer of caesium radionuclides in the food chain

and the associated internal doses are low. In these

areas the relative decrease in the total dose is close

to the DRF value. In contrast, in sandy and peaty

soil areas, where long term internal exposure

dominates, the relative decrease in the total dose

due to village decontamination is expected to be

less significant.

4.2.2. Chernobyl experience

Large scale decontamination was performed

between 1986 and 1989 in the cities and villages of

the USSR most contaminated after the Chernobyl

accident. This activity was performed usually by

military personnel and included washing of buildings

with water or special solutions, cleaning of

residential areas, removal of contaminated soil,

cleaning and washing of roads, and decontamination

of open water supplies. Special attention was paid to

kindergartens, schools, hospitals and other buildings

frequently visited by large numbers of persons. In

total, about one thousand settlements were treated;

this included cleaning tens of thousands of

residences and public buildings and more than a

thousand agricultural farms [4.18, 4.21, 4.22].

In the early period following the accident,

inhalation of resuspended radioactive particles of

soil and nuclear fuel could contribute significantly

to internal dose. To suppress dust formation,

dispersion of organic solutions over contaminated

plots was used in order to create an invisible

polymer film after drying. This method was

implemented at the Chernobyl nuclear power plant

and in the CEZ during the spring and summer of

1986. Streets in cities were watered to prevent dust

formation and to remove radionuclides to the

sewerage system. The effectiveness of early decon-

tamination efforts in 1986 still remains to be

quantified. However, according to Los and

Likhtarev [4.23] daily washing of streets in Kiev

decreased the collective external dose to its

three million inhabitants by 3000 man Sv, and

decontamination of schools and school areas saved

another 600 man Sv.

Depending on the decontamination technol-

ogies used, the dose rate over different measured

plots was reduced by a factor of 1.5–15. However,

the high cost of these activities hindered their

comprehensive application on contaminated areas.

Due to these limitations, the actual effectiveness of

the decrease in annual external dose was 10–20%

for the average population and ranged from about

30% for children visiting kindergartens and schools

to less than 10% for outdoor workers (herders,

foresters, etc.). These data were confirmed by

individual external dose measurements conducted

before and after large scale decontamination

campaigns in 1989 in the Bryansk region of the

Russian Federation [4.18].

Regular monitoring of decontaminated plots

in settlements over five years showed that after 1986

there was no significant recontamination and that

the exposure rate was decreasing over the long

term, as described in Section 5.1 of this report. The

averted collective external dose to 90 000

inhabitants of the 93 most contaminated settlements

of the Bryansk region was estimated to be about

1000 man Sv [4.18].

Since 1990 large scale decontamination in the

countries of the former USSR has been stopped, but

particular contaminated plots and buildings with

measured high contamination levels have been

specifically cleaned. Some decontamination

activities still continue in Belarus, aimed mostly at

public buildings and areas: hospitals, schools,

recreation areas, etc. However, in some contami-

nated Belarusian villages, cleanup of dwellings and

farms has also been performed [4.22].

Another area of continuing decontamination

activity is the cleanup of industrial equipment and

premises contaminated as a result of ventilation

systems being operated during the release/

deposition period in 1986 and immediately

afterwards. Some 20 to 30 industrial buildings and

ventilation systems have been decontaminated

annually in Belarus [4.22].

4.2.3. Recommended decontamination

technologies

In accordance with present radiation

protection methodology, a decision on intervention

(decontamination) and selection of optimal decon-

tamination technologies should be made giving

consideration to the costs of all actions and to social

factors. The calculated cost should address the

various decontamination technologies for which an

assessment of the averted dose has been made. The

benefit (averted collective effective dose) and

detriment (expenses, collective dose to decontami-

nation workers) are to be compared for each decon-

tamination technology by means of a cost–benefit

analysis [4.9] or multiattribute analysis [4.24], which

may include qualitative social factors.

The priorities that different procedures would

be given in a decontamination strategy should be

environment specific. Nevertheless, based on

accumulated experience and research, the following

generic set of the major simple decontamination

procedures can be recommended for the long term:

(a) Removal of the upper 5–10 cm layer

(depending on the activity–depth distribution)

of soil in courtyards in front of residential

buildings, around public buildings, schools and

kindergartens, and from roadsides inside a

settlement. The removed, most contaminated,

layer of soil should be placed into holes

specially dug on the territory of a private

homestead or on the territory of a settlement.

The clean soil from the holes should be used to

cover the decontaminated areas. Such a

technology excludes the formation of special

burial sites for radioactive waste.

(b) Private fruit gardens should be treated by

deep ploughing or removal of the upper

10 cm layer of soil. By now, vegetable

gardens have been ploughed many times, and

the activity distribution in soil will be uniform

in a layer 20–30 cm deep.

courtyards, etc., with a layer of clean sand, or,

where possible, with a layer of gravel to

attenuate residual radiation (see item (a)).

(d) Cleaning or replacement of roofs.

These procedures can be applied both for

decontaminating single private gardens and houses

and for decontaminating settlements as a whole. It

is evident that, in the latter case, the influence of the

decontamination on further reduction in external

radiation dose will be greater. Achievable decon-

tamination factors for various urban surfaces are

presented in Table 4.4. Detailed data on the

efficiency, technology, necessary equipment, cost

and time expenses, quantity of radioactive waste,

and other parameters of decontamination

procedures are contained in Ref. [4.25].

Radioactive waste generated from urban

decontamination should be disposed of in

accordance with established regulatory require-

ments. In the event of large scale decontamination,

temporary storage should be arranged in special

isolated areas from which future activity release

into the environment will be negligible. The site

should be marked by the international symbol of

radiation hazard.

4.3. AGRICULTURAL

COUNTERMEASURES

The implementation of agricultural counter-

measures after the Chernobyl accident has been

extensive, both in the most severely affected

countries of the former USSR and in western

Europe. The main aim of agricultural counter-

measures was the production of food products with

radionuclide activity concentrations below action

levels

. The application of countermeasures in

intensive agricultural production systems was

largely confined to Belarus, the Russian Federation

and Ukraine, although some food bans were

initially applied in western Europe. Many counter-

measures were used extensively in the first few

years after the accident, and their application

continues today. In addition, in these three

countries countermeasures have been applied to

private food production from unimproved

meadows, where high

137

Cs activity concentrations

have persisted for many years [4.3, 4.4, 4.7].

High and persistent transfer of

137

Cs has also

occurred in many contaminated areas of western

Europe. In these countries countermeasures have

largely been focused on animal food products, for

example for grazing animals on unimproved

pastures.

4.3.1. Early phase

From 2–5 May 1986 about 50 000 cattle, 13 000

pigs, 3300 sheep and 700 horses were evacuated

from the CEZ together with the people [4.26]. In

the CEZ more than 20 000 agricultural and

domestic animals, including cats and dogs,

remaining after the evacuation were killed and

buried. Due to a lack of forage for the evacuated

animals and difficulties in managing the large

number of animals in the territories to which they

had been moved, many of the evacuated animals

were also slaughtered [4.27, 4.28]. In the acute

period after the accident it was not possible to

differentiate between the different levels of contam-

ination of animals, and in the period May–July 1986

the total number of slaughtered animals reached

95 500 cattle and 23 000 pigs.

Many carcasses were buried and some were

stored in refrigerators, but this produced great

hygiene, practical and economic difficulties.

Condemnation of meat was an immediately

available and effective countermeasure to reduce

ingested dose from animal products and was widely

used in the USSR and elsewhere. However, this was

very expensive and resulted in large quantities of

contaminated waste.

In the first weeks after the accident, the main

aim of countermeasure application in the USSR was

to lower

131

I activity concentrations in milk or to

prevent contaminated milk from entering the food

chain. The following were recommendations [4.29]

on how to achieve this:

Referred to in the countries of the former USSR

as temporary permissible levels (TPLs).

TABLE 4.4. ACHIEVABLE DECONTAMINATION FACTORS (DIMENSIONLESS)

FOR VARIOUS URBAN SURFACES [4.25]

Technique DRRF

Windows Washing 10

Walls Sandblasting 10–100

Roofs Hosing and/or sandblasting 1–100

Gardens Digging 6

Gardens Removal of surface 4–10

Trees and shrubs Cutting back or removal ~10

Streets Sweeping and vacuum cleaning 1–50

Streets (asphalt) Lining >100

(a) Exclusion of contaminated pasture grasses

from the animals’ diet by changing from

pasture to indoor feeding of uncontaminated

feed;

(b) Radiation monitoring at processing plants and

subsequent rejection of milk in which

131

activity concentrations were above the action

level (3700 Bq/L at that time);

milk to storable products such as condensed or

dried milk, cheese or butter).

In the first few days after the accident the

countermeasures were largely directed towards

milk from collective farms, and few private farmers

were involved. Information on countermeasures for

milk was only given to managers and local

authorities and was not distributed to the private

farming system of the rural population. This

resulted in limited application of countermeasures,

especially for privately produced milk in rural

settlements, resulting in a low effectiveness in some

areas.

Within a few weeks of the accident, feeding of

animals with ‘clean’ fodder began because this had

the potential to reduce

137

Cs in cattle to acceptable

levels within a period of 1–2 months. However, this

countermeasure was not in widespread use at this

stage, partly due to a lack of availability of uncon-

taminated feed early in the growing season.

As early as the beginning of June 1986 maps

were constructed of the density of radioactive

deposition in the contaminated regions. This

allowed estimates to be made of the extent of the

contamination of pasture and identification of

where contaminated milk would be found.

During the growing period of 1986, when

there was still substantial surface contamination of

plants, the major countermeasures in agriculture

were of a restrictive nature. In the first few months

severely contaminated land was taken out of use

and recommendations were developed on suitable

countermeasures that would allow continued

production on less heavily contaminated land. In

the more heavily affected regions, a ban was

imposed on keeping dairy cattle. To reduce contam-

ination levels in crops, an effective method was to

delay harvesting of forage and food crops.

Radiation control of products was introduced at

each stage of food production, storage and

processing [4.3, 4.30].

Based on a radiological survey performed

from May to July 1986, approximately 130 000,

17 300 and 57 000 ha of agricultural land were

initially excluded from economic use in Belarus, the

Russian Federation and Ukraine, respectively

[4.31].

From June 1986 other countermeasures aimed

at reducing

137

Cs uptake into farm products were

implemented as follows:

(i) Banning cattle slaughter in regions where

137

Cs contamination levels exceeded 555 kBq/

(animals had to be fed clean food for

1.5 months before slaughter);

(ii) Minimizing external exposure and formation

of contaminated dust by omitting some

procedures normally used in crop production;

(iii) Limiting the use of contaminated manure for

fertilization;

(iv) Preparation of silage from maize instead of

hay;

(v) Restricting the consumption of milk produced

in the private sector;

(vi) Obligatory radiological monitoring of agricul-

tural products;

(vii) Obligatory milk processing.

Decontamination by removal of the top soil

layer was not found to be appropriate for agricul-

tural lands because of its high cost, destruction of

soil fertility and severe ecological problems related

to burial of the contaminated soil.

As early as August–September 1986 each

collective farm received maps of contamination

levels of their agricultural land and guidance on

potential contamination of products, including

instructions on the farming of private plots [4.3,

4.30].

In western Europe advice was initially given

on avoiding the consumption of drinking water

from local supplies in some countries.

Sweden received some of the highest levels of

deposition outside of the countries of the former

USSR. Initially, Sweden imposed action levels on

131

I and

137

Cs activities in imported and domestic

foods (see Section 4.1.2). A range of other

responses were applied: (a) cattle were not put on to

pasture if the ground deposition exceeded 10 kBq/

131

I and 3 kBq/m

of radiocaesium; (b) advice

was given not to consume fresh leafy vegetables and

to wash other fresh vegetables; (c) restrictions were

placed on the use of sewage sludge as fertilizer for

soil; (d) deep ploughing was recommended; and (e)

a higher cutting level for harvesting of grass was

advised.

In Norway, crops in fields were monitored

after harvesting, and those with radiocaesium above

600 Bq/kg fresh weight were discarded and ploughed

in. Also, hay and silage harvested in June were

monitored, and that with activity concentrations

exceeding the guidelines was not used as forage.

In Germany some milk in Bavaria was

diverted into food processing plants to be converted

into milk powder. It was intended to use the milk

powder as feed for pigs, but this was not done due to

the high radiocaesium content.

In the UK advice was issued for the regulation

of the consumption of red grouse, and restrictions

were imposed on the movement and slaughter of

upland sheep from a number of the more contami-

nated areas of the UK.

In Austria there was advice not to feed fresh

grass to cows for a short period in May 1986 [4.32].

4.3.2. Late phase

Radiological surveys of agricultural products

showed that by the end of 1986 four regions of the

Russian Federation (Bryansk, Tula, Kaluga and

Orel), five regions of Ukraine (Kiev, Zhytomyr,

Rovno, Volyn and Chernigov) and three regions of

Belarus (Gomel, Mogilev and Brest) had food

products that exceeded the action levels for radio-

caesium. In the more contaminated areas of the

Gomel, Mogilev, Bryansk, Kiev and Zhytomyr

regions in the first year after the accident, the

proportion of grain and milk exceeding the action

levels was about 80% [4.3, 4.7, 4.26].

Additionally, in the early 1990s in Ukraine,

101 285 ha of agricultural land was withdrawn from

agricultural use (about 30% of this area had a

137

contamination level above 555 kBq/m

). Privately

owned cattle were moved with the people from

some settlements. Provision of ‘clean’ foodstuffs

produced in the collective sector or imported from

‘clean’ regions was organized for those residents not

resettled.

In the Russian Federation in 1987–1988

further evacuations of agricultural animals were

carried out, but on a more elective basis than in

Ukraine. All sheep in the areas contaminated at

over 555 kBq/m

were removed, because of the high

transfer of radiocaesium to these ruminants. Of

cattle in the regions above 555 kBq/m

, 6880

animals were removed, but many families retained

their animals.

In Belarus in 1989, 52 settlements were

relocated after decontamination and counter-

measure use were found to be inadequate to lower

doses to an acceptable level. Additionally, in 1991

under two new laws, some people were allowed to

resettle away from contaminated areas, and some

settlements were moved. In total, 470 settlements

were moved. In all these resettlements, the agricul-

tural animals accompanied their owners to the new

locations where possible.

Application of countermeasures in contami-

nated areas had two major radiation protection

aims. The first was to guarantee foodstuff

production corresponding to the action levels and to

ensure an annual effective dose to local inhabitants

of less than 1 mSv. The second was to minimize the

total flux of radionuclides in agricultural

production. Generally, the earlier agricultural

countermeasures were applied, the more cost

effective they were [4.33].

From 1987 high radiocaesium activity concen-

trations in agricultural products were only observed

in animal products; application of countermeasures

aimed at lowering

137

Cs activity concentrations in

milk and meat was the key focus of the remediation

strategy for intensive agriculture. Potatoes and root

vegetables were being produced with acceptably

low radiocaesium levels. In the second year after the

accident the radiocaesium activity concentration in

grain was much lower than in the first year, and

countermeasure application ensured that most grain

was below the action levels. By 1991 less than 0.1%

of grain in all three countries had radiocaesium

contents above 370 Bq/kg.

The most difficult issue remaining was the

production of milk in compliance with the

standards. However, large scale application of a

range of countermeasures (described below) made

it possible to achieve a sharp decrease in the

amount of animal products with radiocaesium

activity concentrations above the action levels in all

three countries. Changes with time of milk and meat

exceeding the action levels can be seen in Fig. 4.2. It

should be noted that the values of the action levels

have been reduced with time in each of the three

countries, so the data are not directly comparable.

Changes in the action levels in each country are

shown in Fig. 4.3.

The differences in the time trend shown in

Fig. 4.2 among the countries mainly relate to

changes in the action levels but also to the scale of

countermeasure application. This is particularly

clear for Russian milk, where radiocaesium

activity concentrations rose after 1997 due to a

reduction in countermeasure use. The recent

reduction in the amounts of meat above the action

levels in Ukraine and Belarus is because animals

are monitored before slaughter to ensure that the

meat is below the required level. In the Russian

Federation, where animals are also monitored

before slaughter, the concentration data are

higher, because they refer to both privately and

collectively produced meat.

The maximum effect from countermeasure

application was achieved in 1986–1992. Thereafter,

because of financial constraints in the mid-1990s,

the use of agricultural countermeasures was

drastically reduced. However, by optimizing

available resources,

137

Cs countermeasure

effectiveness remained at a level sufficient to

maintain an acceptable

137

Cs content in most animal

products (Fig. 4.2).

4.3.3. Countermeasures in intensive agricultural

production

The main countermeasures used in the USSR,

and later in the three independent countries, are

briefly described below. The priority was on

chemical amendments to improve soil fertility and

to reduce the uptake of radiocaesium by crops and

plants used for fodder. The extent to which each

measure was used varied among the three countries.

The recommendations on countermeasures have

been repeatedly revised and updated [4.35–4.37].

100

1000

6891

8891

0991

2991

499

991

8991

0002

2002

4002

Year

Russian Federation

Belarus

(a)

100

1000

10 000

100 000

6891

8891

0991

2991

4991

6991

8991

0002

2002

Year

Russian Federation

Ukraine

Belarus

(b)

t (1000)

FIG. 4.2. Amounts of milk (a) and meat (b) exceeding the action levels in the Russian Federation (collective and private),

Ukraine and Belarus (only milk and meat entering processing plants) after the Chernobyl accident [4.26].

100

150

200

250

300

350

400

1985 1990 1995 2000 2005 2010

Year

Russian

Federation

Ukraine

Belarus

500

1000

1500

2000

2500

3000

3500

4000

1985 1990 1995 2000 2005 2010

Year

Russian

Federation

Ukraine

Belarus

TPL for meat (Bq/kg)

TPL for milk (Bq/L)

FIG. 4.3. Changes with time in the action levels (TPL) in the USSR and later in the three independent countries [4.34].

4.3.3.1. Soil treatment

Soil treatment reduces uptake of radio-

caesium (and radiostrontium). The procedure can

involve ploughing, reseeding and/or the application

of nitrogen, phosphorus, potassium (NPK)

fertilizers and lime. Ploughing dilutes the

radioactive contamination originally in the upper

soil layers, where most plant roots absorb their

nutrients. Both deep and shallow ploughing were

used extensively, and skim and burial ploughing

were also used. The use of fertilizers increases plant

production, thereby diluting the radioactivity in the

plant. In addition, the use of fertilizers reduces root

uptake into plants by decreasing the Cs:K ratio in

the soil solution [4.30].

When soil treatment includes all the above

measures it is commonly called radical

improvement; this has been found to be the most

efficient and practical countermeasure for meadows

contaminated by Chernobyl fallout. In the first few

years after the accident the focus was on radical

improvement, including greatly increased fertili-

zation rates. Commonly, high value legumes and

cereal grasses were grown on the treated land. The

nature of the treatment and the efficiency of the

radical improvement of hay meadows and pastures

strongly depend on the type of meadow and the soil

properties. Traditional surface improvement,

involving soil discing, fertilization and surface

liming, was less effective. Some marshy plots were

drained, deep ploughed, improved and used as

grassland. In the 1990s there was a greater focus on

site specific characteristics to ensure that the soil

treatment used was the most appropriate and

effective for the prevailing conditions. With time,

repeated fertilization of already treated soils was

necessary, but the appropriate application rates

were carefully assessed. However, actual rates of

application were sometimes constrained by availa-

bility of funds [4.30, 4.38].

Areas that received additional fertilizers in

each of the three most affected countries are shown

in Fig. 4.4; areas receiving radical improvement are

shown in Fig. 4.5. The average amount of additional

potassium fertilizers added was about 60 kg/ha of

O annually between 1986 and 1994. In the mid-

1990s the productivity of arable land fell because a

worsening economic condition prevented the imple-

mentation of countermeasures at the previous rates;

this resulted in an increasing proportion of contami-

nated products. In some areas of the Russian

500

1000

1500

2000

2500

3000

1986–1990 1991–1995 1996–2000 2001–2003

Year

Russian Federation

Ukraine

Belarus

1000

2000

3000

4000

5000

6000

7000

8000

Year

Russian Federation

Ukraine

Belarus

1986–1990 1991–1995 1996–2000 2001–2003

(a) (b)

ha (1000)

FIG. 4.4. Changes in the extent of agricultural areas treated with liming (a) and mineral fertilizers (b) in the countries mos

affected by the Chernobyl accident [4.34].

100

200

300

400

500

600

1986–1990 1991–1995 1996–2000 2001–2003

Year

Russian Federation

Ukraine

Belarus

ha (1000)

FIG. 4.5. Areas of radical improvement in the countries

most affected by the Chernobyl accident [4.34].

Federation this halted the previous decrease in the

amounts of milk and meat exceeding the radiation

safety standards (see Fig. 4.2); for example, in the

more contaminated areas, such as Novozybkov

(Bryansk region), because of insufficient use of

potassium fertilizers,

137

Cs activity concentrations in

agricultural products in 1995–1996 increased by

more than 50% compared with the period of

optimal countermeasure application (1991–1992).

The effectiveness of soil treatment is

influenced by soil type, nutrient status and pH, and

also by the plant species selected for reseeding. In

addition, the application rates of NPK fertilizers

and lime affect the reduction achieved. Several

studies have shown that the reduction factors

achieved for soil–plant transfer of radiocaesium

following radical improvement, liming and fertili-

zation were in the range of two to four for poor,

sandy soils and three to six for more organic soils.

An added benefit was the reduction in external dose

rate by a factor of two to three due to the dilution of

the surface contamination layer after ploughing.

Even though the radiological problems

associated with

Sr are less acute than those of

137

Cs, some countermeasures have been developed

and a reduction of two to four in the soil–plant

transfer of radiostrontium following discing,

ploughing and reseeding has been achieved.

Despite these countermeasures, in the more

highly contaminated areas of the Bryansk region

the radiocaesium contamination of 20% of the

pasture and hay on farms still exceeded the action

levels in 1997–2000. Concentrations of

137

Cs in hay

varied between 650 and 66 000 Bq/kg dry weight.

4.3.3.2. Change in fodder crops grown on

contaminated land

Some plant species take up less radiocaesium

than others, as can be seen from experimental data

collated in Belarus from 1997 until 2002 (Fig. 4.6).

The extent of the difference is considerable, and

fodder crops such as lupin, peas, buckwheat and

clover, which accumulate high amounts of radio-

caesium, were completely or partly excluded from

cultivation.

In Belarus rapeseed is grown on contaminated

areas with the aim of producing two products:

edible oil and protein cake for animal fodder.

Varieties of rapeseed are grown that have a twofold

to threefold lower

137

Cs and

Sr uptake rate than

other varieties. When the rapeseed is grown,

additional fertilizers (liming with 6 t/ha and fertili-

zation with N

180

) are used to reduce radio-

caesium and radiostrontium uptake into the plant

by a factor of about two. This reduces contami-

nation of the seed that is used for the protein cake.

During processing of the rapeseed, both radio-

caesium and radiostrontium are effectively

removed, and negligible amounts remain. The

production of rapeseed oil in this way has proved to

be an effective, economically viable way to use

contaminated land and is profitable for both the

farmer and the processing industry. During the past

decade the area under rapeseed cultivation has

increased fourfold to 22 000 ha [4.40].

4.3.3.3. Clean feeding

The provision of uncontaminated feed or

pasture to previously contaminated animals for an

appropriate period before slaughter or milking

(‘clean’ feeding) effectively reduces radionuclide

contamination, respectively, in meat and milk at a

rate that depends on the animal’s biological half-life

for each radionuclide. Radiocaesium activity

concentration in milk responds rapidly to changes in

diet, as the biological half-life is a few days. For

meat the response time is longer, due to the longer

biological half-life in muscle [4.28].

Clean feeding reduces uptake of the contami-

nating radionuclides; it has been one of the most

important and frequently used countermeasures

after the Chernobyl accident for meat from agricul-

tural animals in both the countries of the former

USSR and western Europe. Official estimates of the

number of cattle treated are between 5000 and

20 000 annually in the Russian Federation and

20 000 in Ukraine (supported by the government up

100

Carrot Cabbage Tomato Potato Haricot Table Cucumber Radish Pea

beet

Percentage compared with peas

FIG. 4.6. Comparison of

137

Cs uptake in different crops,

normalized to that of peas [4.39].

to 1996) [4.3]. Clean feeding is routinely used in all

three countries for meat production and is

combined with live monitoring of animals, so that if

animal flesh is above the action levels the animals

can be returned to the farm for further clean

feeding.

4.3.3.4. Administration of caesium binders

Hexacyanoferrate compounds (commonly

referred to as Prussian blue) are highly effective

radiocaesium binders. They may be added to the

diet of dairy cows, sheep and goats, as well as to

meat producing animals, to reduce radiocaesium

transfer to milk and meat by reducing absorption in

the gut. They have a low toxicity and are therefore

safe to use. Many different formulations of hexa-

cyanoferrates have been developed in different

countries, partly to identify the most effective

compound and partly to produce a cheaper, locally

available product. Hexacyanoferrate compounds

can achieve reduction factors in animal products of

up to ten [4.41].

Prussian blue can be added to the diet of

animals as a powder, incorporated into pelleted

feed during manufacturing or mixed with sawdust.

A locally manufactured hexacyanoferrate called

ferrocyn (a mixture of 5% KFe[Fe(CN)

] and 95%

[Fe(CN)

]) has been developed in the Russian

Federation. It has been administered as 98% pure

powder, salt licks (10% ferrocyn) and in sawdust

with 10% adsorbed ferrocyn (called bifege) [4.42].

The number of cattle treated annually with

Prussian blue in each of the three countries is shown

in Fig. 4.7. In addition, slow release boli containing

hexacyanoferrate have been developed that are

introduced into the animals’ rumen and gradually

release the caesium binder over a few months. The

boli, originally developed in Norway, consist of a

compressed mixture of 15% hexacyanoferrate, 10%

beeswax and 75% barite [4.43].

Prussian blue has been used to reduce the

137

Cs contamination of animal products since the

beginning of the 1990s. Prussian blue application

has been especially useful and effective in

settlements where there is a lack of meadows

suitable for radical improvement. In initial trials,

Prussian blue reduced

137

Cs transfer from fodder to

milk and meat by a factor of 1.5–6.0 [4.44]. In

Belarus a special concentrate with Prussian blue is

distributed at a rate of 0.5 kg of concentrate per cow

daily, and an average value of three for the

reduction factor for milk has been achieved.

Prussian blue has not been used as extensively

in Ukraine as in the Russian Federation and

Belarus, and its use was confined to the early 1990s.

This is because in Ukraine no local source of

Prussian blue is available and the cost of purchasing

it from western Europe was considered to be too

high. Therefore, instead, locally available clay

mineral binders have been used on a small scale.

These were cheaper but somewhat less effective

than Prussian blue.

4.3.4. Summary of countermeasure effectiveness

in intensive production

The effectiveness of the different agricultural

countermeasures in use on farms is summarized in

Table 4.5. The reduction factors (ratio of radio-

caesium activity concentration in the product before

and after countermeasure application) achieved by

each measure are given.

4.3.5. Countermeasures in extensive production

Extensive production in the three countries of

the former USSR is largely confined to the grazing

of privately owned cattle on poor, unimproved

meadows. Owing to the poor productivity of these

areas, radiocaesium uptake is relatively high

compared with land used by collective farms.

Radical improvement of meadows used by privately

owned cattle has been applied in all three countries

since the early 1990s. Clean feeding is not generally

used by private farmers, although, on occasions,

collective farms may supply private farmers with

uncontaminated feed or pastures. Prussian blue is

used by private farmers in both the Russian

Federation and Belarus. In the Russian Federation

all three Prussian blue delivery systems are used,

according to availability and preference [4.46].

Russia

Ukraine

Belarus

Russian Federation

Ukraine

Belarus

Year

Number of treated cattle (1000)

1991

1992

1993

1994

1995

1996

1997

1998

1999

2000

2001

2002

2003

FIG. 4.7. Changes with time in the use of Prussian blue in

the three countries of the former USSR (provided by

Forum participants from official national sources).

In extensive systems such as upland grazed

areas in western Europe, the most commonly used

countermeasures for free ranging animals have

been clean feeding, administration of caesium

binders, monitoring of live animals, management

restrictions and changes in slaughter times. Many of

these countermeasures were still in use in 2004. The

application of long term countermeasures has been

most extensive in Norway and Sweden, but long

term countermeasures have also been applied in the

UK and Ireland.

AFCF is a highly effective hexacyanoferrate

compound achieving up to a fivefold reduction in

radiocaesium in lamb and reindeer meat and up to a

threefold reduction in cow’s milk and a fivefold

reduction in goat’s milk. The use of AFCF has been

temporarily authorized in the EU and in some other

countries. AFCF as a caesium binder is effective in

extensive production systems, in contrast to many

other countermeasures whose applicability is

limited. Boli are particularly favourable for infre-

quently handled free grazing animals, as the boli can

be administered when animals are gathered for

routine handling operations. For use in extensive

systems, the boli can be given a protective surface

coating of wax to delay the onset of AFCF release,

so that effectiveness is increased at the time when

animals are collected for slaughter [4.47]. It has

been estimated that the use of boli as a counter-

measure for sheep was 2.5 times as cost effective as

feeding with uncontaminated feed [4.48]. Salt licks

containing AFCF have also been used, but are less

effective [4.49].

Management regimes have been modified for

some animals in contaminated areas; for example,

slaughter times are modified to ensure that the

137

activity concentrations are relatively low. In the UK

the movement and slaughter of upland sheep are

restricted in some areas. The animals are monitored

to ensure that their

137

Cs activity concentrations are

below the action level before they are slaughtered.

Live monitoring of animal derived products

(monitoring of live animals and/or of milk and

tissues after slaughter) has been used to ensure that

countermeasures have been effective. The use of

monitoring is also important in maintaining public

confidence in the products from affected areas.

An example of the long term consequences of

the accident can be seen in Fig. 4.8, which shows the

number of reindeer in Sweden that had radio-

caesium activity concentrations above the action

level and the number of slaughtered animals. The

TABLE 4.5. SUMMARY OF THE REDUCTION FACTORS ACHIEVED WITH THE

DIFFERENT COUNTERMEASURES USED IN THE THREE COUNTRIES OF THE

FORMER USSR [4.30, 4.34, 4.40, 4.45]

Caesium-137 Strontium-90

Normal ploughing (first year) 2.5–4.0 —

Skim and burial ploughing 8–16 —

Liming 1.5–3.0 1.5–2.6

Application of mineral fertilizers 1.5–3.0 0.8–2.0

Application of organic fertilizers 1.5–2.0 1.2–1.5

Radical improvement:

First application

Further applications

1.5–9.0

2.0–3.0

1.5–3.5

1.5–2.0

Surface improvement:

First application

Further applications

2.0–3.0

1.5–2.0

2.0–2.5

1.5–2.0

Change in fodder crops 3–9 —

Clean feeding 2–5 (time dependent) 2–5

Administration of caesium binders 2–5 —

Processing milk to butter 4–6 5–10

Processing rapeseed to oil 250 600

For wet peat, up to 15 with drainage.

high number of slaughtered animals in the first year

was in part due to the low action level of 300 Bq/kg

fresh weight, which was subsequently increased to

1500 Bq/kg from 1987. The decline has been

achieved partly by extensive use of counter-

measures, including clean feeding and change of

slaughter time.

4.3.6. Current status of agricultural

countermeasures

Currently in all three countries of the former

USSR clean feeding remains an important counter-

measure to ensure that meat from intensive farms

can be marketed.

In Belarus, fertilization with phosphorus–

potassium is used on collective farms, and milk

above the action level from the farms is processed

into butter. Radical improvement is used on private

farms together with Prussian blue for milk.

Rapeseed production is currently limited by

processing capacity, although this may be increased

in the future.

In Ukraine the only remaining countermeasure

used on intensive systems is clean feeding of meat

producing animals prior to slaughter. Any milk

above the action level is used within the settlements,

partially to feed pigs. All other countermeasures are

directed at private farmers. These countermeasures

currently comprise the radical improvement of

meadows and the use of clay mineral caesium

binders for privately produced milk.

In the Russian Federation, fertilizers (largely

potassium) are supplied to intensive farms. For

private farms, Prussian blue is provided for

privately produced milk and, on request, for

privately produced meat intended for market.

In all contaminated settlements a service for

the monitoring of local produce exists, although the

capacity and availability of the service varies.

In western Europe countermeasures for

animals in extensive systems are still used in

Norway and Sweden, and the movement and

slaughter of upland sheep are still restricted in

certain areas of the UK.

4.3.7. A wider perspective on remediation,

including socioeconomic issues

Experience after the Chernobyl accident has

shown that account has to be taken, in developing

restoration strategies, of a wide range of different

issues to ensure the long term sustainability of large

and varied types of contaminated areas [4.51]. The

selection of robust and practicable restoration

strategies should take into account not only radio-

logical criteria but also: (a) practicability, including

effectiveness, technical feasibility and the accepta-

bility of the countermeasure; (b) cost–benefit; (c)

ethical and environmental considerations; (d)

requirements for effective public communication;

(e) the spatial variation in many of these factors;

and (f) the contrasting needs of people in urban,

rural and industrial environments [4.52]. When not

only radiological factors but also social and

economic factors are taken into account, better

acceptability of countermeasures by the public can

be achieved.

A number of European Commission (EC) and

United Nations projects have applied some of the

above considerations in trying to provide

appropriate information to, and interaction with,

people in contaminated territories and in involving

them in making decisions about responding to

enhanced radiation doses and about ways of living

sustainably in contaminated areas. In particular, this

introduces the possibility of self-help and the

opportunity for people to decide for themselves

whether they wish to modify their behaviour to

reduce their doses. The EC ETHOS project [4.53,

4.54] identified the dissemination of a practical

radiological culture within all segments of the

population as a prerequisite, especially for profes-

sionals in charge of public health. The EC Tacis

Programme ENVREG project [4.55] in Belarus and

Ukraine sought to minimize the environmental and

secondary medical effects resulting from the

20 000

40 000

60 000

80 000

100 000

120 000

88/78

09/98

29/19

39/2991

59/4991

79/6991

99/89

10/00

20/1002

30/20

40/3002

Year (July–June)

Total slaughter

Part of slaughter above the action level

Number of slaughtered reindeer per year

FIG. 4.8. Change with time in the number of reindeer in

Sweden with radiocaesium activity concentrations above

the action level and in the number of slaughtered animals

[4.50].

Chernobyl accident by improving the public

perception and awareness of these effects.

Most recently, the EC CORE project [4.56]

was initiated to address long term rehabilitation and

sustainable development in the Bragin, Chechersk,

Slavgorod and Stolin areas of Belarus. CORE

community projects include health care, radio-

logical safety, information and education. In

addition, critical socioeconomic constraints are

being addressed, specifically using a crediting

system for small businesses and farmers, the cost

effective production of ‘clean’ products, the

creation of a rural entrepreneurs’ centre and the

promotion of community economic initiatives.

The Chernobyl debate is increasingly about

socioeconomic issues and the communication of

technical information in an understandable way.

The ETHOS, ENVREG and CORE projects all

have a strong community focus and target

Chernobyl affected communities and other local

stakeholders. Feedback from the communities

should indicate which approaches are proving

successful and to what extent. The holistic

philosophy of these projects of considering both

environmental and social problems is in line with

the recent United Nations initiative known as

Strategy for Recovery [4.57].

4.3.8. Current status and future of abandoned

land

In this section the extent of recovery of

abandoned land is summarized for each of the three

countries of the former USSR. In 2004, 16 100,

11 000 and 6095 ha of previously abandoned land in

Belarus, the Russian Federation and Ukraine,

respectively, were returned to economic use [4.26].

In general, there is currently little effort being

devoted to any further rehabilitation of abandoned

areas.

4.3.8.1. Exclusion and resettlement zones in Belarus

The CEZ covers a total of 215 000 ha in

Belarus. The people who used to reside there were

evacuated in 1986. Since May 1986, lands in the

CEZ have been taken out of agricultural and other

production. The Polessye State Radioecological

Reserve (PSRR) was set up by a government decree

in 1988 and comprises mainly the CEZ but also

includes some other areas with high levels of trans-

uranium radionuclide contamination. Access to the

PSRR is forbidden and very few, mostly old, people

are currently present, without permission, in the

area. Pursuant to the law on the legal regime of

territories contaminated as a result of the

Chernobyl nuclear power plant accident [4.58],

most of the land in the CEZ cannot be brought back

into economic production within a millennium,

because of contamination with long lived trans-

uranium radionuclides. In the CEZ only activities

related to ensuring radiation safety, fighting forest

fires, preventing the transfer of radioactive

substances, protecting the environment and

scientific research and experimental work are

permitted.

While the CEZ (i.e. the Bragin, Khoiniki and

Narovlya areas of the Gomel region) is the most

contaminated area adjacent to the Chernobyl

nuclear power plant, a further resettlement zone

was identified in the early 1990s from which more

people were evacuated; this zone covers a total area

of 450 000 ha.

A total area of agricultural land of 265 000 ha

received a deposition of

137

Cs at levels in excess of

1480 kBq/m

and/or of

Sr in excess of 111 kBq/m

and/or of plutonium isotopes in excess of

3.7 kBq/m

. All this land is excluded from agricul-

tural use.

The remaining abandoned agricultural land in

the resettlement zone could be used for agriculture

in the future. The present state of the ecosystems

and the economic infrastructure of the resettlement

zone are characterized by a general deterioration in

the former agricultural lands, drainage systems and

roads. Due to lack of drainage, there has also been a

gradual elevation in the water table. Normal

ecological succession has led to an increase in the

number of perennial weeds and shrubs. Unlike in

the CEZ, in the resettlement zone limited access for

certain maintenance activities, such as the activities

needed to maintain roads, electricity transmission

lines, etc., is permitted.

In Belarus it is considered to be important to

bring lands back to agricultural use, if possible. At

the request of collective and State farms, if

supported by local authorities, surveys of former

agricultural lands were conducted to determine

whether it is possible to rehabilitate the land for

agricultural use. This was based on radiological

considerations only.

By 2001 a total of 14 600 ha of previously

withdrawn land had been returned to use [4.34], and

recently this has been increased to about 16 000 ha.

This land is closely adjacent to populated

settlements. In these rehabilitated sites the soil

fertility has been restored and a variety of counter-

measures has been used to minimize radiocaesium

and radiostrontium uptake based on official

guidelines [4.37].

Most of the agricultural and other land of the

resettlement zone was transferred to the authority

of the Ministry of Forestry. This is because much of

the resettlement zone is considered suitable for

forest production.

According to an assessment by Bogdevitch et

al. [4.59], a total of about 35 000 ha of the more

fertile agricultural land may be suitable for further

rehabilitation. However, economic support for

recovery and use of countermeasures has declined

significantly over recent years. Use of counter-

measures is now confined to radical improvement of

meadows, feeding of Prussian blue to cattle, liming

and fertilization.

Methodologies for the rehabilitation of

abandoned land are being developed and improved,

in particular with respect to economic evaluation.

The main obstacles to the renewed agricultural use

of abandoned land are the destroyed infrastructure,

the high production cost and the low market

demand for the agricultural goods. A large scale

rehabilitation of excluded land will only be possible

if there is a general improvement in the economic

situation of the country.

4.3.8.2. Rehabilitation of contaminated lands in

Ukraine

The first priority was the rehabilitation of land

on which people are living. Consideration has since

been given to the potential rehabilitation of

abandoned areas. Such areas can be rehabilitated if

this procedure is expedient with respect to

economic and social criteria. The main condition for

human occupancy of such areas without restrictions

is that the additional annual effective dose should

not exceed 1 mSv.

The efficiency of countermeasures is

determined by the following criteria:

(a) Radiological: reduction of radionuclide

content in local products and in the associated

individual and collective dose.

(b) Economic: increased product market value.

given countermeasure.

In 2004, on the basis of radiological criteria

alone, a significant part of the abandoned agricul-

tural lands (more than 70%) could be returned to

economic use. When economic and social criteria

are assessed, the amount of land that could be

rehabilitated declines (see Table 4.6). Table 4.6

TABLE 4.6. REHABILITATION OF ZONES OF OBLIGATORY

RESETTLEMENT (OUTSIDE THE CHERNOBYL EXCLUSION ZONE)

Area Abandoned land (ha)

Can be rehabilitated judged

on radiological, economic

and social criteria (ha)

Kiev region

1998–2000 (done) — 3475

2001–2005 — 4720

Total 29 342 8205

Zhytomyr region

1998–2000 (done) — 2620

2001–2005 — 4960

Total 71 943 7580

Provided by Forum participants from official national sources.

shows a scheme for rehabilitation based on

technical criteria over a seven year period The first

phase, from 1998 until 2000, was implemented, but

that for the second phase was not, due to changing

economic and social conditions.

In the CEZ, the limiting radionuclide is now

Sr rather than

137

Cs. On the basis of radiological

considerations, the south-west part of the zone can

be used without restrictions. However, in reality,

legal restrictions, the lack of a suitable infra-

structure and consideration of economic and social–

psychological factors prevent its rehabilitation.

The same restrictions apply to the other

abandoned areas, where legal restrictions are also in

force that, together with deteriorating economic

conditions, currently prevent the application of

countermeasures in the remaining identified

abandoned areas. The pressure to bring the

abandoned land back into production is also

reduced by the current abundance of agriculturally

productive land in Ukraine and the presence in

southern Ukraine of land that is much more

productive.

Some people have returned to abandoned

areas to live, and others live outside them but use

the land for agricultural activities such as hay

production. Countermeasures are not being applied

in the abandoned areas, but there is sanitary and

regulatory control of these activities.

4.3.8.3. Abandoned zones in the Russian Federation

Areas in the Russian Federation with high

levels of radioactive soil contamination were

abandoned in stages from 1986 until 1989, and in

total 17 000 ha of agricultural land was excluded

from economic use. The abandoned areas belonged

to 17 rural settlements with about 3000 inhabitants

(at the time of the accident) and 12 collective farms.

In 1987–1989 considerable efforts were made

to keep the highly contaminated areas in economic

use, and hence most of the abandoned areas were

subjected to intensive countermeasure application.

However, these efforts were only partially

successful, and the land was gradually abandoned;

in the 1990s, the intensity of countermeasure

application was reduced. Overall, about 11 000 ha

was returned to agricultural use by 1995. These

decisions to return land to agricultural use were

made individually for each contaminated field.

Special attention was paid to highly contaminated

fields surrounded by fields with relatively low levels

of contamination, because there was a natural

inclination to use these fields. The assessments were

based on Russian radiation safety standards,

including standards governing the quality of agricul-

tural products (TPL-93) [4.60].

Between 1995 and 2004 there was no further

rehabilitation of the abandoned areas. Officially

they are abandoned but, unofficially, some local

people are living in these areas and using them for

agricultural production, but without the benefit of

countermeasures.

Recently, a technical project of gradual

rehabilitation of the remaining abandoned areas, in

which the mean

137

Cs soil deposition varies from

1540 to 3500 kBq/m

, has been proposed by the

Russian Institute of Agricultural Radiology and

Agroecology. The criteria for agricultural

production include ensuring that

137

Cs activity

concentrations would be less than the TPL, as well

as a requirement that application of counter-

measures for each contaminated field would be

optimized. During the first planned stage, up to

2015, it is proposed to produce grain and potatoes

using agricultural workers who live elsewhere but

would come into the contaminated area as

necessary. Soil based countermeasures (liming,

potassium fertilization) should allow the production

of plant products with sufficiently low levels of

137

on most of the abandoned area. From 2015 the

implementation of animal breeding is planned, and

from 2025 the re-establishment of populated

settlements could commence. Thus by 2045 all

abandoned land could be used once more, although

application of different countermeasures would be

needed up to 2055 to ensure that annual doses to

the local inhabitants were less than 1 mSv.

4.4. FOREST COUNTERMEASURES

Countermeasures for forested areas contami-

nated with radionuclides are only likely to be

implemented if they can be accepted by foresters or

landowners on a practical basis (i.e. actions are

likely to fit in with normal forest management

practices). For countermeasures to be successful

they must also be accepted by the general public. As

forest countermeasures are labour consuming and

expensive, they cannot be implemented quickly and

must be planned carefully. They are likely to be long

term activities and their beneficial effects take time

to be realized.

4.4.1. Studies on forest countermeasures

Generally, prior to the Chernobyl accident

countermeasures to offset doses due to large scale

contamination of forests had not been given very

much attention. Several international projects in

the 1990s gave rise to a number of publications in

which suggestions and recommendations were

made for possible countermeasures to be applied

in forests [4.61–4.64]. However, in the three

countries of the former USSR, actions had already

been taken to restrict activities in the more

contaminated zones, which included significant

areas of forestry [4.65]. These actions were, in

general, rather simple and involved restrictions on

basic activities such as accessing forests and

gathering wild foods and firewood. A major

question remains as to whether any more complex

or technologically based countermeasures can be

applied in practice, and whether the ideas

developed by researchers will remain as

theoretical possibilities rather than as methods

that can be applied in real forests on a realistic

scale. The following section describes some of the

more feasible countermeasures that have been

devised for forests contaminated with radio-

caesium. This is illustrated in Section 4.4.3 by

studies in which countermeasures have actually

been put into practice.

4.4.2. Countermeasures for forests contaminated

with radiocaesium

There are several categories of counter-

measure that are, in principle, applicable to forest

ecosystems [4.66, 4.67]. A selection of these is

shown in Table 4.7. These can be broadly

categorized into (a) management based and (b)

technology based countermeasures.

4.4.2.1. Management based countermeasures

Under the broad heading of management

based countermeasures, the principal remedial

methods applied after the Chernobyl accident

involved restrictions on various activities normally

carried out in forests. Restriction of access to

contaminated forests and restriction of the use of

forest products were the main countermeasures

applied in the USSR and later in the three

independent countries [4.65]. These restrictions can

be categorized as follows:

(a) Restricted access, including restrictions on

public and forestworker access. This has been

assisted by the provision of information from

local monitoring programmes and education

on issues such as food preparation [4.65].

(b) Restricted harvesting of food products by the

public. The most commonly obtained food

products include game, berries and

mushrooms. The relative importance of these

varies from country to country. In the three

countries of the former USSR, mushrooms are

particularly important and can often be

severely contaminated.

public. This not only exposes people to in situ

gamma radiation while collecting firewood

but can also lead to further exposures in the

home and garden when the wood is burned

and the ash is disposed of, sometimes being

used as a fertilizer.

(d) Alteration of hunting practices. The

consumption of fungi by animals such as roe

deer leads to strong seasonal trends in their

body content of radiocaesium (see Section

3.3). Thus excessive exposures can be avoided

by eating the meat only in seasons in which

fungi are not available as a food source for the

animals.

(e) Fire prevention is a fundamentally important

part of forest management under any circum-

stances, but it is also important after a large

scale deposition to avoid secondary contami-

nation of the environment, which could result

from burning of trees and especially forest

litter, which is one of the major repositories of

radiocaesium in the forest system (see Section

3.3). One of the ways in which forest fires can

be avoided is by minimizing human presence

in the forest, so this countermeasure is

strongly linked to restricting access, as

described above.

4.4.2.2.

Technology based countermeasures

This category of countermeasures includes the

use of machinery and/or chemical treatments to

alter the distribution or transfer of radiocaesium in

forests. Many mechanical operations are carried out

as part of normal forestry practice; examples of

these have been described by Hubbard et al. [4.69]

with reference to their use as countermeasures.

Similarly, applications of fertilizers and pesticides

may be made at different times in the forest

TABLE 4.7. SELECTED COUNTERMEASURES THAT HAVE BEEN CONSIDERED FOR APPLICA-

TION IN CONTAMINATED FORESTS [4.68]

Countermeasure Category Caveat Benefit Cost

Normal operation Management — No loss of productivity

or amenity

No dose reduction and

negative social costs

Minimum management:

forest fire protection,

disease protection and

necessary hunting

Management — Creation of nature

reserve and reduced

worker dose

Worker dose, loss of

productivity, negative

social costs and costs for

hunting

Delayed cutting of

mature trees

Management/

agrotechnical

Marginal feasibility Reduced

contamination of wood

due to:

— Radioactive decay

— Fixation of caesium

in soil

— Loss from soil and

wood

Delay in revenue

Early clear cutting and

replanting or self-

regeneration

Management/

agrotechnical

Must consider tree age

at time of

contamination;

possibly in combination

with soil mixing

Reduced tree

contamination:

— Lower soil–tree

transfer

— Delayed harvest

time

— Alternative tree

crop

Higher dose to workers

during replanting and

operational costs

Soil improvement:

harrowing after

thinning or clear

cutting

Agrotechnical Cost effectiveness is

dependent on the area

to be treated; possibly

in combination with

fertilizer application

Improved tree growth,

therefore growth

dilution; dilutes

radionuclide activity

concentrations in the

soil surface layer and

decreases them in

mushrooms, berries

and understorey game

Operational costs,

worker doses and

environmental or

ecological costs (e.g.

nitrate and other

nutrients lost)

Application of

phosphorus–potassium

fertilizer and/or liming

Agrotechnical Phosphorus–

potassium: may only be

effective for caesium,

especially effective for

younger stands

Lime: particularly

useful for

Reduction of uptake to

trees, herbs, etc., maybe

better growth and

dilution effect and

higher fixation

Cost of fertilizer, worker

dose and negative

ecological effects

Limiting public access Management Note: people normally

residing in forests not

considered

Reduction in dose,

possible increase in

public confidence

Loss of amenity/social

value, loss of food and

negative social impacts

Salt licks Agrotechnical — Reduction in caesium

uptake by grazing

animals

Continuing cost of

providing licks

Ban on hunting Management — Reduction in dose due

to ingestion of game

Need to find alternative

supply of meat

Ban on mushroom

collection

Management — Reduction in internal

dose

Need to find alternative

mushroom supply

cropping cycle as part of normal management

practice. However, the cost effectiveness of many

technological countermeasures is questionable,

especially when applied on a large scale [4.68]. Thus

it is to be expected that such countermeasures will

be restricted to small scale cases only, if they are

feasible at all. Such cases might include small areas

of urban woodland, such as parkland, which are

likely to be visited by many more people than

extensive and remote forest areas.

Technological countermeasures might include

mechanical removal of leaf litter or scraping of soil

layers, clear cutting and ploughing, and the

application of calcium- and potassium-containing

fertilizers. It is evident, however, that any of these

methods can damage the ecological functioning of

the forest when applied outside of the normal

schedule of forestry operations. This, and the high

economic costs of such operations, means that the

practical use of such techniques as countermeasures

remains largely speculative, and such measures have

not been applied after the Chernobyl accident other

than in small scale experiments. Indeed, the results

of cost–benefit calculations indicate that the

management options likely to result in the least

overall detriment are those which limit access and

consumption of forest foods. Options that involve

technological intervention, application of chemicals

or altering the harvesting patterns in forests are

unlikely to be used in practice.

4.4.3. Examples of forest countermeasures

Case studies in which forest countermeasures,

particularly technology based countermeasures,

have actually been applied in practice are rare. This

illustrates the difficulty of implementing practical

remedial measures in forests, in contrast to

agriculture, in which the application of fertilizers, in

particular, has been used with some success (see

Section 4.3). In practice, restrictive counter-

measures were applied in the USSR, and later in the

three independent countries, as well as in a limited

number of other countries, such as Sweden.

In the Bryansk region of the Russian

Federation, individual restrictions on forestry and

on the population living near forests were

recommended according to the level of

137

deposition. For forests receiving depositions greater

than 1480 kBq/m

, access was only allowed for

forest conservation, fire fighting and control of

pests and diseases. All forestry activity was stopped,

and public access, including for collection of forest

plants, was prohibited. In forests receiving

depositions between 555 and 1480 kBq/m

collection of forest products was also prohibited,

but limited forestry activities continued. At

deposition levels between 185 and 555 kBq/m

harvesting of trees was continued on the basis of

radiological surveys that were used to identify

individual areas in which external doses to forestry

workers and contamination of wood were

acceptable. However, the collection of berries and

mushrooms by the public was only permitted in

forests with deposition levels less than 74 kBq/m

One of the major effects of the restrictions

that were enforced on a large scale up to 1990 was a

negative impact on rural populations. At the

beginning of 1990 the population began gathering

mushrooms and berries again over the whole

Bryansk region. However, in areas where the

original

137

Cs deposition was between 555 and

1480 kBq/m

, restrictions on gathering forest food

products are still in force. This example illustrates a

major difficulty in implementing countermeasures

involving restrictions on public activities that

inevitably lead to a disturbance of normal societal

behaviour patterns. Furthermore, wood production

is still under the official control of local forest

authorities [4.65]; the currently applicable

permissible levels for contamination of wood and

forest products in the Russian Federation are shown

in Table 4.8. Similar restrictions and permissible

levels have been implemented in different regions

of Belarus, notably the Gomel and Mogilev regions.

The use of caesium binders, particularly

Prussian blue, in domestic animals has been one of

the more effective techniques used to reduce doses

from contaminated forests in the three countries of

the former USSR. The principles underlying this

method are described in Section 4.3; they are

equally applicable to the problem of marginal

grazing of domestic animals in forests. Typically,

reductions in

137

Cs activity concentrations of a

factor of five in milk and a factor of three in meat

can be achieved at optimum dosage [4.65].

One example of intervention in normal forest

related practices in countries outside the former

USSR is the case of roe deer hunting in Sweden. In

1988 the average muscle content of roe deer shot in

the autumn was 12 000 Bq/kg in the Gävle area. The

intervention level for such foodstuffs in Sweden is

1500 Bq/kg. Such high levels of contamination of

roe deer meat were due to the preferential

consumption of fungi by the deer in the autumn. As

a result of experiments, the Swedish authorities

recommended a change of hunting season for roe

deer to the spring; this change was applied

voluntarily by the hunting community in the early

1990s. As a result, the radiocaesium content in roe

deer meat in Gävle was reduced by approximately

six times. The recommendation to shift the hunting

season to the spring has remained in place until the

present day [4.71].

In addition, the management of reindeer by

the Sami people in northern Sweden has been

altered in a variety of ways to help reduce the radio-

caesium content of animals before slaughter. This

includes provision of clean fodder for sufficient time

to reduce the body burden below the intervention

level. A similar result can be achieved by altering

the time of slaughter, sometimes in combination

with feeding of clean fodder [4.72].

4.5. AQUATIC COUNTERMEASURES

There are a number of different intervention

measures that can be employed following fallout of

radioactive material to reduce doses to the public

via the surface water pathway. These actions may be

grouped into two main categories: those aimed at

reducing doses from radionuclides in drinking water

and those aimed at reducing doses from the

consumption of contaminated aquatic foodstuffs.

In the context of the atmospheric fallout of

radionuclides to both terrestrial and aquatic systems,

it has been shown [4.73–4.75] that doses from terres-

trial foodstuffs are in general much more significant

than doses from drinking water and aquatic

foodstuffs. However, in the Dnieper River system,

the river water transported radionuclides to areas

that were not significantly contaminated by atmos-

pheric fallout. This created a significant stress in the

population and a demand to reduce radionuclide

fluxes from the affected zone via the aquatic system.

Many remediation measures were put in place but,

because actions were not taken on the basis of dose

reduction, most of these measures were ineffective.

Moreover, radiation exposures of workers imple-

menting these countermeasures were high.

Measures to reduce doses via drinking water

may, however, be required, particularly in the short

term (timescale of weeks) after fallout, when

activity concentrations in surface waters are

relatively high. Owing to the importance of short

lived radionuclides, early intervention measures,

particularly the changing of supplies, can signifi-

cantly reduce radiation doses to the population.

Measures to reduce doses due to freshwater

foodstuffs may be required over longer timescales

as a result of the bioaccumulation of radionuclides

in the aquatic food chain.

Reviews of aquatic countermeasures (e.g.

Refs [4.76–4.79]) have considered both direct

(restrictions) and indirect intervention measures to

reduce doses:

(a) Restrictions on water use or changing to

alternative supplies;

(b) Restrictions on fish consumption;

drainage systems);

(d) Reduction of uptake by fish and aquatic

foodstuffs from contaminated water;

(e) Preparation of fish prior to consumption.

There is no evidence that countermeasures

were required, or applied, in marine systems after

the Chernobyl accident.

4.5.1. Measures to reduce doses at the water

supply and treatment stage

Restrictions were placed on the use of water

from the Dnieper River for the first year after the

TABLE 4.8. TEMPORARY PERMISSIBLE

LEVELS FOR CAESIUM-137 IN WOOD AND

FOREST FOOD PRODUCTS IN THE RUSSIAN

FEDERATION [4.70]

TPL (Bq/kg)

Round wood, including bark 11 100

Unsawn timber with bark removed 3 100

Sawn wood (planks) 3 100

Construction wood 370

Wood used for pulp and paper

production

3 100

Wood products for household use and

industrial processing

2 200

Wood products for packing and food

storage

1 850

Firewood 1 400

Mushrooms and berries (fresh weight) 1 480

Mushrooms and berries (dry weight) 7 400

Medical plants and medical raw material 7 400

Seeds of trees and bushes 7 400

accident. Abstraction of drinking water for Kiev

was switched to the Desna River with use of a

pipeline built during the first weeks after the

accident. A summary of the measures taken by the

Ukrainian authorities to switch to alternative

supplies from less contaminated rivers and from

groundwater can be found in Refs [4.76, 4.79].

Radionuclides may be removed from drinking

water supplies during the water treatment process.

Suspended particles are removed during water

treatment, and filtration can remove dissolved

radionuclides. In the Dnieper waterworks station,

activated charcoal and zeolite were added to water

filtration systems. It was found that activated

charcoal was effective in removing

131

I and

106

Ru,

and zeolite was effective in removing

137

Cs,

134

and

Sr. These sorbents were effective for the first

three months, after which they became saturated

and their efficiency declined [4.80, 4.81]. The

average removal of these radionuclides from water

(dissolved phase) was up to a factor of two.

After the accident, the upper gates of the Kiev

reservoir dam were opened to release surface water.

It was believed at the time that the surface water

was relatively low in radionuclide content, because

suspended particles had sunk to deeper waters.

Therefore, the release of water would allow room in

the reservoir to contain runoff water from the

inflowing rivers, which was believed to be highly

contaminated. In fact, because of direct

atmospheric deposition to the reservoir surface, the

surface waters in the reservoir were much more

contaminated than the deep waters. As noted by

Voitsekhovitch et al. [4.80], “a better approach to

lowering the water level within the Kiev reservoir

would have been to open the bottom dam gates and

close the surface gates. This would have reduced the

levels of radioactivity in downstream drinking water

in the first weeks after the accident.” Although this

countermeasure was not efficiently implemented

after the Chernobyl accident, regulation of flow,

given the correct information on contamination,

could effectively reduce activity concentrations in

drinking water, as it takes some time (days or more)

for lakes and reservoirs to become fully mixed.

In a large river–reservoir system such as the

Dnieper, control of water flows in the system can

significantly reduce transfers of radioactive material

downstream [4.82]. In the Dnieper River, the time it

takes for water to travel from the Kiev reservoir to

the Black Sea varies between three and ten months.

Over the time that the water takes to travel

downstream, radioactive pollution is reduced by

decay of short lived radionuclides and transfers to

reservoir bed sediments (particularly of radio-

caesium) [4.82].

4.5.2. Measures to reduce direct and secondary

contamination of surface waters

Standard antisoil erosion measures can be

used to reduce runoff of radionuclides attached to

soil particles. Note, however, that typically less than

50% of radiocaesium and less than 10% of radio-

strontium and radioiodine were in the particulate

phase, and this limits the potential effectiveness of

this countermeasure. It should also be noted that

the dissolved, rather than particulate, form of these

radionuclides is important in determining activity

concentrations in drinking water and freshwater

biota.

Dredging of canal bed traps to intercept

suspended particles in contaminated rivers was

carried out in the Pripyat River [4.79]. These canal

bed traps were found to be highly inefficient for two

reasons: (a) the flow rates were too high to trap the

small suspended particles carrying much of the

radionuclide contamination; and (b) a significant

proportion of the radionuclide activity (and most of

the ‘available’ activity) was in dissolved form and

thus would not have been intercepted by the

sediment traps.

One hundred and thirty zeolite-containing

dykes were constructed on smaller rivers and

streams around Chernobyl in order to intercept

dissolved radionuclides. These were found to be

very ineffective: only 5–10% of the

Sr and

137

Cs in

the small rivers and streams was adsorbed by these

zeolite barriers [4.80]. In addition, the rivers and

streams on which they were placed were later found

to contribute only a few per cent to the total radio-

nuclide load in the Pripyat–Dnieper system.

After the Chernobyl accident, spring flooding

of the highly contaminated Pripyat floodplain

resulted in increases in

Sr activity concentrations

in the Pripyat River from annual average activity

concentrations of around 1 Bq/L to a maximum of

around 8 Bq/L for a flood event covering an approx-

imately two week period [4.83]. In 1993 a dyke was

constructed around the highly contaminated

floodplain on the left bank of the Pripyat. This

prevented flooding of this area and proved effective

in reducing

Sr wash-off to the river during flood

events [4.80]. A second dyke was constructed on the

right bank of the Pripyat in 1999. The annual

average

Sr activity concentration in Kiev reservoir

water, however, was below 1 Bq/L in all years from

1987 onwards. The radiological significance of the

Sr activity concentrations in Kiev reservoir water,

even during the short flood events, is therefore very

low, although it has been argued that the averted

collective dose to the large number of users of the

river–reservoir system is significant.

It is potentially possible to increase the

sedimentation of radionuclides from lakes and

reservoirs by the introduction of a strongly sorbing

material such as a zeolite or an (uncontaminated)

mineral soil. This method has not been tested. Using

a model for the removal of radiocaesium from lakes

by settling of suspended particles, Smith et al. [4.78]

identified two problems with this method: (a) large,

deep lakes would require extremely large amounts

of sorbent; and (b) secondary contamination of the

lake by remobilization of activity from the

catchment and/or bottom sediments would require

repeat applications in most systems.

4.5.3. Measures to reduce uptake by fish and

aquatic foodstuffs

Bans on the consumption of freshwater fish

have been applied in the limited zones affected by

the Chernobyl accident [4.84]. In some areas,

selective bans on the more contaminated predatory

fish have been applied. It is believed that such bans

are often ignored by fishermen. Bans on the sale of

freshwater fish were applied in some areas of

Norway [4.85]. Farmed fish could be used as an

alternative source of freshwater fish in areas

affected by fishing bans, since farmed fish fed with

uncontaminated food do not accumulate radionu-

clides significantly [4.86].

The addition of lime to reduce radionuclide

levels in fish was tested in 18 Swedish lakes [4.87].

The results of the experiments showed that liming

had no significant effect on the uptake of

137

Cs in

fish in comparison with control lakes. Although the

uptake of

Sr was not studied in these experiments,

it is expected that increased calcium concentration

in lakes may have an effect on the

Sr concen-

tration in fish. Experience of lake liming, in

conjunction with artificial feeding of fish in

Ukraine, has been summarized by Voitsekhovitch

[4.79].

It is known that the concentration factor for

radiocaesium in fish is inversely related to the

potassium content of the surrounding water. After

the Chernobyl accident, potassium was added to 13

lakes in Sweden, either as potash or as an additive in

mixed lime [4.87]. The results of the potash

treatment were somewhat inconclusive, with a small

reduction in activity concentrations in perch fry

observed during the two year experiment. It was

found that in lakes with short water retention times

it was difficult to maintain high levels of K

in the

lake.

In an experiment on Lake Svyatoe (a closed

lake) in Belarus, Kudelsky et al. [4.88, 4.89] added

potassium chloride fertilizer on to the frozen lake

surface. Results showed a significant (factor of

three) overall reduction in

137

Cs concentration in

fish during the first years after the experiment.

However, as expected, the

137

Cs in the water

increased by a factor of two to three after the

countermeasure application. It is likely that

potassium treatment is only feasible in lakes with

very long water residence times, which allow

increased potassium concentrations to be

maintained. Also, the increased

137

Cs in water is

unlikely to be acceptable in lakes that have water

abstracted for drinking.

Manipulation of the aquatic food web by

intensive fishing was carried out in four lakes in

Sweden [4.87], and as a complementary measure in

an additional three lakes. This resulted in a

reduction of the fish population by about 5–10 kg/

ha. The species reduced were mainly pike, perch

and roach. No effect of intensive fishing on

137

concentrations in fish was observed. Fertilization

was carried out in two Swedish lakes using

Osmocoat (5% phosphorus and 15% nitrogen). The

concentrations of total phosphorus generally

showed no change in the long term mean value: it

appears that the fertilization treatment was not

carried out sufficiently effectively. No effect was

observed on

137

Cs activity concentrations in fish.

Different methods of food preparation may

affect the quantity of radionuclides in consumed

food [4.90]. Ryabov suggested bans on the

consumption of smoked and dried fish, because

these processes increase concentrations of radionu-

clides (per unit of weight consumed) [4.84]. Other

preparation processes may reduce radionuclide

levels in fish by approximately a factor of two. An

effective measure to reduce the consumption of

radiostrontium is to remove the bony parts of fish

prior to cooking, since strontium is mainly concen-

trated in the bones and skin. Various other food

preparation methods are discussed in Ref. [4.91].

4.5.4. Countermeasures for groundwater

There is no evidence that measures have ever

been taken to protect groundwater supplies after an

atmospheric deposition of radioactivity.

Groundwater residence times are long enough that

shorter lived radionuclides such as

131

I will have

decayed long before they affect drinking water.

Only very small amounts of radiostrontium and

radiocaesium percolate from surface soils to

groundwater after atmospheric deposition. A study

[4.77] has shown that, after the Chernobyl

accident, exposure to

Sr and

137

Cs via the

groundwater pathway was insignificant in

comparison with other pathways (food, external

exposure, etc.).

Measures were taken to protect groundwater

from seepage of radionuclides from the shelter and

from radioactive waste sites in the CEZ. These

measures focused mainly on the construction of

engineering and geochemical barriers around the

local hot spots to reduce groundwater fluxes to the

river network. Actions to stop precipitation from

entering the shelter, and drainage of rainwater

collected in the bottom rooms of the shelter, have

also to be considered as preventive measures to

reduce groundwater contamination around the

Chernobyl nuclear power plant industrial site.

4.5.5. Countermeasures for irrigation water

As discussed previously, irrigation did not add

significantly to the radionuclide contamination of

crops that had previously been affected by the

atmospheric deposition of radionuclides. Thus, in

practice, no countermeasures were directly applied

to irrigation waters. However, the experience

described in Ref. [4.79] shows that the change from

sprinkling to drainage irrigation of agricultural

plants (e.g. vegetables) can reduce the transfer of

radionuclides from water to crops by several times.

This, in combination with improved fertilization of

irrigated lands, can effectively reduce radionuclide

levels in crops irrigated with water from reservoirs

affected by radioactive pollution.

4.6. CONCLUSIONS AND

RECOMMENDATIONS

The Chernobyl accident prompted the intro-

duction of an extensive set of short and long term

environmental countermeasures by the authorities

in the most affected countries to reduce its negative

consequences. Unfortunately, there was not always

openness and transparency towards the public, and

information was withheld. This can, in part, explain

some of the problems experienced later in commu-

nication with the public, and the mistrust of the

competent authorities. Similar behaviour in many

other countries outside Belarus, the Russian

Federation and Ukraine led to a distrust in

authority that, in many countries, has prompted

investigations on how to deal with such major

accidents in an open and transparent way and on

how the affected people can be involved in decision

making processes.

The unique experience of countermeasure

application after the Chernobyl accident has

already been widely used both at the national and

international levels in order to improve prepar-

edness against future nuclear and radiological

emergencies [4.12, 4.14, 4.41, 4.91, 4.92].

4.6.1. Conclusions

(a) The Chernobyl accident prompted the intro-

duction of an extensive set of short and long

term environmental countermeasures by the

USSR and, later, independent country

authorities, aimed at reducing the accident’s

negative consequences. The countermeasures

involved large amounts of human, economic

and scientific resources.

(b) When social and economic factors along with

the radiological factors are taken into account

during the planning and application of

countermeasures, better acceptability of these

measures by the public is achieved.

quences of the Chernobyl accident required

the development of some additional national

and international radiation safety standards,

to take account of changes of radiation

exposure conditions.

(d) Countermeasures applied in the early phase of

the Chernobyl accident were only partially

effective in reducing radioiodine intake via

milk, because of the lack of timely information

about the accident and advice on appropriate

actions, particularly for private farmers.

(e) The most effective countermeasures in the

early phase were exclusion of contaminated

pasture grasses from animal diets and

rejection of milk (with further processing)

based on radiation monitoring data. Feeding

animals with ‘clean’ fodder was effectively

performed in some affected countries. The

slaughtering of cattle was unjustified from a

radiological point of view and had great

hygienic, practical and economic implications.

(f) The greatest long term problem has been radio-

caesium contamination of milk and meat. In

the USSR, and later in the three independent

countries, this has been addressed by the

treatment of land used for fodder crops, clean

feeding and application of caesium binders to

animals, which enabled most farming practices

to continue in affected areas.

(g) Decontamination of settlements was widely

applied in contaminated regions of the USSR

during the first years after the Chernobyl

accident as a means of reducing the external

exposure of the public; this was cost effective

with regard to external dose reduction when

its planning and implementation were

preceded by a remediation assessment based

on cost–benefit considerations and external

dosimetry data.

(h) The decontamination of urban environments

has produced a considerable amount of low

level radioactive waste, which creates a

problem of disposal. However, secondary

contamination of cleaned up plots has not

been observed.

(i) The following forest related restrictions

widely applied in the USSR and later in the

three independent countries and partially in

Scandinavia have reduced human exposure

due to residence in radioactively contami-

nated forests and the use of forest products:

(i) Restrictions on public and forest worker

access as a countermeasure against

external exposure.

(ii) Restricted harvesting by the public of

food products such as game, berries and

mushrooms contributed to a reduction in

internal dose. In the affected countries

mushrooms are a common dietary

component, and therefore this restriction

has been particularly important.

(iii) Restricted collection of firewood by the

public to prevent exposures in the home

and garden when the wood is burned and

the ash is disposed of or used as a

fertilizer.

(iv) Alteration of hunting practices, aimed at

avoiding consumption of meat with high

seasonal levels of radiocaesium.

(v) Fire prevention, especially in areas with

large scale radionuclide deposition, in

order to avoid secondary contamination

of the environment.

(j) Experience has shown that forest restrictions

can result in significant negative social conse-

quences, and advice from the authorities to

the general public may be ignored as a result.

This situation can be offset by the provision of

suitable educational programmes targeted at

the local scale to explain the purpose of the

suggested changes in the use of some forest

areas.

(k) It is unlikely that any technology based forest

countermeasures (i.e. the use of machinery

and/or chemical treatments to alter the distri-

bution or transfer of radiocaesium in forests)

will be practicable on a large scale.

(l) Numerous countermeasures put in place in

the months and years after the accident to

protect water systems from transfers of radio-

nuclides from contaminated soils were, in

general, ineffective and expensive and led to

relatively high exposures of the workers

implementing the countermeasures.

(m) The most effective countermeasure for

aquatic pathways was the early restriction of

drinking water abstraction and the change to

alternative supplies. Restrictions on the

consumption of freshwater fish have proved

effective in Scandinavia and Germany;

however, in Belarus, the Russian Federation

and Ukraine such restrictions may not always

have been adhered to.

(n) It is unlikely that any future countermeasures

to protect surface waters will be justifiable in

terms of economic cost per unit of dose

reduction. It is expected that restrictions on

the consumption of fish will be retained in a

few cases (in closed lakes) for several more

decades.

4.6.2. Recommendations

4.6.2.1. Countries affected by the Chernobyl

cident

(a) Long term remediation measures and

countermeasures in the areas contaminated

with radionuclides should be applied if they

are radiologically justified and optimized.

(b) Authorities and the general public should be

particularly informed on radiation risk factors

and the technological possibilities to reduce

them in the long term by means of

remediation and countermeasures. Local

authorities and the public should be involved

in related discussions and decision making.

remediation measures and regular counter-

measures should be maintained where they

remain efficient and justified — mainly in

agricultural areas with poor (sandy and peaty)

soils and resulting high radionuclide transfer

from soil to plants.

(d) Particular attention must be given to private

farms in several hundred settlements and to

about 50 intensive farms in Belarus, the

Russian Federation and Ukraine, where radio-

nuclide concentrations in milk still exceed the

national action levels.

(e) Emphasis should be on the most efficient long

term remediation measures; these are the

radical improvement of pastures and

grasslands and the draining of wet peaty areas.

The most efficient regular agricultural

countermeasures are the pre-slaughter clean

feeding of animals accompanied by in vivo

monitoring, the application of Prussian blue to

cattle and the enhanced application of mineral

fertilizers in plant cultivation.

(f) Restricting harvesting of wild food products

such as game, berries, mushrooms and fish

from closed lakes by the public still may be

needed in areas where their activity concen-

trations exceed the national action levels.

(g) Advice should continue to be given on

individual diets, as a way of reducing

consumption of highly contaminated wild

food products, and on simple cooking

procedures to remove radioactive caesium.

(h) It is necessary to identify sustainable ways to

make use of the most affected areas that

reflect the radiation hazard, but also to revive

their economic potential for the benefit of the

community.

4.6.2.2. Worldwide

(a) The unique experience of countermeasure

application after the Chernobyl accident

should be carefully documented and used for

the preparation of international guidance for

authorities and experts responsible for

radiation protection of the public and the

environment.

(a) Practically all the long term agricultural

countermeasures implemented on a large

scale on contaminated lands of the three most

affected countries can be recommended for

use in the event of future accidents. However,

the effectiveness of soil based

countermeasures varies at each site. Analysis

of soil properties and agricultural practices

before application is therefore of great

importance.

(b) Recommendations on the decontamination of

the urban environment in the event of large

scale radioactive contamination should be

distributed to the owners and operators of

nuclear facilities that have the potential for

substantial accidental radioactive release

(nuclear power plants and reprocessing

plants) and to authorities in adjacent regions.

4.6.2.3. Research

(a) Generally, the physical and chemical

processes involved in environmental counter-

measures and remediation technologies, both

of a mechanical nature (radionuclide removal,

mixing with soil, etc.) and a chemical nature

(soil liming, fertilization, etc.), are understood

sufficiently to be modelled and applied in

similar circumstances worldwide. Much less

understood are the biological processes that

could be used in environmental remediation

(e.g. reprofiling of agricultural production,

bioremediation, etc.). These processes require

more research.

(b) An important issue that requires more socio-

logical research is the perception by the public

of the introduction, performance and

withdrawal of countermeasures in the event of

an emergency, as well as the development of

social measures aimed at involving the public

in these processes at all stages, beginning with

the decision making process.

national and national radiological criteria and

safety standards applicable to the remediation

of areas affected by environmental contami-

nation with radionuclides. The experience of

radiological protection of the public after the

Chernobyl accident has clearly shown the

need for further international harmonization

of appropriate radiological criteria and safety

standards.

REFERENCES TO SECTION 4

[4.1] UNITED NATIONS, Sources, Effects and Risks

of Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(1988) 309–374.

[4.2] INTERNATIONAL ATOMIC ENERGY

AGENCY, Recovery Operations in the Event of a

Nuclear Accident or Radiological Emergency

(Proc. Symp. Vienna, 1989), IAEA, Vienna

(1990).

[4.3] INTERNATIONAL ADVISORY COMMIT-

TEE, International Chernobyl Project: Technical

Report, IAEA, Vienna (1991).

[4.4] INTERNATIONAL ATOMIC ENERGY

AGENCY, One Decade after Chernobyl:

Summing up the Consequences of the Accident

(Proc. Int. Conf. Vienna, 1996), IAEA, Vienna

(1996).

[4.5] EUROPEAN COMMISSION, Council Regula-

tion (EEC) No. 1707/86 OJ No. L 146 of 31 May

1986 (1986) 88–90.

[4.6] UNITED NATIONS, Sources and Effects of

Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(2000).

[4.7] INTERNATIONAL ATOMIC ENERGY

AGENCY, Present and Future Environmental

Impact of the Chernobyl Accident, IAEA-

TECDOC-1240, IAEA, Vienna (2001).

[4.8] INTERNATIONAL COMMISSION ON

RADIOLOGICAL PROTECTION, A

Framework for Assessing the Impact of Ionising

Radiation on Non-human Species, Pergamon

Press, Oxford and New York (2003).

[4.9] INTERNATIONAL COMMISSION ON

RADIOLOGICAL PROTECTION, Recommen-

dations of the International Commission on Radi-

ological Protection, Publication 26, Pergamon

Press, Oxford and New York (1977).

[4.10] INTERNATIONAL COMMISSION ON

RADIOLOGICAL PROTECTION, Protection

of the Public in the Event of Major Radiation

Accidents: Principles for Planning, Publication 40,

Pergamon Press, Oxford and New York (1984).

[4.11] INTERNATIONAL ATOMIC ENERGY

AGENCY, INTERNATIONAL LABOUR

ORGANISATION, OECD NUCLEAR

ENERGY AGENCY, WORLD HEALTH

ORGANIZATION, Basic Safety Standards for

Radiation Protection, Safety Series No. 9, IAEA,

Vienna (1982).

[4.12] CODEX ALIMENTARIUS COMMISSION,

Guideline Levels for Radionuclides in Foods

Following Accidental Nuclear Contamination for

Use in International Trade, Rep. CAC/GL 5-1989,

Codex Alimentarius Commission, Rome (1989).

[4.13] INTERNATIONAL COMMISSION ON

RADIOLOGICAL PROTECTION, 1990

Recommendations of the International Commis-

sion on Radiological Protection, Publication 60,

Pergamon Press, Oxford and New York (1991).

[4.14] INTERNATIONAL COMMISSION ON

RADIOLOGICAL PROTECTION, Principles

for Intervention for Protection of the Public in a

Radiological Emergency, Publication 63,

Pergamon Press, Oxford and New York (1993).

[4.15] INTERNATIONAL COMMISSION ON

RADIOLOGICAL PROTECTION, Protection

of the Public in Situations of Prolonged Radiation

Exposure, ICRP Publication 82, Pergamon Press,

Oxford and New York (1999).

[4.16] MINISTRY OF HEALTH, Standards of

Radiation Safety SRS-76, Atomizdat, Moscow

(1977) (in Russian).

[4.17] BALONOV, M., “Overview of doses to the Soviet

population from the Chernobyl accident and

protective actions applied”, The Chernobyl Papers

(MERWIN, S., BALONOV, M., Eds), Research

Enterprises, Richland, WA (1993) 23–45.

[4.18] BALONOV, M.I., GOLIKOV, V.Y., ERKIN,

V.G., PARCHOMENKO, V.I., PONOMAREV,

A., “Theory and practice of a large-scale

programme for the decontamination of the settle-

ments affected by the Chernobyl accident”, Proc.

Int. Sem. on Intervention Levels and Counter-

measures for Nuclear Accidents, Rep. EUR 14469,

Office for Official Publications of the European

Communities, Luxembourg (1992) 397–415.

[4.19] HUBERT, P., ANISIMOVA, L., ANTSIPOV, G.,

RAMZAEV, V., SOBOTOVICH, V. (Eds), Strat-

egies for Decontamination, Final Report on

Experimental Collaboration Project No. 4, Rep.

EUR 16530 EN, Office for Official Publications of

the European Communities, Luxembourg (1996).

[4.20] ROED, J., et al., Mechanical Decontamination

Tests in Areas Affected by the Chernobyl

Accident, Rep. Riso-R-1029(EN), Risø National

Lab., Roskilde (1998).

[4.21]

VOVK, I., BLAGOEV, V., LYASHENKO, A.,

KOVALEV, I., Technical approaches to decon-

tamination of terrestrial environments in the CIS

(former USSR), Sci. Total Environ. 137 (1993) 49–

63.

[4.22] ANTSIPOV, G., TABACHNY, L., BALONOV,

M., ROED, J., “Evaluation of the effectiveness of

decontamination activities in the CIS countries for

objects contaminated as a result of the Chernobyl

accident”, Proc. Workshop on Restoration of

Contaminated Territories Resulting from the

Chernobyl Accident, Rep. EUR 18193 EN, Office

for Official Publications of the European Commu-

nities, Luxembourg (2000) 10–15.

[4.23] LOS, I., LIKHTAREV, I., The peculiarities of

urban environmental contamination and assess-

ment of actions aimed at reduction of public

exposure, Int. J. Radiat. Hyg. 1 (1993) 51–59.

[4.24] INTERNATIONAL COMMISSION ON RADI-

OLOGICAL PROTECTION, Optimization and

Decision-making in Radiological Protection,

Publication 55, Pergamon Press, Oxford and New

York (1989).

[4.25] ROED, J., ANDERSSON, K., PRIP, H., Practical

Means for Decontamination 9 Years After a

Nuclear Accident, Risoe-R-828(EN), Risø

National Lab., Roskilde (1995).

[4.26] NADTOCHIY, P.P., et al., Experience of Liquida-

tion of the Chernobyl Accident Consequences,

Svit, Kiev (2003) (in Ukrainian).

[4.27] ALEXAKHIN, R.A., KORNEEV, N.A. (Eds),

Agricultural Radioecology, Ecology, Moscow

(1991) (in Russian).

[4.28] PRISTER, B., PEREPELYATNIKOV, G.P.,

PEREPELYATNIKOVA, L.V., Countermeasures

used in the Ukraine to produce forage and animal

food products with radionuclide levels below

intervention limits after the Chernobyl accident,

Sci. Total Environ. 137 (1993) 183–198.

[4.29] USSR STATE AGROINDUSTRIAL

COMMITTEE, Guidance for Agricultural

Workers and the Population Inhabiting the Area

of the Radioactive Trail of the Chernobyl NPP,

USSR State Agroindustrial Committee, Moscow

(1986) (in Russian).

[4.30] ALEXAKHIN, R.M., Countermeasures in agri-

cultural production as an effective means of miti-

gating the radiological consequences of the

Chernobyl accident, Sci. Total Environ. 137 (1993)

9–20.

[4.31] BARYAKHTAR, V.G., Chernobyl Catastrophe,

Export Publishing House, Kiev (1997).

[4.32] MÜCK, K., “Environmental restoration by

natural effects — Advantages and limits”,

Radiation Legacy of the 20th Century: Environ-

mental Restoration, IAEA-TECDOC-1280,

IAEA, Vienna (2002) 115–126.

[4.33] PRISTER, B., ALEKSAKHIN, R., FIRSA-

KOVA, S., HOWARD, B., “Short and long term

environmental assessment”, Proc. Workshop on

Restoration of Contaminated Territories

Resulting from the Chernobyl Accident, Rep.

EUR 18193 EN, Office for Official Publications of

the European Communities, Luxembourg (2000)

103–114.

[4.34] SHEVCHUK, V.E., GOURACHEVSKIY, V.L.

(Eds), 15 Years After the Chernobyl Catastrophe:

Consequences in the Republic of Belarus and

their Overcoming, National Report, Committee

on the Problems of the Consequences of the

Accident at the Chernobyl NPP, Minsk (2001).

[4.35] ALEXAKHIN, R.M. (Ed.), Recommendations

for 1991–1995 on Agriculture Management in

Areas Subjected to Contamination as a Result of

the Accident at the Chernobyl NPP, State

Commission of the USSR Council of Ministers on

Food and Procurement, Moscow (1991) (in

Russian).

[4.36] PRISTER, B. (Ed.), Recommendations on Agri-

culture Management on Contaminated Territories,

Ukrainian Institute of Agricultural Radiology,

Kiev (1998) (in Ukrainian).

[4.37] BOGDEVITCH, I.M., Guidelines on Agricultural

and Industrial Production Under Radioactive

Contamination in the Republic of Belarus, Minsk

(2003) (in Russian).

[4.38] VIDAL, M., et al., Soil- and plant-based counter-

measures to reduce

137

Cs and

Sr uptake by

grasses in natural meadows: The REDUP project,

J. Environ. Radioact. 56 (2001) 139–156.

[4.39] BOGDEVITCH, I.M., PUTYATIN, Y.V.,

SHMIGELSKAYA, I.D., SERAYA, T.M.,

CENCEVICKI, F.A., Food Crop Production on

Personal Plots of Land Contaminated with Radio-

nuclides, Institute for Soil Science and Agrochem-

istry of the National Academy of Sciences of

Belarus, Minsk (2003) (in Russian).

[4.40] BOGDEVITCH, I., SANZHAROVA, N.,

PRISTER, B., TARASIUK, S., “Countermeasures

on natural and agricultural areas after Chernobyl

accident”, Role of GIS in Lifting the Cloud off

Chernobyl (KOLEJKA, J., Ed.), Kluwer,

Dordrecht (2002) 147–158.

[4.41] INTERNATIONAL ATOMIC ENERGY

AGENCY, The Use of Prussian Blue to Reduce

Radicaesium Contamination of Milk and Meat

Produced on Territories Affected by the

Chernobyl Accident, IAEA-TECDOC-926,

IAEA, Vienna (1997).

[4.42] RATNIKOV, A.N., et al., The use of hexacyanof-

errates in different forms to reduce radiocaesium

contamination of animal products in Russia, Sci.

Total Environ. 223 (1998) 167–176.

[4.43] HOVE, K., HANSEN, H.S., Reduction of radio-

caesium transfer to animal products using

sustained release boli with ammoniumiron(III)–

hexacyanoferrate(II), Acta Vet. Scand. 34 (1993)

287–297.

[4.44] HOVE, K., et al., “Use of caesium binders to

reduce radiocaesium contamination of milk and

meat in Belarus, Russia and Ukraine”, Environ-

mental Impact of Radioactive Releases (Proc. Int.

Conf. Vienna, 1995), IAEA, Vienna (1995) 539–

547.

[4.45] DEVILLE-CAVELIN, G., et al., “Countermeas-

ures in agriculture: Assessment of efficiency”,

Fifteen Years after the Chernobyl Accident:

Lessons Learned (Proc. Int. Conf. Kiev), Cherno-

bylinterinform, Kiev (2001) 118–128.

[4.46] JACOB, P., et al., Remediation strategies for rural

territories contaminated by the Chernobyl

accident, J. Environ. Radioact. 56 (2001) 51–76.

[4.47] HANSEN, H.S., HOVE, K., BARVIK, K., The

effect of sustained release boli with ammonium-

iron (III)–hexacyanoferrate(II) on radiocaesium

accumulation in sheep grazing contaminated

pasture, Health Phys. 71 (1996) 705–712.

[4.48] BRYNILDSEN, L.I., SELNAES, T.D., STRAND,

P., HOVE, K., Countermeasures for radiocaesium

in animal products in Norway after the Chernobyl

accident — Techniques, effectiveness, and costs,

Health Phys. 70 (1996) 665–672.

[4.49] HOVE, K., Chemical methods for reduction of the

transfer of radionuclides to farm animals in semi-

natural environments, Sci. Total Environ. 137

(1993) 235–248.

[4.50] ÅHMAN, B., personal communication: Data from

the Swedish Board of Agriculture, Jönköping, 2005.

[4.51] OUGHTON, D., FORSBERG, E.-M., BAY, I.,

KAISER, M., HOWARD, B., An ethical

dimension to sustainable restoration and long-

term management of contaminated areas, J.

Environ. Radioact. 74 (2004) 171–183.

[4.52] Sustainable Restoration and Long-term Manage-

ment of Contaminated Rural, Urban and Indus-

trial Ecosystems, www.strategy-ec.org.uk/

[4.53] HERIARD DUBREUIL, G., et al., Chernobyl

post-accident management: The ETHOS project,

Health Phys. 77 (1999) 361–372.

[4.54] LOCHARD, J., “Living in contaminated territo-

ries: A lesson in stakeholder involvement”,

Current Trends in Radiation Protection, EDP

Sciences, Les Ulis (2004) 211–220.

[4.55] ENVREG 9602, http://mns.gov.ua/chornobyl/

envreg-9602icd/?m=17

[4.56] CORE Program, http://www.core-chernobyl.org/

eng/

[4.57] UNITED NATIONS DEVELOPMENT

PROGRAMME, UNITED NATIONS CHIL-

DREN’S FUND, The Human Consequences of

the Chernobyl Nuclear Accident: Strategy for

Recovery, UN, New York (2002).

[4.58] On the Legal Regime of Territories Contaminated

as a Result of the Chernobyl NPP Catastrophe,

Bulletin of the Supreme Soviet of the Republic of

Belarus No. 35 (1991) 622 (in Russian).

[4.59] BOGDEVITCH, I.M., SHMIGELSKAYA, I.D.,

TARASIUK, S.V., Rational using of contaminated

soils after the Chernobyl accident in Belarus, Nat.

Resour. (1998) 15–29 (in Russian).

[4.60] STATE COMMITTEE OF RUSSIA FOR

SANITARY INSPECTION, TPL-1993

Temporary Permissible Levels of Caesium-134

and -137 and Strontium-90 in Food Products,

Hygienic Standard GS 2.6.005-93, State

Committee of Russia for Sanitary Inspection,

Moscow (1993).

[4.61] GUILLITTE, O., WILLDRODT, C., An assess-

ment of experimental and potential countermeas-

ures to reduce radionuclide transfers in forest

ecosystems, Sci. Total Environ. 137 (1993) 273–

288.

[4.62] GUILLITTE, O., TIKHOMIROV, F.A., SHAW,

G., VETROV, V., Principles and practices of coun-

termeasures to be carried out following radio-

active contamination of forest areas, Sci. Total

Environ. 157 (1994) 399–406.

[4.63] AMIRO, B.D., GREBEN’KOV, A., VANDEN-

HOVE, H., “Countermeasures and risks associ-

ated with contaminated forests”, Contaminated

Forests: Recent Developments in Risk Identifica-

tion and Future Perspectives (LINKOV, I.,

SCHELL, W.R., Eds), NATO Science Series,

Vol. 58, Kluwer, Dordrecht (1999) 395–401.

[4.64] RAFFERTY, B., SYNNOTT, H., FORECO —

Countermeasures Applied in Forest Ecosystems

and their Secondary Effects, Serie Documenti 6/

1998, Agenzia Nazionale per la Protezzione

dell’Ambiente, Rome (1998).

[4.65] FESENKO, S., BROWN, J., Review of Counter-

measures Options for Semi-natural Environments:

Forest and Natural Meadows, Rep. NRPB-M1123,

National Radiological Protection Board, Didcot,

UK (2000).

[4.66] TIKHOMIROV, F.A., SHCHEGLOV, A.I., Main

investigation results on the forest radioecology in

the Kyshtym and Chernobyl accident zones, Sci.

Total Environ. 157 (1994) 45–57.

[4.67] PANFILOV, A., “Countermeasures for radioac-

tively contaminated forests in the Russian Federa-

tion”, Contaminated Forests: Recent

Developments in Risk Identification and Future

Perspectives (LINKOV, I., SCHELL, W.R., Eds),

NATO Science Series, Vol. 58, Kluwer, Dordrecht

(1999) 271–279.

[4.68] SHAW, G., ROBINSON, C., HOLM, E.,

FRISSEL, M.J., CRICK, M., A cost–benefit

analysis of long-term management options for

forests following contamination with

137

Cs, J.

Environ. Radioact. 56 (2001) 185–208.

[4.69] HUBBARD, L., RANTAVAARA, A.,

ANDERSSON, K., ROED, J., Tools for Forming

Strategies for Remediation of Forests and Park

Areas in Northern Europe after Radioactive

Contamination: Background and Techniques, Rep.

NKS-52, Nordisk Kernesikkerhedsforskning,

Roskilde (2002).

[4.70] MINISTRY OF HEALTH, Permissible Levels of

137

Cs and

Sr Contents in Products of Forestry:

Hygienic Standards, GH 2.6.1.670-97, Ministry of

Health, Moscow (1997) (in Russian).

[4.71] JOHANSSON, K., “Radiocaesium in game

animals in the Nordic countries”, Nordic Radio-

ecology: The Transfer of Radionuclides Through

Nordic Ecosystems to Man (DAHLGAARD, H.,

Ed.), Elsevier, Amsterdam (1994) 287–301.

[4.72] ÅHMAN, B., Radiocaesium in Reindeer

(Rangifer Tarandus Tarandus) after Fallout from

the Chernobyl Accident, Thesis, Swedish Univer-

sity of Agricultural Sciences, Uppsala (1994).

[4.73] BERKOVSKI, V., VOITSEKHOVITCH, O.A.,

NASVIT, O., ZHELEZNYAK, M., SANSONE,

U., “Exposures from aquatic pathways”, The

Radiological Consequences of the Chernobyl

Accident (Proc. Int. Conf. Minsk, 1996), Rep.

EUR 16544 EN, Office for Official Publications of

the European Communities, Luxembourg (1996)

283–294.

[4.74] VOITSEKHOVITCH, O.V., SANSONE, U.,

ZHELESNYAK, M., BUGAI, D., “Water quality

management of contaminated areas and its effects

on doses from aquatic pathways”, The Radiolog-

ical Consequences of the Chernobyl Accident

(Proc. Int. Conf. Minsk, 1996), Rep. EUR 16544

EN, Office for Official Publications of the

European Communities, Luxembourg (1996) 401–

410.

[4.75] STONE, D., SMITH, J.T., JACKSON, D.,

IBBOTSON, A.T., Scoping Study on the Relative

Importance of Freshwater Dose Pathways

Following a Major Nuclear Accident, Westlakes

Research Institute, Whitehaven, UK (1997).

[4.76] VOITSEKHOVITCH, O., et al., “Hydrological

processes and their influence on radionuclide

behaviour and transport by surface water

pathways as applied to water protection after

Chernobyl accident”, Hydrological Impact of

Nuclear Power Plants (Proc. Int. Workshop, Paris,

1992), UNESCO, Paris (1993) 85–105.

[4.77] WATERS, R., et al., “A review of post-accident

measures affecting transport and isolation of radi-

onuclides released from the Chernobyl accident”,

Environmental Contamination in Central and

Eastern Europe (Proc. Int. Symp. Budapest, 1994),

Florida State University, Tallahassee (1996) 728–

730.

[4.78] SMITH, J.T., VOITSEKHOVITCH, O.V.,

HÅKANSON, L., HILTON, J., A critical review

of measures to reduce radioactive doses from

drinking water and consumption of freshwater

foodstuffs, J. Environ. Radioact. 56 (2001) 11–32.

[4.79] VOITSEKHOVITCH, O.V., Management of

Surface Water Quality in the Areas Affected by

the Chernobyl Accident, Ukrainian Hydromete-

orological Institute, Kiev (2001) (in Russian).

[4.80] VOITSEKHOVITCH, O.V., NASVIT, O., LOS’Y,

Y., BERKOVSKI, V., “Present thoughts on

aquatic countermeasures applied to regions of the

Dnieper River catchment contaminated by the

1986 Chernobyl accident”, Freshwater and

Estuarine Radioecology (DESMET, G., et al.,

Eds), Elsevier, Amsterdam (1997) 75–86.

[4.81] TSARIK, N., Supplying water and treating sewage

in Kiev after the Chernobyl accident, J. Am. Water

Works Assoc. 85 (1993) 42–45.

[4.82] ZHELEZNYAK, M., SHEPELEVA, V., SIZO-

NENKO, V., MEZHUEVA, I., Simulation of

countermeasures to diminish radionuclide fluxes

from the Chernobyl zone via aquatic pathways,

Radiat. Prot. Dosim. 73 (1997) 181–186.

[4.83] VAKULOVSKY, S.M., et al., Cs-137 and Sr-90

contamination of water bodies in the areas

affected by releases from the Chernobyl nuclear

power plant accident: An overview, J. Environ.

Radioact. 23 (1994) 103–122.

[4.84] RYABOV, I.N., “Analysis of countermeasures to

prevent intake of radionuclides via consumption

of fish from the region affected by the Chernobyl

accident”, Intervention Levels and Countermeas-

ures for Nuclear Accidents (Proc. Int. Sem. Cada-

rache, 1991), Rep. EUR-14469, Office for Official

Publications of the European Communities,

Luxembourg (1992) 379–390.

[4.85] BRITTAIN, J.E., STORRUSTE, A., LARSEN,

E., Radiocaesium in brown trout (Salmo trutta)

from a subalpine lake ecosystem after the

Chernobyl reactor accident, J. Environ. Radioact.

14 (1991) 181–191.

[4.86] CAMPLIN, W.C., LEONARD, D.R.P., TIPPLE,

J.R., DUCKETT, L., Radioactivity in Freshwater

Systems in Cumbria (UK) Following the

Chernobyl Accident, MAFF Fisheries Research

Data Report No. 18, Ministry of Agriculture,

Fisheries and Food, London (1989).

[4.87] HÅKANSON, L., ANDERSSON, T., Remedial

measures against radioactive caesium in Swedish

lake fish after Chernobyl, Aquat. Sci. 54 (1992)

141–164.

[4.88] KUDELSKY, A.V., SMITH, J.T., PETROVICH,

A.A., An experiment to test the addition of

potassium to a non-draining lake as a counter-

measure to

137

Cs accumulation in fish, Radiopro-

tection-Colloques 37 (2002) 621–626.

[4.89] SMITH, J.T., KUDELSKY, A.V., RYABOV, I.N.,

HADDERINGH, R.H., BULGAKOV, A.A.,

Application of potassium chloride to a Chernobyl-

contaminated lake: Modelling the dynamics of

radiocaesium in an aquatic ecosystem and decon-

tamination of fish, Sci. Total Environ. 305 (2003)

217–227.

[4.90] RANTAVAARA, A.H., “Transfer of radio-

nuclides during processing and preparation of

foods: Finnish studies since 1986”, Radioactivity

Transfer during Food Processing and Culinary

Preparation (Proc. Int. Sem. Cadarache, 1989),

Office for Official Publications of the European

Communities, Luxembourg (1989) 69–94.

[4.91] FOOD AND AGRICULTURE ORGANIZA-

TION OF THE UNITED NATIONS, INTER-

NATIONAL ATOMIC ENERGY AGENCY,

Guidelines for Agricultural Countermeasures

Following an Accidental Release of Radionu-

clides, Technical Reports Series No. 363, IAEA,

Vienna (1994).

[4.92] FOOD AND AGRICULTURE ORGANIZA-

TION OF THE UNITED NATIONS, INTER-

NATIONAL ATOMIC ENERGY AGENCY,

INTERNATIONAL LABOUR ORGANISA-

TION, OECD NUCLEAR ENERGY AGENCY,

PAN AMERICAN HEALTH ORGANIZA-

TION, WORLD HEALTH ORGANIZATION,

International Basic Safety Standards for Protec-

tion against Ionizing Radiation and for the Safety

of Radiation Sources, Safety Series No. 115,

IAEA, Vienna (1996).

100

5. HUMAN EXPOSURE LEVELS

5.1. INTRODUCTION

5.1.1. Populations and areas of concern

Following the Chernobyl accident, both

workers and the general public were affected by

radiation that carried a risk of adverse health

effects. UNSCEAR selected the following three

categories of exposed populations: (a) workers

involved in the accident, either during the

emergency period or during the cleanup phase; (b)

the inhabitants of contaminated areas who were

evacuated in 1986; and (c) the inhabitants of

contaminated areas who were not evacuated [5.1].

In this section consideration is given primarily

to members of the general public exposed to radio-

nuclides deposited in the environment. The workers

involved in the emergency response to the accident

or in the cleanup following the accident and

exposed predominantly on-site (i.e. at the

Chernobyl nuclear power plant and in the CEZ) are

not considered here. For information on Chernobyl

worker populations, the reader is referred to the

comprehensive material provided by UNSCEAR

[5.1, 5.2] and by the Chernobyl Forum in its report

considering the human health effects [5.3].

Information on the radiation doses received

by members of the general public, both those

evacuated from the accident area and those who

live permanently in contaminated areas, is required

for the following health related purposes:

(a) Substantiation of countermeasures and

remediation programmes;

(b) Forecast of expected adverse health effects

and justification of corresponding health

protection measures;

(d) Epidemiological and other medical studies of

radiation caused adverse health effects.

In this section the methodologies and data

specifically required for the estimation of mean

doses to population groups living in particular

settlements and selected by the factors influencing

either external or internal dose or both are

presented. These factors are usually age, sex,

occupation, food habits, etc. Dose distributions

among group members and collective doses are also

considered. Individual doses to members of the

public, used mainly in analytical epidemiological

studies, are presented in the Chernobyl Forum

report on the health consequences of the Chernobyl

accident [5.3]. On these subjects, substantial

progress has been achieved since publication of the

comprehensive UNSCEAR report in 2000 [5.1].

As mentioned in Section 3.1, atlases have

been prepared that show the deposition of

137

Cs and

other radionuclides throughout the former USSR

and other countries of Europe [5.4, 5.5]. These

indicate that the most affected countries are

Belarus, the Russian Federation and Ukraine. In

addition, the countries of Austria, Bulgaria, Finland,

Greece, Italy, Norway, Republic of Moldova,

Slovenia, Sweden and Switzerland had areas that

can be considered to have been ‘contaminated’ —

that is, at the level of more than 37 kBq/m

(>1 Ci/

) of

137

Cs (see Table 3.2).

5.1.2. Exposure pathways

Following the Chernobyl accident there were

several pathways by which humans were exposed to

radioactive material (Fig. 5.1). The main pathways

are listed below in the approximate time sequence

in which the doses were received:

(a) External dose from cloud passage;

(b) Internal dose from inhalation during cloud

passage and of resuspended material;

upon soil and other surfaces;

FIG. 5.1. Pathways of exposure of humans to environ-

mental releases of radioactive material.

101

(d) Internal dose from the consumption of

contaminated food and water.

Under most exposure conditions for members

of the general public the two most important

pathways are dose from radiation from the decay of

radionuclides deposited upon the soil and other

surfaces and dose from the ingestion of contami-

nated food and water. If persons are evacuated

quickly after passage of the initial cloud, then the

most important pathways are the first two in the list,

because the latter two pathways have been

prevented.

5.1.3. Concepts of dose

Methods of calculating radiation dose have

been refined over the years, and specific concepts

have evolved [5.1, 5.6]. The fundamental measure of

radiation dose to an organ or tissue is the absorbed

dose, which is the amount of energy absorbed by

that organ or tissue divided by its weight. The inter-

national unit of absorbed dose is the gray (Gy),

which is equal to one joule per kilogram. Since this

is a rather large amount of dose, it is common to use

units of mGy (one thousandth of a gray) or µGy

(one millionth of a gray).

Since many organs and tissues were exposed

as a result of the Chernobyl accident, it has been

very common to use an additional concept, effective

dose, which is the sum of the products of absorbed

dose to each organ multiplied by a radiation

weighting factor and a tissue weighting factor. The

former varies by radiation type and is related to the

density of ionizations created; the latter is an

approximation of the relative probability that an

absorbed dose to a particular organ might lead to

the production of a cancer. The sum of all tissue

weighting factors is equal to 1.0.

The concepts mentioned above are applied to

individuals. Where many individuals have been

exposed to an event, such as happened following the

Chernobyl accident, an additional concept, the

collective dose, can be used. The collective dose is

the sum of the doses to all individuals within a

particular group, which may be the residents of a

particular country or the persons involved in some

type of activity, such as cleaning up the conse-

quences of the accident. This concept is most often

applied to effective doses, and the common unit of

the collective effective dose is the man-Sv.

Finally, UNSCEAR has employed the concept

of dose commitment to examine the long term

consequences of a practice or accident [5.1]; for

example, at the very moment that the Chernobyl

accident occurred, it can be considered that a dose

commitment occurred at the moment of the release

of the radioactive material. This is true even though

it will take many years for the doses to be received

by the persons alive at that time and by persons not

yet born or conceived.

5.1.4. Background radiation levels

Living organisms are continually exposed to

ionizing radiation from natural sources, which

include cosmic rays and terrestrial radionuclides

(such as

238

232

Th and their progeny, including

222

Rn (radon)). Table 5.1 shows the average annual

dose and typical dose range worldwide from natural

sources.

In addition to natural sources, radiation

exposure occurs as a result of human activities.

Table 5.2 shows the annual individual effective

TABLE 5.1. RADIATION DOSES FROM NATURAL SOURCES [5.1]

Worldwide average annual effective dose

(mSv)

Typical range (mSv)

External exposure

Cosmic rays 0.4 0.3–1.0

Terrestrial gamma rays 0.5 0.3–0.6

Internal exposure

Inhalation (mainly radon) 1.2 0.2–10

Ingestion 0.3 0.2–0.8

Total 2.4 1–10

102

doses in 2000 on a worldwide basis. Diagnostic

medical exposure is the largest non-natural source

of radiation. The residual global effects of the

Chernobyl accident are now very small but, of

course, are higher in European countries and

especially in areas of Belarus, the Russian

Federation and Ukraine.

5.1.5. Decrease of dose rate with time

To calculate the radiation dose for particular

time periods, it is necessary to predict the decrease

of dose rate with time. The most obvious

mechanism acting to cause such a decrease is the

radioactive decay of the radionuclides. Additional

DRRFs are usually called ecological half-lives; for

example, external gamma exposure rates decrease

with time due to the weathering of long lived radio-

nuclides such as

137

Cs into the soil and subsequent

migration down the soil column, which results in

increased absorption of the emitted radiations

within the soil. Typically, a two component

exponential function describes this process [5.7,

5.8].

The availability of

137

Cs for ingestion also

decreases with time at a rate faster than radioactive

decay. This additional long term decrease is due

mainly to the adsorption of

137

Cs to soil particles

from which the caesium atoms are no longer biolog-

ically available. As with the external dose rate, the

decrease of

137

Cs in milk or in humans living in areas

contaminated by the Chernobyl accident also shows

a two component exponential decrease with time

[5.9, 5.10].

5.1.6. Critical groups

In all situations that involve the exposure of

large segments of the population to natural or

human-made radioactive material, there is always a

significant spread in the radiation dose received by

various members of the population living within the

same geographical area. Those individuals with the

higher doses are frequently called the critical group,

and these persons may have doses twice or even

higher than the average dose to all members of the

population considered. Usually such persons can be

identified in advance, and, in some cases, special

protective measures may be considered.

For external dose, members of the critical

group are those who spend a considerable amount

of time outdoors, either for occupational or recrea-

tional reasons; also, people living and/or working in

buildings with minimal shielding might be members

of the critical group. For exposure to radioiodine

isotopes, the critical group is often infants drinking

goat’s milk. Infants have a thyroid gland weighing

only two grams that concentrates roughly 30% of

the radioiodine consumed; goats are more efficient

than cows at secreting radioiodine into milk. For

exposure to radiocaesium, critical groups have been

identified as those who consume large quantities of

local animal products such as milk and meat and

wild products such as game meat, mushrooms, wild

berries and lake fish.

TABLE 5.2. EFFECTIVE DOSES IN 2000 FROM NATURAL AND HUMAN SOURCES [5.1]

Worldwide average annual

per caput effective dose (mSv)

Range or trend in exposure

Natural background 2.4 Typical range of 1–10 mSv

Diagnostic medical examinations 0.4 Ranges from 0.04 to 1 mSv at the lowest and

highest levels of health care

Atmospheric nuclear testing 0.005 Has decreased from a maximum of 0.15 mSv in

1963; higher in the northern hemisphere

Chernobyl accident 0.002 Has decreased from a maximum of 0.04 mSv in

1986 (in the northern hemisphere); higher at

locations nearer the accident site

Nuclear power production 0.0002 Has increased with expansion of the nuclear

programme but decreased with improved practice

103

5.2. EXTERNAL EXPOSURE

5.2.1. Formulation of the model for external

exposure

In any situation of human external exposure

caused by releases of radioactive substances into the

environment, the following three types of data are

necessary for assessment of organ or effective

doses:

(a) Parameters that describe the external gamma

radiation field;

(b) Parameters describing human behaviour in

this field;

or effective dose.

The basic model for human external exposure

in the event of radioactive contamination of the

environment is the model for exposure above an

open plot of undisturbed soil; the absorbed dose in

air D(t) at a height of 1 m above the soil surface is

used as the basic parameter to describe the radiation

field. The value of this basic parameter is influenced

not only by the surface activity of deposited radio-

nuclides but also by such natural factors as the initial

penetration of radionuclides in soil and their

radioactive decay, vertical migration of long lived

radionuclides and the presence of snow cover.

Radiation exposure is influenced by altered or

disturbed environments. In models this factor is

taken into account by using location factors. The

location factor LF

is defined as the ratio of the dose

rate in air at point i inside a settlement to a similar

value above a plot of undisturbed soil [5.11].

Human behaviour in the radiation field is described

by occupancy factors OF

, which represent the

fraction of time spent by individuals of the kth

population group at the ith point of the settlement

of interest. The third type of data necessary for

assessment of the effective external dose are

conversion factors CF

, which convert measured

values (the absorbed dose in air) to a parameter

that can be directly related to health effects — the

effective dose to the kth population group.

On this basis, a deterministic model for the

assessment of the effective external dose rate E

for

representatives of the kth population group is

represented in Fig. 5.2.

5.2.2. Input data for the estimation of effective

external dose

Numeric values for the parameters listed

above have been determined from long term

dosimetric investigations in the most highly contam-

inated regions after the Chernobyl accident.

5.2.2.1. Dynamics of external gamma dose rate over

open undisturbed soil

Immediately after the accident, external

gamma exposure rates were relatively high, and

contributions from many short lived radionuclides

were important. Thus in the contaminated areas

outside the Chernobyl nuclear power plant

boundaries the initial dose rate over lawns and

meadows ranged between 3 and 10 µGy/h in areas

contaminated at about 37 kBq/m

(1 Ci/km

) of

137

Cs and up to 10 000 µGy/h within the CEZ with

higher deposition levels. Exposure rates decreased

rapidly, due to the radioactive decay of short lived

radionuclides, as shown in Fig. 5.3.

Owing to different isotopic compositions of

radionuclide fallout in different geographical areas

[5.8, 5.13, 5.14], the contribution of short lived

radionuclides to the overall dose rate was highly

variable. In the CEZ,

132

Te +

132

131

I and

140

Ba +

140

La dominated during the first month and then

Zr +

Nb for another half year before

137

Cs and

134

Cs became dominant (Fig. 5.4). In contrast, in the

far zone only the radioiodine isotopes dominated

during the first month; afterwards

137

Cs and

134

dominated, with a moderate contribution from

103

Ru and

106

Ru (Fig. 5.5). Since 1987 more than

∑

∫

dtLFCFOFE

Location

factorover soil

()

Dose conversion

factor

Occupancy

factor

∑

∫

ikik

Dose rate

External dose to humans

()

FIG. 5.2. Model of external exposure of the kth population

roup (i is a location index) [5.9].

104

90% of the dose rate in air has come from the

gamma radiation of long lived

137

Cs and

134

Cs. Thus

the radionuclide composition of the deposited

activity was a major factor in determining the

external exposure of the population in the early

period of time after the accident. Model estimates

of the gamma dose rate in free air (90% confidence

interval) based on the radionuclide composition of

the deposited activity agree well with the measured

values during the first month after deposition (see

Fig. 5.6).

The influence of radionuclide migration into

soil on the gamma dose rate has been determined

using gamma spectrometric analyses of over 400 soil

samples taken during 1986–1999 in the contami-

nated areas of Germany (Bavaria), the Russian

Time after the accident (days)

0.1

1.0

10.0

100.0

1000.0

1 10 100 1000 10 000

North-west direction, CEZ

South direction, CEZ

Far zone (>100 km)

Absorbed dose rate in air (nGy/h per 1 kBq/m

137

Cs)

FIG. 5.3. Dynamics of standardized dose rate in air ove

undisturbed soil after the Chernobyl accident in differen

eographical areas [5.12].

Time after the accident (days)

North-west direction, CEZ

Contribution to absorbed dose rate in air 1 m above ground (%)

Cs-137; Cs-134; Cs-136

Zr-95 + Nb-95

Te-132 + I-132

I-131; I-133

Ba-140 + La-140

Ru-103; Ru-106

1 10 100

100

FIG. 5.4. Relative contribution of gamma radiation from

individual radionuclides to the external gamma dose rate in

air during the first year after the Chernobyl acciden

(north-west direction, CEZ) [5.12].

Time after the accident (days)

100

1 10 100

Cs-137; Cs-134; Cs-136

Te-132 + I-132

Ru-103; Ru-106

Ba-140 + La-140

I-131; I-133

Zr-95 + Nb-95

Far zone (>100 km)

Contribution to absorbed dose rate in air 1 m above ground (%)

FIG. 5.5. Relative contribution of gamma radiation from

individual radionuclides to the external gamma dose rate in

air during the first year after the Chernobyl accident (fa

zone — more than 100 km from the Chernobyl nuclea

ower plant) [5.12].

Time after the accident (days)

0 5 10 15 20 25 30 35

Mean

95%

Dose rate (relative units)

FIG. 5.6. Dose rate in air during the first days after the

accident in several rural settlements in the Bryansk and

Tula regions of the Russian Federation (normalized to the

dose rate on 10 May 1986). Points indicate dose rate meas-

urements and curves represent calculated values accordin

to the isotopic composition [5.7].

105

Federation, Sweden and Ukraine [5.7, 5.8, 5.15].

The analysis also included data on the

137

Cs distri-

bution in soil at sites in the north-east region of the

USA, whose contamination was attributed to

nuclear tests at the Nevada test site [5.16], and in

Bavaria (Germany), where contamination was due

to global fallout. The last two data sets were

obtained 20 to 30 years after deposition; this allows

for long term predictions to be applied to the

Chernobyl depositions. The measurement sites were

considered to be representative of reference sites

(i.e. open, undisturbed fields).

For a few years after the accident, the dose

rate over open plots of undisturbed soil decreased

by a factor of 100 or more compared with the initial

level (see Fig. 5.3). At that time, the dose rate was

mainly determined by gamma radiation of caesium

radionuclides (i.e.

137

Cs (half-life 30 years) and

134

(half-life 2.1 years), and later, one decade and more

after the accident, mainly the longer lived

137

Cs).

Long term studies of external gamma exposure

rates during the past 17 years have shown that the

external gamma exposure rate is decreasing faster

than that due to radioactive decay alone. Golikov et

al. [5.7] and Likhtarev et al. [5.8] have calculated a

reference function for

137

Cs gamma radiation dose

rate that has 40–50% of the exposure rate

decreasing with an ecological half-life of 1.5–2.5

years and the remaining 50–60% decreasing with an

ecological half-life of 40–50 years, as indicated in

Fig. 5.7. The latter value is rather uncertain. It

corresponds to an effective half-life of 17–19 years

that takes into account both the radioactive decay

137

Cs and its gradual deepening in soil.

5.2.2.2. Dynamics of external gamma dose rate in

anthropogenic areas

In settlements in urban and rural areas, the

characteristics of the radiation field differ consid-

erably from those over an open plot of undisturbed

land, which is used as the reference site and starting

point for calculation of external dose to people from

deposited activity. These differences are attrib-

utable to varying source distributions as a result of

deposition, runoff, weathering and shielding. All

such effects can be summarized by the term

‘location factors’.

Location factors for typical western European

buildings have been assessed [5.11, 5.17, 5.18].

Gamma spectrometric measurements performed in

Germany and Sweden [5.19–5.22] allowed the

determination of location factors in urban environ-

ments and their variation with time over several

years after the Chernobyl accident. The character-

istic feature, and advantage, of these investigations

is that they began immediately after the accident,

whereas systematic investigations of location factors

in the contaminated areas of Belarus, the Russian

Federation and Ukraine began two to three years

after the accident. The results of one such later

investigation in Novozybkov (in the Bryansk region

of the Russian Federation) are presented in Fig. 3.12

(Section 3).

5.2.2.3. Behaviour of people in the radiation field

The influence of the behaviour of different

social population groups on the level of exposure

can be taken into account if the frequency with

which people of the kth population group remain at

the location of the ith type is known. The times

spent in various types of location (indoor, outdoors

on streets or in yards, etc.) by members of different

population groups have been assessed on the basis

of responses to a questionnaire. Data collected

included age, sex, occupation, information about

dwelling, etc. An example of the results is shown in

Table 5.3, where values of occupancy factors for the

summer period are presented for different groups of

the rural populations of Belarus, the Russian

Federation and Ukraine [5.15].

Time after the accident (years)

0.0

0.2

0.4

0.6

0.8

1.0

0 5 10 15 20 25 30 35

95%

Median

‘Chernobyl’ caesium

Bryansk region (Russian Federation)

Caesium from

Nevada test site

(north-west USA)

Global fallout

from Bavaria

(Germany)

r(t) = 0.38 exp(–0.693 t/2.4y) + 0.39 exp(-0.693 t/37y)

Dose rate (relative units)

FIG. 5.7. Reduction of the

137

Cs gamma dose rate in air due

to caesium migration in undisturbed soil relative to the

dose rate caused by a plane source on the air–soil interface

(from Ref. [5.7]).

106

5.2.2.4. Effective dose per unit gamma dose in air

Mean values of conversion factors CF

, which

convert the gamma dose rate in air to the effective

dose rate in a member of population (age) group k,

were obtained for the three groups of population by

use of phantom experiments [5.15] and Monte

Carlo calculations [5.23]. The values were 0.75 Sv/

Gy for adults, 0.80 Sv/Gy for schoolchildren (7–

17 years) and 0.90 Sv/Gy for pre-school children (0–

7 years). For the calculation of effective doses,

conversion factors CF

were used that are

independent of the location and time after the

accident.

5.2.3. Results

5.2.3.1. Dynamics of external effective dose

Shortly after the deposition of the fallout the

gamma radiation field was dominated by emissions

from short lived radionuclides, as discussed above

(see Figs 5.4 and 5.5). As the mixtures at different

locations varied widely, the radionuclide

composition of the deposited activity was a major

factor in determining the external exposure of the

population during the early period after the

accident.

Another relevant parameter in the midterm

period is the dependence of location factors on

time, due to the relatively fast migration processes

of radionuclides during this period. The dose rate

over different urban surfaces caused by gamma

radiation of

137

Cs decreased during the first years

after deposition, with an exponential half-life of one

to two years (see Fig. 3.12). In the five to seven

years after deposition, the change in dose rate with

time had stabilized — this was due to the decay of

the short lived radionuclides and the fixation of

caesium radionuclides within the soil column.

According to measurements and evaluations

within the first year after the accident, the external

dose rate had decreased by a factor of approxi-

mately 30, mainly due to radioactive decay of short

lived radionuclides (see Fig. 5.8). During the

following decade the external dose rate decreased

because of the radioactive decay of

134

Cs and

137

and the migration of radiocaesium into the soil.

Afterwards, the external dose rate was mainly due

137

Cs. In the long term, radiocaesium becomes

fixed within the soil matrix, and this results in a slow

migration into the soil and, correspondingly, in a

slow decrease of the external dose rate. On the basis

of such measurements, it is predicted that, of the

total external dose to be accumulated during 70

years following the accident, about 30% was

accumulated during the first year and about 70%

during the first 15 years (Fig. 5.8) [5.7].

5.2.3.2. Measurement of individual external dose

with thermoluminescent dosimeters

In general, before the Chernobyl accident,

individual external doses were measured only for

occupational exposures. After the Chernobyl

accident, individual external doses to members of

the population were also measured. For this

purpose thermoluminescent dosimeters were

distributed to the inhabitants of the more contami-

nated areas of Belarus, the Russian Federation and

Ukraine [5.24–5.28]. Inhabitants wore thermolumi-

nescent dosimeters for about one month in the

spring and summer periods. Examples of such

results are presented in Figs 5.9 and 5.10 for rural

and urban areas, respectively. According to these

results it can be concluded that the urban

TABLE 5.3. VALUES OF OCCUPANCY FACTORS FOR THE SUMMER PERIOD FOR DIFFERENT

GROUPS OF THE RURAL POPULATIONS OF THE RUSSIAN FEDERATION, BELARUS AND

UKRAINE

[5.15]

Location Indoor workers Outdoor workers Pensioners Schoolchildren

Pre-school

children

Inside houses 0.65/0.77/0.56 0.50/0.40/0.46 0.56/0.44/0.54 0.57/0.44/0.75 0.64/—/0.81

Outside houses

(living area)

0.32/0.19/0.40 0.27/0.25/0.29 0.40/0.42/0.41 0.39/0.45/0.21 0.36/—/0.19

Outside settlements 0.03/0.04/0.04 0.23/0.35/0.25 0.04/0.14/0.05 0.04/0.11/0.04 0/—/0

The first number corresponds to data for the Russian Federation, the second is for Belarus and the third is for Ukraine

[5.15].

107

population has been exposed to a lower dose by a

factor of 1.5–2 compared with the rural population

living in areas with similar levels of radioactive

contamination. This arises because of the better

shielding features of urban buildings and different

occupational habits.

The critical group in relation to external

irradiation is composed of individuals in an

occupation or with habits that result in spending a

significant amount of time outside in areas of

undisturbed soil, in forests or meadows, and who

also live in houses with the least protective

properties. At present, the average external dose

to any population group does not exceed the

average dose in a settlement by more than a factor

of two. Typical critical groups are foresters (factor

1.7), herders (factor 1.6) and field crop workers

(factor 1.3) living in one-storey wooden houses

[5.9, 5.15]

Analysis of the results of measurements of

inhabitants of settlements showed that the distri-

bution of individual doses can be described by a log-

normal function [5.7]. Figure 5.11 presents a

comparison of model calculations with individual

thermoluminescence measurements performed in

1993 in four villages of the Bryansk region (565

measurements). The distributions of the ratio of

individual external doses to the mean value of

measured doses in each of the villages are almost

identical. Thus the resulting log-normal distribution

with a geometric standard deviation of about 1.5

(attributed mainly to the stochastic variability of

individual doses) may be assumed to be typical for

rural settlements in the zone of the Chernobyl

accident.

Time after the accident (years)

0.01

0.10

1.00

10.00

100.00

0.01 0.10 1.00 10.00 100.00

Effective dose rate

Accumulated effective dose

Effective dose rate (10

–6

Sv/h) and accumulated

effective dose (10

–3

Sv) per 1 MBq/m

137

FIG. 5.8. Model prediction of the time dependence of the

external effective gamma dose rate and the accumulated

external effective dose to the urban population of the

Bryansk region of the Russian Federation [5.7].

Time after the accident (years)

5 10 15 20

100

150

200

250

Veprin

Smyalch

Mean effective monthly dose to village residents (mSv)

FIG. 5.9. Results of thermoluminescence measurements of

mean monthly doses among inhabitants living in wooden

houses in the Veprin and Smyalch villages (Bryansk region

of the Russian Federation) in different time periods

ollowing deposition [5.28].

Year

100

150

200

1989 1990 1991 1992 1993

Inhabitants living in wooden houses

Inhabitants living in brick houses

Mean effective monthly dose (10

–6

Sv)

FIG. 5.10. Effective dose rates for indoor workers in

Novozybkov (Russian Federation). The points with erro

bars represent average values and 95% confidence

intervals (+/– two standard errors) of thermoluminescence

measurements [5.28].

108

5.2.3.3. Levels of external exposure

To illustrate actual levels of external exposure

and differences in the level of exposure among

various population groups, Table 5.4 presents

calculated values of effective external doses during

different time intervals for rural and urban

populations in the Russian Federation and Ukraine,

and Table 5.5 presents the ratio of average effective

doses in separate population groups to the mean

dose in a settlement. Calculations of dose for

different time intervals were performed on the basis

of the model described above for the assessment of

external dose in a population.

At present, the average annual external dose

to residents of a rural settlement with a current

137

soil deposition of ~700 kBq/m

(~20 Ci/km

) is

0.9 mSv. For the critical group, the dose value

exceeds the annual dose limit of 1 mSv set for the

population under normal conditions. The external

dose due to Chernobyl deposition accumulated to

the present time is 70–75% of the total lifetime dose

(70 years) for persons born in 1986 and living all the

time in contaminated areas.

Individual dose/mean dose in village

0.0 0.3 0.6 0.9 1.2 1.5 1.8 2.1 2.4 2.7 3.0

Veprin

Kozhany

Smyalch

Vnukovichi

Relative frequency (%)

FIG. 5.11. Frequency distributions of monthly effective

external doses to individual persons as measured in the

ummer of 1993 with thermoluminescent dosimeters in

our villages of the Bryansk region of the Russian Federa-

tion (points) and calculated by the stochastic model

(curve). Doses are normalized to the arithmetic mean of

the individual doses determined for each of the villages

(from Ref. [5.7]).

TABLE 5.4. AVERAGE NORMALIZED EFFECTIVE EXTERNAL DOSE TO THE ADULT

POPULATION IN THE INTERMEDIATE (100 km < DISTANCE < 1000 km) ZONE OF CHERNOBYL

CONTAMINATION

Population

E/s

137

(mSv ◊ kBq

–1

◊ m

–2

137

Cs)

1986 1987–1995 1996–2005 2006–2056 1986–2056

Russian Federation Rural 1425101968

[5.7, 5.28] Urban 9 14 5 9 37

Ukraine [5.8] Rural 24 36 13 14 88

Urban 1725 91061

137

is given as for 1986.

TABLE 5.5. RATIO OF AVERAGE EFFECTIVE EXTERNAL DOSES IN SEPARATE POPULATION

GROUPS TO THE MEAN DOSE IN A SETTLEMENT [5.9]

Type of dwelling Indoor workers Outdoor workers Herders, foresters Schoolchildren

Wooden 0.8 1.2 1.7 0.8

One to two storey, brick 0.7 1.0 1.5 0.9

Multistorey 0.6 0.8 1.3 0.7

109

5.3. INTERNAL DOSE

5.3.1. Model for internal dose

The general form of models used to calculate

internal dose is shown in Fig. 5.12 [5.9]. The main

pathways of radionuclide intake into the body of a

person of age k are inhalation, with average

inhalation rate IR

of air with time dependent

concentration AC

of radionuclide r, and ingestion

of the set of f food products and water with

consumption rates CR

with time dependent

specific activity SA

Data on air concentrations and food activity

concentrations have been discussed in previous

sections and will be summarized briefly below. Data

on food consumption rates are taken from the

literature [5.2, 5.10] or from special surveys of the

affected populations [5.29, 5.30]. Other data needed

for dosimetric calculations are taken from publica-

tions of the International Commission on Radio-

logical Protection for age specific inhalation rates

[5.31] and for age specific dose coefficients [5.32].

The latter values are for both inhalation and

ingestion and give the dose per unit radionuclide

inhaled or ingested. These values are calculated in

terms of committed dose; that is, the dose that will

be received over the next 50 years for adults or until

age 70 for younger persons. For most radionuclides,

but not for

Sr or

239

Pu, the biological residence

time within the body is short, and the committed

dose is only slightly larger than the dose accrued

over the course of one year. Strontium and

plutonium nuclides, and a few others, are

metabolized slowly, and the full committed dose is

not actually received for many years.

Another method of calculating internal dose is

to use direct measurements of the radionuclide of

interest in the human body. This was done for

131

I in

the human thyroid in the three most affected

countries [5.33–5.35] and for

137

Cs (e.g. Refs [5.10,

5.36]). Especially for the thyroid, direct measure-

ments are not sufficient to calculate doses, and such

information must be supplemented by suitable

intake models to determine the past and future

concentration of the radionuclide in the body and

its organs.

Predictions of future intakes of long lived

radionuclides into the body must be made in order

to predict future doses. Information on the long

term transfer of the important radionuclide

137

from the environment to the human body can be

made on the basis of experience with this radionu-

clide in global and local fallout [5.1]. Also, enough

time has now passed since the Chernobyl accident

that measurements specific to Chernobyl can be

used to predict the future course of concentration of

137

Cs in foods and the human body; for example,

Likhtarev et al. [5.10] on the basis of 126 000

samples of milk collected during 1987–1997

observed a two component exponential loss curve

with 90% of

137

Cs activity disappearing with a half-

life of 2.9 ± 0.3 years and 10% with 15 ± 7.6 years.

The second value is very uncertain, due to the short

time of observation compared with the radiological

half-life of

137

Cs of 30 years. These data are in

general agreement with those observed in the

Russian Federation [5.37, 5.38].

5.3.2. Monitoring data as input for the assessment

of internal dose

A unique feature of Chernobyl related

monitoring of human internal exposure was the

extensive application of whole body measurements

of radionuclide content in the human body and its

organs (mainly thyroid); these measurements were

performed along with regular measurements of

radionuclides in food, drinking water and other

components of the environment. This combination

of various kinds of monitoring data allowed

substantial improvement in the precision of the

reconstruction of internal dose.

To assess internal dose from inhalation, the air

concentration measurements described in a

previous section have been used. The most

important aspect was assessment of dose for the

()

∑

∫

rrkkk

dttACDCIRE

Inhalation

()

tAC

()

tSA

()

∑

∫

frfkrkk

dttSACRCDE

Ingestion

∑

Air

concentration

Inhalation

rate

Inhalation dose

coefficient

Food specific

activity

Food

consumption rate

Ingestion dose

coefficient

∑

FIG. 5.12. Model for calculation of internal exposure fo

ersons exposed to Chernobyl fallout [5.9].

110

first days after the accident, when the concentration

of radionuclides in air was relatively high. Later, the

assessment of doses via inhalation was needed in

relation to the resuspension of radionuclides with

low mobility in the food chain, such as plutonium.

Assessment of radionuclide intake with food

and drinking water was primarily based on the

numerous measurements of

131

134,137

Cs and

Sr,

which have been performed all over Europe and

especially in the three most affected countries

(Belarus, the Russian Federation and Ukraine).

Gamma spectroscopy for

131

I and

134,137

Cs and radio-

chemical analyses for

Sr have been the main types

of measurement. In some laboratories beta

spectroscopy was successfully applied to determine

different radionuclides in samples; when radionu-

clide composition was well known, total beta

activity measurements were also made. In most of

the measurements,

137

Cs in raw animal products

(milk, meat, etc.) was determined; the number of

these measurements performed since 1986 and

available for dose estimation comprises a few

million. Generic data on radionuclide measure-

ments in food are presented in the sections

pertaining to the terrestrial environment.

Activity concentrations of soluble radio-

nuclides (mainly

131

134,137

Cs,

Sr) in drinking water

were determined in 1986 both in surface and

underground sources (see Section 3.5). Later, these

activity concentrations declined to relatively low

levels, and their contribution to internal dose was

usually negligible compared with that associated

with the intake of food.

In May–June 1986,

131

I activities were

measured in the thyroids of residents of areas with

substantial radionuclide deposition. In total, more

than 300 000

131

I measurements in the thyroid were

performed in the three most affected countries, and

a substantial number of measurements were also

performed in other European countries. Special

attention was paid to measurements of children and

adolescents. After careful calibration, data on large

scale measurements were used as the main basis for

the reconstruction of thyroid dose.

Most of the numerous whole body measure-

ments performed since 1986 in different European

countries have been aimed at the determination of

134,137

Cs. The number of measurements exceeded

one million, most of which were performed in the

three most affected countries. The measurement

data were widely used both for model validation

concerning radionuclide intake and evaluation of

the effectiveness of countermeasures. In the most

contaminated regions of Belarus, the Russian

Federation and Ukraine, whole body measurement

data were used to obtain more precise estimates of

human doses both for radiation protection purposes

and as part of epidemiological studies.

Strontium-90 and plutonium radionuclides,

which do not emit gamma radiation that is readily

detectable by whole body counters, have been

measured in excreta samples, and, since the 1990s, in

samples taken at autopsy. Several hundred samples

of human bone tissue have been analysed by

radiochemical methods for

Sr/

Y content.

Activities of plutonium radionuclides have been

successfully determined in several tens of samples

of human lungs, liver and bones [5.39, 5.40].

Reduced monitoring programmes for

radiation protection purposes, and specifically for

the justification of remediation efforts, are

continuing in the affected areas.

5.3.3. Avoidance of dose by human behaviour

In addition to the countermeasures employed

to reduce levels of contamination in urban environ-

ments and in agricultural foodstuffs, changes in

human habits after the accident were also effective

in reducing doses to residents of the contaminated

areas. The most obvious and highly effective

method immediately after the accident would have

been to stop the consumption of milk to reduce the

intake of

131

I. The effectiveness of this is not well

documented, and it is only in some of the more

affected regions that the residents of the three

countries were advised of this option in a timely

manner.

The longer term option of reducing the

consumption of food products known to be more

highly contaminated by

134,137

Cs appears to have

been more successful, at least during 1987–1993

[5.10, 5.41]. Such foods were typically locally

produced milk and beef or of the ‘wild’ variety,

including game meat, mushrooms and berries.

Later, due to deteriorating economic conditions and

the gradual reduction of the public’s caution over

wild food products, such self-imposed restrictions

became less widespread.

5.3.4. Results for doses to individuals

5.3.4.1. Thyroid doses due to radioiodines

One of the major impacts of the accident was

exposure of the human thyroid. Doses were

111

accumulated rather quickly due to the rapid transfer

of iodine through the food chain and the short half-

life of

131

I of eight days; other radioiodines of

interest in terms of thyroid dose also have short

half-lives. The importance of thyroid doses was

recognized by national authorities throughout the

world, and early efforts focused on this issue.

Estimates of country average individual thyroid

doses to infants and adults have been provided by

UNSCEAR [5.2]. Attention has been paid to

thyroid dose reconstruction since the early 1990s,

when an increase of thyroid cancer morbidity was

discovered in children and adolescents residing in

areas of Belarus, the Russian Federation and

Ukraine contaminated with Chernobyl fallout [5.1,

5.3, 5.42].

In association with radioepidemiological

studies, the main patterns of thyroid dose formation

were clarified and published in the 1990s [5.33–5.35]

and summarized in Ref. [5.1]. Nevertheless,

important new work in this area has appeared

recently [5.43–5.45]. The general approach to

internal dose reconstruction has been elaborated in

Ref. [5.46].

Methodologies of thyroid dose reconstruction

for the Chernobyl affected populations developed

in parallel in Belarus, the Russian Federation and

Ukraine, with the participation of US and EU

experts; these methodologies have a number of

commonalities and some substantial distinctions

that complicate their desired integration. Firstly, in

all three countries there are many tens of thousands

131

I thyroid measurements available, although of

different quality, that are used as the basic data for

thyroid dose reconstruction. In the Russian

Federation, additionally, data on

131

I in milk were

used. Owing to the use of human and environmental

131

I measurements, the reconstructed doses are

realistic rather than conservative.

Another commonality is the use of several age

groups living in one settlement or in a group of close

settlements as a unit for mean thyroid dose recon-

struction. When there is a substantial number of

human and environmental

131

I measurements

available in a settlement, they are used for dose

reconstruction. The subsidiary quantities used for

dose reconstruction in the settlements where

historical

131

I measurements are not available are

137

Cs soil deposition values as indicators of

radioactive contamination in the area.

However, the methodologies for thyroid dose

reconstruction for settlements without

environmental or human

131

I measurements are

substantially different in the three countries. In

Ukraine, where most of the radioiodine was

deposited in dry weather conditions, Likhtarev et al.

[5.45] developed a model with linear dependence of

thyroid dose on

137

Cs soil deposition. In Belarus,

where both dry and wet deposition of radioiodines

occurred, a semiempirical model based on non-

linear dependence of thyroid dose on

137

Cs soil

deposition was developed by Gavrilin et al. [5.35]

and widely applied. In another recently published

paper devoted to the same problem, a

comprehensive radioecological model of

radioiodine environmental transfer was developed

and successfully applied for thyroid dose

reconstruction [5.44]. In the Russian Federation,

where wet deposition of radioiodines dominated, a

linear semiempirical model of dependence of

131

activity concentration in milk and of thyroid dose

137

Cs soil deposition of more than 37 kBq/m

was

developed [5.43] and applied [5.47]. Despite

differences in the applied methodological

approaches, the general agreement, except for areas

of low contamination, is satisfactory [5.3].

The thyroid doses resulting from the

Chernobyl accident comprise four contributions: (a)

internal dose from intakes of

131

I; (b) internal dose

from intakes of short lived radioiodines (

132

133

and

135

I) and of short lived radiotelluriums (

131

and

132

Te); (c) external dose from the deposition of

radionuclides on the ground; and (d) internal dose

from intakes of long lived radionuclides such as

134

Cs and

137

Cs.

For most residents of the Chernobyl affected

areas, the internal thyroid dose resulting from

intakes of

131

I is by far the most important and has

received almost all of the attention. The dose from

131

I was mainly due to the consumption of fresh

cow’s milk and, to a lesser extent, of green

vegetables; children on average received a dose that

was much higher than that received by adults,

because of their small thyroid mass and a

consumption rate of fresh cow’s milk that was

similar to that of adults.

An example of age and sex dependence of the

mean thyroid dose to inhabitants of a settlement,

based on 60 000 measurements of

131

I in thyroids

performed in Ukraine in May 1986, is presented in

Fig. 5.13 [5.48]. The mean thyroid dose to infants is

larger by a factor of about seven than to young

adults (19–30 years) residing in the same rural or

urban settlements; this ratio decreases monotoni-

cally with age as an exponential function, with some

deviation in adolescents. Differences in age

112

dependence between males and females seem to be

insignificant. Similar patterns were revealed both

from Belarusian and Russian measurements of

131

in the thyroid [5.34, 5.35].

As the rural population living in contaminated

areas depends more on local agricultural production

than does an urban population, thyroid doses

caused predominantly by the consumption of

contaminated milk and dairy products are higher in

rural than in urban populations by a factor of about

two [5.1].

Although the largest contribution to thyroid

dose resulted from intakes of

131

I, it is also

important to take into consideration the internal

dose from short lived radioiodines (

132

133

I and

135

I). Among members of the public, the highest

relative contribution to the thyroid doses from short

lived radionuclides was expected among the

residents of Pripyat. This cohort was exposed to

radioiodines via inhalation only and was evacuated

about 1.5 days after the accident. Analysis of direct

thyroid and lung spectrometric measurements

performed on 65 Pripyat evacuees has shown that

the contribution of short lived radionuclides to

thyroid dose is about 20% for persons who did not

employ stable iodine to block their thyroids and

more than 50% for persons who took KI pills soon

after the accident [5.49]. The total thyroid dose

among the Pripyat evacuees, however, was

relatively small compared with populations

consuming contaminated food.

For populations permanently residing in

contaminated areas, the contribution of short lived

radionuclides to thyroid dose was minor, as most of

the thyroid exposure resulted from the week long

consumption of contaminated milk and other

foodstuffs. During transport of radioiodines along

food chains, short lived radioiodines decayed, and

the contribution of short lived radioiodines is

estimated to have been of the order of 1% of the

131

thyroid dose [5.49, 5.50].

The distribution of individual thyroid doses is

illustrated in Table 5.6 for children and adolescents

residing in the northern regions of Ukraine (i.e. the

Kiev, Zhytomyr and Chernigov regions) most

affected by radiation after the Chernobyl accident

[5.45]. The dose distributions presented in Table 5.6

are based on about 100 000 human thyroid measure-

ments. The range of thyroid dose in all groups is

wide, between less than 0.2 Gy and more than

10 Gy. The latter dose group includes about 1% of

younger children, less than 0.1% of children of five

to nine years, and less than 0.01% of adolescents.

Doses to adults are lower by a factor of about 1.5

than those to adolescents (see Fig. 5.13). In all age

groups presented in Table 5.6, and especially in the

younger ones, doses were high enough to cause both

short term functional thyroid changes and thyroid

cancer in some individuals [5.1, 5.3, 5.42].

Similar data for Belarus and the Russian

Federation are available [5.35, 5.47]. Substantially

more detail on the calculation of thyroid doses to

individuals is provided in the dosimetry section of

the Chernobyl Forum report on health effects [5.3].

Generally, it can be stated that adequate

methodologies for thyroid dose reconstruction for

people who resided in the spring of 1986 in the

contaminated areas of Belarus, the Russian

Federation and Ukraine have been developed and

published. These estimates of both individual and

collective doses are being widely used by research

scientists and national health authorities both in

forecasts of thyroid morbidity and in radioepidemi-

ological studies.

5.3.4.2. Long term internal doses from terrestrial

pathways

Inhabitants of areas contaminated with radio-

nuclides in 1986 are still experiencing internal

exposure due to consumption of local foodstuffs

containing

137

Cs and, to a lesser extent,

Sr.

According to model estimates and direct human

measurements [5.39], inhalation of plutonium radio-

nuclides and

241

Am does not significantly contribute

to human dose in this context.

Generic dose conversion parameters have

been developed to reconstruct the past, assess the

current and forecast future average effective

internal doses. Examples for the adult rural

0 5 10 15 20

Relative dose

Age (years)

Male

Female

FIG. 5.13. Age–sex dependence of the mean thyroid dose to

inhabitants of a settlement standardized to the mean dose

to adults from the same settlement [5.48].

113

population of a settlement located in the interme-

diate (100 km < distance < 1000 km) zone of

contamination based on experimental data and

models developed in the Russian Federation and

Ukraine are given in Table 5.7 [5.9, 5.10, 5.15].

Values for each indicated time period are given

separately for various soil types as the ratios of the

mean internal dose (E) to the mean

137

Cs soil

deposition in a settlement as of 1986 (s

137

)

(µSv · kBq

·m

In a series of experimental whole body

measurements and associated annual internal dose

calculations it was found that long term doses to

children caused by ingestion of food containing

caesium radionuclides are usually lower by a factor

of about 1.1 to 1.5 than those to adults and

adolescents (see, for example, Refs [5.51, 5.52]).

The mean internal doses to residents of rural

settlements strongly depend on soil properties. For

assessment purposes, soils are classified into three

TABLE 5.6. DISTRIBUTION OF INDIVIDUAL THYROID DOSES FOR AGE GROUPS OF

CHILDREN AND ADOLESCENTS FROM THE KIEV, ZHYTOMYR AND CHERNIGOV REGIONS

OF UKRAINE, BASED ON IODINE-131 IN THYROID MEASUREMENTS [5.45]

Category and age group Number of measurements

Per cent of children with thyroid dose (Gy) in interval

£0.2 >0.2–1 >1–5 >5–10 >10

Settlements not evacuated

Rural areas

1–4 years 9 119 40 43 15 1.7 0.9

5–9 years 13 460 62 31 6.5 0.44 0.07

10–18 years 26 904 73 23 3.7 0.16 <0.01

Urban areas

1–4 years 5 147 58 33 7.5 1.0 0.7

5–9 years 11 421 82 15 2.6 0.23 0.04

10–18 years 24 442 91 7.7 1.4 0.12 <0.01

Evacuated settlements

1–4 years 1 475 30 45 22 2.7 1.0

5–9 years 2 432 55 36 8.4 0.6 0.08

10–18 years 4 732 73 23 3.6 0.13 0.02

TABLE 5.7. RECONSTRUCTION AND PROGNOSIS OF THE AVERAGE EFFECTIVE INTERNAL

DOSE TO THE ADULT RURAL POPULATION IN THE INTERMEDIATE (100 km < DISTANCE

< 1000 km) ZONE OF CHERNOBYL CONTAMINATION

Soil type

E/s

137

(mSv ◊ kBq

–1

◊ m

–2

137

Cs)

1986 1987–1995 1996–2005 2006–2056 1986–2056

Russian Federation

[5.9]

Soddy podzolic sandy

Black

180

Ukraine

[5.10, 5.15]

Peat bog 19 167 32 31 249

Sandy 19 28 5 5 57

Clay 19 17 3 3 42

Black 19 6 1 1 27

137

is given as for 1986.

114

major soil types: (a) black or chernozem soil; (b)

podzol soil (including both podzol sandy and podzol

loam soils); and (c) peat bog or peat soil. Due to the

environmental behaviour of

137

Cs, internal exposure

exceeds external dose in areas with peaty soil.

Contributions due to internal and external exposure

are comparable in areas with light sandy soil, and

the contribution of internal exposure to the total

(external and internal) dose does not exceed 10% in

areas with dominantly black soil. According to

numerous studies, the contribution of

Sr to the

internal dose regardless of natural conditions is

usually less than 5%.

The parameters obtained from independent

sets of Russian and Ukrainian data significantly

differ for some soil types and time periods (see

Table 5.7). Some of these discrepancies can be

explained by the different meteorological

conditions (mainly dry deposition in Ukraine and

wet deposition in the Russian Federation) that

occurred in different parts of the Chernobyl

affected areas and by different food consumption

habits.

Multiplication of the parameters presented in

Table 5.7 by the mean

137

Cs soil deposition (as of

1986) gives an estimate of the internal effective dose

caused by radiation from

137

Cs and

134

Cs (for the

Russian Federation, also from

Sr and

Sr) but not

from radioiodines. Dose estimates are given on the

assumption that countermeasures against internal

exposure were not applied. In broad terms, the most

important factors controlling internal dose to the

rural population are the dominant soil type and the

amount of

137

Cs deposition.

In towns and cities, internal dose is partially

determined by radioactive contamination of

foodstuffs produced in surrounding districts.

However, importation of foodstuffs from non-

contaminated areas has significantly reduced the

intake of radionuclides, and internal doses received

by urban populations are typically a factor of two to

three less than in rural settlements with an equal

level of radioactive contamination.

The deviation in dose to critical groups

compared with settlement average values varies by

a factor of about three for internal exposure. The

group most subjected to internal exposure from

137

Cs is adults consuming both locally produced

agricultural animal foods (e.g. milk, dairy products,

etc.) and natural foods (e.g. mushrooms, lake fish,

berries, etc.) in amounts exceeding average

consumption rates.

At present, inhabitants of areas of low

contamination (less than 0.04 MBq/m

137

Cs) are

receiving up to 0.004 mSv/a from ingestion of local

foods in black soil areas, up to 0.04 mSv/a in sandy

soil areas and about 0.1 mSv/a in villages located in

peaty soil areas. In the period 2002–2056, they will

receive an additional internal dose of less than

0.1 mSv in black soil areas, up to 0.7 mSv in sandy

soil areas and about 1–2 mSv in villages located in

peaty soil areas.

To avoid the presentation of dosimetric data

on a site by site basis, mean effective doses to adult

residents of rural and urban localities have been

determined as a function of soil

137

Cs deposition and

predominant soil type; such data are given in

Tables 5.8 and 5.9. The

137

Cs soil deposition is

subdivided into two ranges: 0.04–0.6 MBq/m

15 Ci/km

) and above 0.6 MBq/m

(i.e. 0.6–4

MBq/m

(15–100 Ci/km

)) in 1986. The level

0.04 MBq/m

is considered as a conventional border

between ‘non-contaminated’ and ‘contaminated’

areas. In areas contaminated with

137

Cs above

TABLE 5.8. PAST (1986–2000) AND FUTURE (2001–2056) MEAN CHERNOBYL RELATED

EFFECTIVE INTERNAL DOSES (mSv) TO ADULT RESIDENTS OF AREAS WITH CAESIUM-137

SOIL DEPOSITION ABOVE 0.04 MBq/m

(1 Ci/km

) IN 1986 [5.53]

Population

Caesium-137 in soil

(MBq/m

)

Soil type/time period

Black Podzol Peat

1986–2000 2001–2056 1986–2000 2001–2056 1986–2000 2001–2056

Rural 0.04–0.6 1–10 0.1–1 3–30 0.5–7 8–100 2–30

0.6–4 — 30–100 7–50 —

Urban 0.04–0.6 1–8 0.1–0.6 2–20 0.3–5 6–80 1–20

115

0.6 MBq/m

, application of active countermeasures

(i.e. agricultural restrictions, decontamination

measures, recommendations to restrict

consumption of locally gathered natural foods

(forest mushrooms and berries, lake fish, etc.)) has

been mandatory.

Dosimetric models predict that by 2001 the

residents had already received at least 75% of their

lifetime internal dose due to

137

Cs,

134

Cs,

Sr and

Sr (see Table 5.8). In the coming years (2001–

2056) they will receive the remaining 25% (i.e. less

than 1 mSv for black soil, up to 7 mSv for podzol soil

and up to 30 mSv for peat soil). In the more contam-

inated podzol soil areas, an effective dose of up to

50 mSv can still be expected.

As can be seen from Table 5.9, the more

elevated internal doses in some of the settlements

are above the national action level of 1 mSv/a. For

some population groups in contaminated areas, wild

foods (forest mushrooms, game, forest berries, fish)

can make an important contribution to dose [5.9,

5.15, 5.30]. Studies of

137

Cs intake of the rural

population in the Bryansk region of the Russian

Federation indicated that natural foods contributed

about 20% of total uptake in 1987, but up to 80% in

1994–1999 [5.29]. The relative contribution of wild

foods to internal dose has risen gradually because of

the substantial reduction of radionuclide content in

agricultural foods derived from vegetables and

animals, combined with a much slower decrease in

the contamination of wild foods. In the latter

period, the highest contributions to

137

Cs intake

(and, by inference, internal dose) came from forest

mushrooms, followed by forest berries, game and

lake fish.

Similar trends were found in residents of

Kozhany (Bryansk region), located on the coast of a

highly contaminated lake, where natural foods

contributed an average of 50–80% of

137

Cs intake

[5.30]. Men were more likely to eat natural foods

than women, and there was a positive correlation

between consumption of mushrooms and fish that

indicated a liking by many inhabitants for ‘gifts of

nature’. The average annual internal dose due to

137

Cs was estimated to be 1.2 mSv for men and 0.7

mSv for women in 1996.

5.3.4.3. Long term doses from aquatic pathways

Human exposure via the aquatic pathway

occurs as a result of consumption of drinking water,

fish and agricultural products grown using irrigation

water from contaminated water bodies. Use of

water bodies as a source of drinking water for

livestock and flooding of agricultural land can also

lead to human exposure via terrestrial pathways.

In the middle and lower areas of the Dnieper

River catchment, which were not significantly

subjected to direct radionuclide contamination in

1986, a significant proportion (10–20%) of the

Chernobyl related exposures was attributed to

aquatic pathways [5.53]. Although these doses were,

in fact, estimated to be very low, there was an

inadequate appreciation by the local population of

the risks of using water from contaminated aquatic

systems. This created an (unexpected) stress in the

population concerning the safety of the water

supply system. In areas close to Chernobyl,

radiation exposures via the aquatic pathway are

much higher, but are again minor in comparison

with terrestrial pathways.

Three pathways of exposure due to aquatic

systems need to be considered [5.53]:

(a) Consumption of drinking water from rivers,

lakes, reservoirs and wells in the contaminated

areas. The most significant exposures via

consumption of drinking water resulted from

the use of water from the Dnieper River basin,

and, in particular, the reservoirs of the

TABLE 5.9. ANNUAL (2001) MEAN CHERNOBYL RELATED EFFECTIVE

INTERNAL DOSES (mSv) TO ADULT RESIDENTS OF AREAS WITH

CAESIUM-137 SOIL DEPOSITION ABOVE 0.04 MBq/m

(1 Ci/km

) IN 1986 [5.53]

Population

Caesium-137 in soil

(MBq/m

)

Soil type

Black Podzol Peat

Rural 0.04–0.6 0.004–0.06 0.03–0.4 0.1–2

0.6–4 — 0.4–2 —

Urban 0.04–0.6 0.003–0.04 0.02–0.2 0.1–1

116

Dnieper River system. The Dnieper cascade is

a source of drinking water for more than eight

million people. The main consumers of

drinking water from the Dnieper River live in

the Dnipropetrovsk and Donetsk regions. In

Kiev, water from the Dnieper and Desna

Rivers is used by about 750 000 people. The

remaining part of the population use water

mainly derived from groundwater sources.

(b) Consumption of fish. The Dnieper River

reservoirs are used intensively for commercial

fishing. The annual catch is more than 25 000 t.

There was no significant decrease in fishing in

most of these reservoirs during the first

decade after the accident. During the first two

to three years, however, restrictions were

placed on the consumption of fish from the

Kiev reservoir. In some smaller lakes, both in

the former USSR and in parts of western

Europe, fishing was prohibited during the first

months and even years after the accident.

on land irrigated with water from the Dnieper

reservoirs. In the Dnieper River basin there is

more than 1.8 × 10

ha of irrigated agricultural

land. Almost 72% of this territory is irrigated

with water from the Kakhovka reservoir in the

Dnieper River–reservoir system. Accumu-

lation of radionuclides in plants in irrigated

fields can take place because of root uptake of

the radionuclides introduced with irrigation

water and due to direct incorporation of radio-

nuclides through leaves following sprinkler

irrigation. However, recent studies have

shown that, in the case of irrigated land in

southern Ukraine, irrigation water did not add

significant amounts of radioactive material to

crops in comparison with that which had been

initially deposited in atmospheric fallout and

subsequently taken up from the soil.

The contribution of aquatic pathways to the

dietary intake of

137

Cs and

Sr is usually quite small,

even in areas that were seriously affected by

Chernobyl fallout. For the relatively large rural

population that consumes fish from local rivers and

lakes, however, exposures could be significant. In

addition, collective doses to the large urban and

rural populations using water from the Pripyat–

Dnieper River–reservoir system were relatively

high. Owing to the high fallout within the catchment

of the Pripyat and Dnieper Rivers, this system has

been intensively monitored, and doses via aquatic

pathways have been estimated [5.53].

Contaminated rivers could potentially have

led to significant doses in the first months after the

accident through consumption of drinking water,

mainly through contamination with short lived

radionuclides. The most significant individual dose

was from

131

I and was estimated to be up to 0.5–

1.0 mSv for the citizens of Kiev during the first few

weeks after the Chernobyl accident [5.53].

After the end of the first month following the

accident, the main contributors to doses via aquatic

pathways became

137

Cs and

Sr. Estimated doses

due to these radionuclides in the Dnieper River–

reservoir system were made on the basis of

monitoring data and predictions of flood

frequencies. A worst case scenario of a series of high

floods during the first decade after the accident

(1986–1995) was assumed. Estimates were that

individual doses via aquatic pathways would not

have exceeded 1–5 mSv/a. Thus long term doses via

the drinking water pathway were small in

comparison with doses (mainly from short lived

radionuclides) in the early phase [5.53].

The contribution of different exposure

pathways to dose is shown in Fig. 5.14 for the village

of Svetilovichy in the Gomel region of Belarus. In

this case, consumption of freshwater fish forms an

important part of the diet, and hence doses via this

pathway can be significant for some individuals.

5.4. TOTAL (EXTERNAL AND INTERNAL)

EXPOSURE

The generalized data for both external and

internal (not including dose to the thyroid)

External from

soil

50.3%

External from

shore

4.0%

External from

swimming

0.01%

Internal from

food ingestion

23.9%

Internal from fish

ingestion

20.6%

Internal from

drinking water

1.2%

FIG. 5.14. Contributions of different pathways to the

effective dose to the critical group of the population of

Svetilovichy in the Gomel region of Belarus [5.53, 5.54].

117

exposures of the general public presented in

Tables 5.4 and 5.9, respectively, have been

summarized in Table 5.10 in order to estimate

broadly the mean individual total (external and

internal) effective doses accumulated by residents

of radioactively contaminated areas during 1986–

2000 and to forecast doses for 2001–2056. Table 5.11

gives estimates of the annual total dose in 2001. In

both tables data are given for levels of

137

Cs soil

deposition existing in 1986 in currently inhabited

areas of Belarus, the Russian Federation and

Ukraine, separately for rural and urban populations

and for different soil types, with no account taken of

current countermeasures. Both accumulated and

current annual total doses are presented for adults,

since, in the long term, children generally receive

lower external and internal doses from

137

environmental contamination (in contrast to

thyroid internal doses from radioiodine intake),

because of their occupancy (see Tables 5.3 and 5.5),

food habits and metabolic features.

As can be seen from Table 5.10, both

accumulated and predicted mean doses in

settlement residents vary over two orders of

magnitude depending on the radioactive contami-

nation of the area, soil type and settlement type.

Thus in 1986–2000 the dose range was from 2 mSv in

towns located in black soil areas, up to 300 mSv in

villages located in areas with podzol sandy soil.

According to the forecast, the doses expected in

2001–2056 are substantially lower than the doses

already received (i.e. in the range of 1–100 mSv). In

total, if countermeasures were not applied, the

populations of some of the more contaminated

villages in Belarus and the Russian Federation

would receive lifetime effective doses of up to

400 mSv, not including dose to the thyroid.

However, intensive application of countermeasures

such as settlement decontamination and agricultural

countermeasures has reduced dose levels by a factor

of about two. For comparison, a worldwide average

lifetime dose from natural background radiation is

about 170 mSv, with a typical range of 70–700 mSv

in various regions.

Based on local demographic data [5.51],

137

soil deposition maps (see Section 3.1) and the

TABLE 5.10. PAST (1986–2000) AND FUTURE (2001–2056) MEAN CHERNOBYL RELATED TOTAL

EFFECTIVE DOSES (mSv) TO ADULT RESIDENTS OF AREAS WITH CAESIUM-137 SOIL DEPOSI-

TION ABOVE 0.04 MBq/m

(1 Ci/km

) IN 1986 [5.53]

Population

Caesium-137

in soil (MBq/m

)

Soil type

Black Podzol Peat

1986–2000 2001–2056 1986–2000 2001–2056 1986–2000 2001–2056

Rural 0.04–0.6 3–40 1–14 5–60 1–20 10–150 3–40

0.6–4 — 60–300 20–100 —

Urban 0.04–0.6 2–30 1–9 4–40 1–13 8–100 2–20

TABLE 5.11. ANNUAL (2001) MEAN CHERNOBYL RELATED TOTAL EFFECTIVE DOSES (mSv) TO

ADULT RESIDENTS OF AREAS WITH CAESIUM-137 SOIL DEPOSITION ABOVE 0.04 MBq/m

(1 Ci/

) IN 1986 [5.53]

Population Caesium-137 in soil (MBq/m

)

Soil type

Black Podzol Peat

Rural 0.04–0.6 0.05–0.8 0.1–1 0.2–2

0.6–4 — 1–5 —

Urban 0.04–0.6 0.03–0.4 0.05–0.6 0.1–1

118

current level of countermeasure application (see

Section 4), the vast majority of the five million

people currently residing in the contaminated areas

of Belarus, the Russian Federation and Ukraine

(see Table 3.2) — that is, in the early 2000s —

receive annual effective doses of less than 1 mSv

(i.e. less than the national action levels in the three

countries). For comparison, a worldwide average

annual dose from natural background radiation is

about 2.4 mSv, with a typical range of 1–10 mSv in

various regions [5.1].

The number of residents of the contaminated

areas in the three most affected countries that

currently receive more than 1 mSv annually can be

estimated to be about 100 000 persons. As the future

reduction of both the external dose rate and radio-

nuclide (mainly

137

Cs) activity concentrations in

food will be rather slow (see Sections 5.2 and 3.3–

3.5), the reduction of human exposure levels is

expected to be slow (i.e. about 3–5%/a with

currently applied countermeasures).

5.5. COLLECTIVE DOSES

5.5.1. Thyroid

A summary of the collective doses to the

thyroid for the three most contaminated countries,

based on the thyroid dose reconstruction techniques

described in Section 5.3.4.1, is shown in Table 5.12.

The total thyroid collective dose is 1.6 × 10

man Gy,

with nearly half received by the group of persons

exposed in Ukraine. The present estimate of the

collective thyroid dose does not differ from that

made in Ref. [5.1].

5.5.2. Total (external and internal) dose from

terrestrial pathways

Estimates of collective dose accumulated in

1986–2005 via the terrestrial pathways of external

irradiation and ingestion of contaminated foods are

given in Table 5.13 for the three most affected

countries. The total collective dose was estimated to

be 43 000 man Sv in 1986–1995, including

24 000 man Sv from external exposure and 19 000

man Sv from internal exposure, according to

UNSCEAR [5.1], annex J, table 34. According to

the models of exposure dynamics presented above

[5.7], the estimated collective effective external

doses in 1986–2005 are about a factor of 1.2 higher,

and collective effective internal doses are higher by

a factor of 1.1–1.5 (depending on soil properties and

applied countermeasures), than those obtained in

1986–1995. In total, the collective dose increased by

9000 man Sv, or by 21%, during the second decade,

compared with the first decade, after the accident,

TABLE 5.12. COLLECTIVE THYROID DOSES

IN THE THREE COUNTRIES MOST CONTAM-

INATED BY THE CHERNOBYL ACCIDENT

[5.1]

Collective thyroid dose

(10

man Gy)

Russian Federation 300

Belarus 550

Ukraine 740

Total 1600

TABLE 5.13. ESTIMATED COLLECTIVE EFFECTIVE DOSES IN 1986–2005 TO THE POPULATIONS

OF THE CONTAMINATED AREAS OF BELARUS, THE RUSSIAN FEDERATION AND UKRAINE

(CAESIUM-137 SOIL DEPOSITION IN 1986 MORE THAN 37 kBq/m

)

Population

(millions of people)

Collective dose (10

man Sv)

External Internal Total

Belarus 1.9 11.9 6.8 18.7

Russian Federation 2.0 10.5 6.0 16.5

Ukraine 1.3 7.6 9.2 16.8

Total 5.2 30 22 52

Excluding thyroid dose. (Modified from Ref. [5.1], annex J, table 34, using dosimetric models presented in this report.)

119

and reached 52 000 man Sv. This is in good

agreement with the predictions made by

UNSCEAR in 1988 [5.2].

The recent estimate of collective dose based

on both human and environmental measurements

implicitly accounts for substantial, but not specified,

amounts of collective dose saved by the institution

of countermeasures that included evacuation,

relocation, prohibition on the use of foodstuffs and

longer term remediation of contaminated areas.

5.5.3. Internal dose from aquatic pathways

The most important aquatic system (the

Dnieper River basin) occupies a large area with a

population of about 32 million people who use the

water for drinking, fishing and irrigation. Estimates

have been made of the collective dose to people

from these three pathways for a period of 70 years

after the accident (i.e. from 1986 to 2056) [5.55,

5.56]. A long term hydrological scenario has been

analysed using a computer model [5.57]. Historical

data were used to account for the natural variability

in river flow. Dose assessment studies were carried

out to estimate the collective dose from the three

pathways [5.58]. The results of the calculations are

given in Table 5.14.

Dose estimates for the Dnieper River system

show that if there had been no action to reduce

radionuclide fluxes to the river, the collective dose

commitment for the population of Ukraine (mainly

due to radiocaesium and radiostrontium) could

have reached 3000 man Sv. Protective measures (see

Section 4) carried out during 1992–1993 on the left

bank floodplain of the Pripyat River decreased

exposure by approximately 700 man Sv. Other

protective measures on the right bank in the CEZ

(during 1999–2001) will further reduce collective

doses by 200–300 man Sv [5.59].

5.6. CONCLUSIONS AND

RECOMMENDATIONS

5.6.1. Conclusions

(a) The collective effective dose (not including

dose to the thyroid) received by about five

million residents living in the areas of Belarus,

the Russian Federation and Ukraine contami-

nated by the Chernobyl accident (

137

deposition on soil of >37 kBq/m

) was

approximately 40 000 man Sv during the

period 1986–1995. The groups of exposed

TABLE 5.14. COLLECTIVE DOSE COMMITMENT (CDC

) DUE TO STRONTIUM-90 AND

CAESIUM-137 FLOWING FROM THE PRIPYAT RIVER TO THE DNIEPER RIVER AND

DOWNSTREAM [5.56, 5.58]

Region

Population

(millions of people)

Strontium-90 CDC

(man-Sv)

Caesium-137 CDC

(man-Sv)

Ratio

Sr CDC

137

CDC

Chernigov 1.4 4 2 2

Kiev 4.5 290 190 1.5

Cherkassy 1.5 115 50 2.3

Kirovograd 1.2 140 40 3.5

Poltava 1.7 130 60 2.2

Dnepropetrovsk 3.8 610 75 8

Zaporozhe 2 320 35 9

Nikolaev 1.3 150 20 8

Kharkov 3.2 60 4 15

Lugansk 2.9 15 1 15

Donetsk 5.3 330 20 17

Kherson 1.2 100 20 5

Crimea 2.5 175 5 35

Total 32.5 2500 500 5

120

persons within each country received an

approximately equal collective dose. The

additional amount of collective effective dose

projected to be received between 1996 and

2006 is about 9000 man Sv.

(b) The collective dose to the thyroid was nearly

2×10

man Gy, with nearly half received by

persons exposed in Ukraine.

exposure were external exposure from radio-

nuclides deposited on the ground and the

ingestion of contaminated terrestrial food

products. Inhalation and ingestion of drinking

water, fish and products contaminated with

irrigation water were generally minor

pathways.

(d) The range in thyroid dose in different

settlements and in all age–gender groups is

large, between less than 0.1 Gy and more than

10 Gy. In some groups, and especially in

younger children, doses were high enough to

cause both short term functional thyroid

changes and thyroid cancer effects in some

individuals.

(e) The internal thyroid dose from the intake of

131

I was mainly due to the consumption of

fresh cow’s milk and, to a lesser extent, of

green vegetables; children, on average,

received a dose that was much higher than

that received by adults, because of their small

thyroid masses and consumption rates of fresh

cow’s milk that were similar to those of adults.

(f) For populations permanently residing in

contaminated areas and exposed predomi-

nantly via ingestion, the contribution of short

lived radioiodines (i.e.

132

133

I and

135

I) to

thyroid dose was minor (i.e. about 1% of the

131

I thyroid dose), since short lived radio-

iodines decayed during transport of the radio-

iodines along the food chains. The highest

relative contribution (20–50%) to the thyroid

dose to the public from short lived radionu-

clides was received by the residents of Pripyat

through inhalation; these residents were

evacuated before they could consume contam-

inated food.

(g) Both measurement and modelling data show

that the urban population was exposed to a

lower external dose by a factor of 1.5–2

compared with the rural population living in

areas with similar levels of radioactive

contamination. This arises because of the

better shielding features of urban buildings

and different occupational habits. Also, as the

urban population depends less on local

agricultural products and wild foods than the

rural population, both effective and thyroid

internal doses caused predominantly by

ingestion were lower by a factor of two to

three in the urban than in the rural

population.

(h) The initial high rates of exposure declined

rapidly due to the decay of short lived radio-

nuclides and to the movement of radio-

caesium into the soil profile. The latter caused

a decrease in the rate of external dose due to

increased shielding. In addition, as caesium

moves into the soil column it binds to soil

particles, which reduces the availability of

caesium to plants and thus to the human food

chain.

(i) The great majority of dose from the accident

has already been accumulated.

(j) Persons who received effective doses (not

including dose to the thyroid) higher than the

average by a factor of two to three were those

who lived in rural areas in single storey homes

and who ate large amounts of wild foods such

as game meats, mushrooms and berries.

(k) The long term internal doses to residents of

rural settlements strongly depend on soil

properties. Contributions due to internal and

external exposure are comparable in areas

with light sandy soil, and the contribution of

internal exposure to the total (external and

internal) dose does not exceed 10% in areas

with predominantly black soil. The contri-

bution of

Sr to the internal dose, regardless

of natural conditions, is usually less than 5%.

(l) The long term internal doses to children

caused by ingestion of food containing

caesium radionuclides are usually lower by a

factor of about 1.1–1.5 than those in adults and

adolescents.

(m) Both accumulated and predicted mean doses

in settlement residents vary in the range of

two orders of magnitude, depending on the

radioactive contamination of the area,

predominant soil type and settlement type. In

the period 1986–2000 the accumulated dose

ranged from 2 mSv in towns located in black

soil areas up to 300 mSv in villages located in

areas with podzol sandy soil. The doses

expected in the period 2001–2056 are substan-

tially lower than the doses already received

(i.e. in the range of 1–100 mSv).

121

(n) If countermeasures had not been applied, the

populations of some of the more contami-

nated villages could have received lifetime (70

years) effective doses of up to 400 mSv.

Intensive application of countermeasures such

as settlement decontamination and agricul-

tural countermeasures has substantially

reduced the doses. For comparison, a

worldwide average lifetime dose from natural

background radiation is about 170 mSv, with a

typical range of 70–700 mSv in various regions

of the world.

(o) The vast majority of the approximately five

million people residing in the contaminated

areas of Belarus, the Russian Federation and

Ukraine currently receive annual effective

doses of less than 1 mSv (equal to the national

action levels in the three countries). For

comparison, a worldwide average annual dose

from natural background radiation is about

2.4 mSv, with a typical range of 1–10 mSv in

various regions of the world.

(p) The number of residents of the contaminated

areas in the three most affected countries that

currently receive more than 1 mSv annually

can be estimated to be about 100 000 persons.

As the future reduction of both the external

dose rate and the radionuclide (mainly

137

Cs)

activity concentrations in food is predicted to

be rather slow, the reduction in the human

exposure levels is also expected to be slow (i.e.

about 3–5%/a with current countermeasures).

(q) Based upon available information, it does not

appear that the doses associated with hot

particles were significant.

(r) The assessment of the Chernobyl Forum

agrees with that of UNSCEAR [5.1] in terms

of the dose received by the populations of the

three most affected countries: Belarus, the

Russian Federation and Ukraine.

5.6.2. Recommendations

(a) Large scale monitoring of foodstuffs, whole

body counting of individuals and provision of

thermoluminescent dosimeters to members of

the general public are no longer necessary.

The critical groups in areas of high contami-

nation and/or high transfer of radiocaesium to

foods are known. Representative members of

these critical groups should be monitored with

dosimeters for external dose and with whole

body counting for internal dose.

(b) Sentinel or marker individuals in more highly

contaminated areas not scheduled for further

remediation might be identified for continued

periodic whole body counting and monitoring

of external dose. The goal would be to follow

the expected continued decrease in external

and internal dose and to determine whether

such decreases are due to radioactive decay

alone or to further ecological elimination.

REFERENCES TO SECTION 5

[5.1] UNITED NATIONS, Sources and Effects of

Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(2000).

[5.2] UNITED NATIONS, Sources, Effects and Risks

of Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(1988).

[5.3] WORLD HEALTH ORGANIZATION, Health

Effects of the Chernobyl Accident and Special

Health Care Programmes, Report of the

Chernobyl Forum Expert Group “Health”

(EGH), WHO, Geneva (in press).

[5.4] IZRAEL, Y. (Ed.), Atlas of Radioactive Contami-

nation of European Russia, Belarus and Ukraine,

Federal Service for Geodesy and Cartography of

Russia, Moscow (1998).

[5.5] DE CORT, M., Atlas of Caesium Deposition on

Europe after the Chernobyl Accident, Rep. 16733,

Office for Official Publications of the European

Communities, Luxembourg (1998).

[5.6] INTERNATIONAL COMMISSION ON RADI-

OLOGICAL PROTECTION, 1990 Recommen-

dations of the International Commission on

Radiological Protection, Publication 60,

Pergamon Press, Oxford and New York (1991).

[5.7] GOLIKOV, V.Y., BALONOV, M.I., JACOB, P.,

External exposure of the population living in areas

of Russia contaminated due to the Chernobyl

accident, Radiat. Environ. Biophys. 41 (2002) 185–

193.

[5.8] LIKHTAREV, I.A., KOVGAN, L.N., JACOB, P.,

ANSPAUGH, L.R., Chernobyl accident: Retro-

spective and prospective estimates of external

dose of the population of Ukraine, Health Phys. 82

(2002) 290–303.

[5.9] BALONOV, M., JACOB, P., LIKHTAREV, I.,

MINENKO, V., “Pathways, levels and trends of

population exposure after the Chernobyl acci-

dent”, The Radiological Consequences of the

Chernobyl Accident (Proc. Int. Conf. Minsk,

1996), Rep. EUR 16544 EN, Office for Official

Publications of the European Communities,

Luxembourg (1996) 235–249.

122

[5.10] LIKHTAREV, I.A., et al., Internal exposure from

the ingestion of foods contaminated by

137

Cs after

the Chernobyl accident — Report 2. Ingestion

doses of the rural population of Ukraine up to 12 y

after the accident (1986–1997), Health Phys. 79

(2000) 341–357.

[5.11] MECKBACH, R., JACOB, P., PARETZKE, H.,

Gamma exposures due to radionuclides deposited

in urban environments. Part I: Kerma rates from

contaminated urban surfaces, Radiat. Prot. Dosim.

25 (1988) 167–169.

[5.12] GOLIKOV, V.Y., personal communication, 2004.

[5.13] IZRAEL, Y.A., et al., Chernobyl: Radioactive

Contamination of the Environment, Gidromete-

oizdat, Leningrad (1990) (in Russian).

[5.14] MÜCK, K., et al., A consistent radionuclide vector

after the Chernobyl accident, Health Phys. 82

(2002) 141–156.

[5.15] JACOB, P., LIKHTAREV, I. (Eds), Pathway

Analysis and Dose Distributions, Joint Study

Project No. 5, Final Report, Rep. EUR 16541,

Office for Official Publications of the European

Communities, Luxembourg (1996).

[5.16] MILLER, K.M., KUIPER, I.L., HELFER, I.K.,

Cs-137 fallout depth distributions in forest versus

field sites: Implication for external dose rates, J.

Environ. Radioact. 12 (1990) 23–47.

[5.17] MECKBACH, R., JACOB, P., Gamma exposures

due to radionuclides deposited in urban environ-

ments. Part II: Location factors for different depo-

sition patterns, Radiat. Prot. Dosim. 25 (1988)

181–190.

[5.18] HEDEMANN-JENSEN, P., Shielding factors for

gamma radiation from activity deposited on struc-

tures and ground surfaces, Nucl. Technol. 68

(1985) 29–39.

[5.19] JACOB, P., MECKBACH, R., MILLER, H.M.,

Reduction of external exposures from deposited

Chernobyl radioactivity due to run-off, weath-

ering, street-cleaning and migration in the soil,

Radiat. Prot. Dosim. 21 (1987) 51–57.

[5.20] KARLBERG, O., Weathering and migration of

Chernobyl fallout in Sweden, Radiat. Prot. Dosim.

21 (1987) 75–78.

[5.21] JACOB, P., MECKBACH, R., “External exposure

from deposited radionuclides”, paper presented at

Proc. Sem. on Methods and Codes for Assessing

the Off-site Consequences of Nuclear Accidents,

Athens, 1990.

[5.22] JACOB, P., MECKBACH, R., Measurements

after the Chernobyl Accident in Relation to an

Exposure of an Urban Population, IAEA-

TECDOC-1131, IAEA, Vienna (2000) 34–41.

[5.23] JACOB, P., ROSENBAUM, H., PETOUSSI, N.,

ZANKL, M., Calculation of Organ Doses from

Environmental Gamma Rays Using Human

Phantoms and Monte Carlo Methods, Part II:

Radionuclides Distributed in the Air or Deposited

on the Ground, GSF-Bericht 12/90, National

Research Center for Environment and Health,

Munich (1990).

[5.24] ERKIN, V.G., LEBEDEV, O.V., “Thermolumi-

nescent dosimeter measurements of external

doses to the population of the Bryansk region

after the Chernobyl accident”, The Chernobyl

Papers (BALONOV, M.I., MERWIN, S.E., Eds),

Research Enterprises, Richland, WA (1993) 289–

311.

[5.25] SKRYABIN, A.M., et al., Distribution of Doses

Received in Rural Areas Affected by the

Chernobyl Accident, Rep. NRPB-R277, National

Radiological Protection Board, Didcot, UK

(1995).

[5.26] LIKHTAREV, I.A., et al., Effective doses due to

external irradiation from the Chernobyl accident

for different population groups of Ukraine, Health

Phys. 70 (1996) 87–98.

[5.27] CHUMAK, V. V. , LIKHTAREV, I . A .,

PAVLENKO, J.V., Monitoring of individual doses

of populations residing in the territories contami-

nated after the Chernobyl accident, Radiat. Prot.

Dosim. 85 (1999) 137–139.

[5.28] GOLIKOV, V., BALONOV, M., ERKIN, V.,

JACOB, P., Model validation for external doses

due to environmental contaminations by the

Chernobyl accident, Health Phys. 77 (1999) 654–

661.

[5.29] BRUK, G.Y., et al., “The role of the forest

products in the formation of internal exposure

doses to the population of Russia after the

Chernobyl accident”, Contaminated Forests:

Recent Developments in Risk Identification and

Future Perspectives (LINKOV, I., SCHELL,

W.R., Eds), NATO Science Series, Vol. 58, Kluwer,

Dordrecht (1999) 343–352.

[5.30] TRAVNIKOVA, I.G., et al., Lake fish as the main

contributor of internal dose to lakeshore residents

in the Chernobyl contaminated area, J. Environ.

Radioact. 77 (2004) 63–75.

[5.31] INTERNATIONAL COMMISSION ON RADI-

OLOGICAL PROTECTION, Human Respira-

tory Tract Model for Radiological Protection,

Publication 66, Pergamon Press, Oxford and New

York (1994).

[5.32] INTERNATIONAL COMMISSION ON RADI-

OLOGICAL PROTECTION, The ICRP

Database of Dose Coefficients: Workers and

Members of the Public, ICRP CD-ROM System,

Version 2.01, Pergamon Press, Oxford and New

York (2001).

[5.33] LIKHTAREV, I.A., SHANDALA, N.K.,

GULKO, G.M., KAIRO, I.A., CHEPURNY, N.I.,

Ukrainian thyroid doses after the Chernobyl

accident, Health Phys. 64 (1993) 594–599.

[5.34] ZVONOVA, I.A., BALONOV, M.I., “Radioio-

dine dosimetry and prediction of consequences of

thyroid exposure of the Russian population

following the Chernobyl accident”, The

Chernobyl Papers (MERWIN, S.E., BALONOV,

123

M.I., Eds), Research Enterprises, Richland, WA

(1993) 71–125.

[5.35] GAVRILIN, Y.I., et al., Chernobyl accident:

Reconstruction of thyroid dose for inhabitants of

the Republic of Belarus, Health Phys. 76 (1999)

105–119.

[5.36] BALONOV, M.I., et al., “Long term exposure of

the population of the Russian Federation as a

consequence of the accident at the Chernobyl

power plant”, Environmental Impact of Radioac-

tive Releases (Proc. Int. Conf. Vienna, 1995),

IAEA, Vienna (1995) 397–411.

[5.37] BRUK, G.Y., SHUTOV, V.N., BALONOV, M.I.,

BASALEYAVA, L.N., KISLOV, M.V., Dynamics

137

Cs content in agricultural food products

produced in regions of Russia contaminated after

the Chernobyl accident, Radiat. Prot. Dosim. 76

(1998) 169–178.

[5.38] FESENKO, S.V., et al., Regularities of changing in

137

Cs activity concentrations in the long term after

the accident at the Chernobyl NPP, Radiat. Biol.

Radioecol. 44 (2004) 35–49.

[5.39] IVANOVA, N.P., SHVYDKO, N.S., ERSHOV,

E.B., BALONOV, M.I., Population doses in

Russia from plutonium fallout following the

Chernobyl accident, Radiat. Prot. Dosim. 58

(1995) 255–260.

[5.40] KUTKOV, V., SKRAYBIN, A., POGODIN, R.,

AREFIEVA, Z., MURAVYEV, Y., “Inhalation

of the aerosol of nuclear fuel particles from the

Chernobyl nuclear power plant by adult persons

from the Gomel region of Belarus”, Environ-

mental Impact of Radioactive Releases (Proc. Int.

Conf. Vienna, 1995), IAEA, Vienna (1995) 107–

115.

[5.41] BALONOV, M.I., TRAVNIKOVA, I.G., “The

role of agricultural and natural ecosystems in the

internal dose formation in the inhabitants of a

controlled area”, Transfer of Radionuclides in

Natural and Semi-natural Environments

(DESMET, G., NASSIMBENI, P., BELLI, M.,

Eds), Elsevier, London and New York (1990) 419–

430.

[5.42] THOMAS, G., KARAOGLOU, A., WILLIAMS,

E.D. (Eds), Radiation and Thyroid Cancer, World

Scientific, Singapore (1999).

[5.43] ZVONOVA, I., et al., “Methodology of thyroid

dose reconstruction for population of Russia after

the Chernobyl accident”, Harmonization of Radi-

ation, Human Life and the Ecosystem (Proc. Int.

Congr. Hiroshima, 2000), International Radiation

Protection Association, Fontenay-aux-Roses

(2000) P-11-265.

[5.44] KRUK, J.E., PROHL, G., KENIGSBERG, J.A.,

A radioecological model for thyroid dose recon-

struction of the Belarus population following the

Chernobyl accident, Radiat. Environ. Biophys. 43

(2004) 101–110.

[5.45] LIKHTAREV, I.A., et al., Post-Chernobyl thyroid

cancers in Ukraine, Report 1: Estimation of

thyroid doses, Radiat. Res. 163 (2005) 125–136.

[5.46] INTERNATIONAL COMMISSION ON RADI-

OLOGICAL UNITS AND MEASUREMENTS,

Retrospective Assessment of Exposures to

Ionising Radiation, J. Int. Comm. Radiat. Prot. 2 2

(2002) 1–136.

[5.47] BALONOV, M.I., ZVONOVA, I.A. (Eds), Mean

thyroid doses for inhabitants of different age who

lived in 1986 in settlements of the Bryansk, Tula,

Orel and Kaluga regions contaminated by radio-

nuclides as a result of the Chernobyl accident,

Radiat. Risk, Special issue (2002) (in Russian).

[5.48] HEIDENREICH, W., et al., Age- and sex-specific

relative thyroid radiation exposure to

131

I in

Ukraine after the Chernobyl accident, Health

Phys. 80 (2001) 242–250.

[5.49] BALONOV, M., et al., Contributions of short-

lived radioiodines to thyroid doses received by

evacuees from the Chernobyl area estimated using

early in-vivo activity measurements, Radiat. Prot.

Dosim. 105 (2003) 593–600.

[5.50] GAVRILIN, Y., et al., Individual thyroid dose

estimation for a case-control study of Chernobyl-

related thyroid cancer among children of Belarus,

Part 1:

131

I, short-lived radioiodines (

132

133

135

I),

and short-lived radiotelluriums (

131m

Te and

132

Te ),

Health Phys. 86 (2004) 565–585.

[5.51] INTERNATIONAL ADVISORY COMMIT-

TEE, The International Chernobyl Project:

Technical Report, IAEA, Vienna (1991).

[5.52] SHUTOV, V.N., BRUK, G.Y., BALONOV, M.I.,

PARHOMENKO, V.I., PAVLOV, I.J., “Caesium

and strontium radionuclide migration in the agri-

cultural ecosystem and estimation doses to the

population”, The Chernobyl Papers (MERWIN,

S.E., BALONOV, M.I., Eds), Research Enter-

prises, Richland, WA (1993).

[5.53] INTERNATIONAL ATOMIC ENERGY

AGENCY, Radiological Conditions in the

Dnieper River Basin, IAEA, Vienna (2006).

[5.54] DROZDOVITCH, V.V., MINENKO, V.F.,

KUKHTA, T.S., ULANOVSKY, A.V.,

BOUVILLE, A., personal communication, 2003.

[5.55] ZHELEZNYAK, M., et al., “Modelling of radio-

nuclides transport in the set of river reservoirs”,

Computational Methods in Water Resources, X,

Vol. 2 (PETERS, A., et al., Eds), Kluwer,

Dordrecht (1994) 1189–1196.

[5.56] BERKOVSKI, V., VOITSEKHOVITCH, O.,

NASVIT, O., ZHELEZNYAK, M., SANSONE,

U., “Exposures from aquatic pathways”, The

Radiological Consequences of the Chernobyl

Accident (Proc. Int. Conf. Minsk, 1996), Rep.

EUR 16544 EN, Office for Official Publications of

the European Communities, Luxembourg (1996)

283–294.

[5.57] ZHELEZNYAK, M.K., et al., Mathematical

modeling of radionuclides dispersion in the

124

Pripyat–Dnieper aquatic system after the

Chernobyl accident, Sci. Total Environ. 112 (1992)

89–114.

[5.58] BERKOVSKI, V., RATIA, G., NASVIT, O.,

Internal doses to Ukrainian population using

Dnieper river water, Health Phys. 71 (1996) 37–44.

125

6. RADIATION INDUCED EFFECTS

ON PLANTS AND ANIMALS

6.1. PRIOR KNOWLEDGE OF RADIATION

EFFECTS ON BIOTA

The effects of radiation on plants and animals

have long been of interest to scientists; in fact, much

of the information on the effects on humans has

evolved from studies on plants and animals.

Additional research followed the development of

nuclear energy and concerns about the possible

impacts of increased, but authorized, discharges of

waste radionuclides into the terrestrial and aquatic

environments. The magnitude of these authorized

releases has always been controlled on the basis of

the limitation of human exposure, but it has been

recognized that animals and plants have also been

exposed — frequently to a higher degree than

humans. By the mid-1970s, sufficient information

had been accrued on the effects of ionizing

radiation on plants and animals that several author-

itative reviews had been compiled to summarize the

findings [6.1–6.4].

Some broad generalizations about the effects

of radiation exposure can be gleaned from the

research that has been conducted over the past 100

years. Foremost are the relatively large differences

in doses required to cause lethality among various

taxonomic groups (Fig. 6.1). Considerable variation

in response occurs within a taxon due to enhanced

radiosensitivity of some individuals or life stages.

Wide ranges in doses are also observed within a

group or taxon when progressing from minor to

severe effects.

Figure 6.2 summarizes information on the

doses required to be delivered over a short time

period to produce damage of different degrees in

various plant communities, soil invertebrates and

rodents. Within the plant kingdom, trees are

generally more sensitive than shrubs, which in turn

are more sensitive than herbaceous species.

Primitive forms such as lichens, mosses and

liverworts are more resistant than vascular plants.

Radiation resistant plants frequently have

molecular and cellular characteristics that enhance

their ability to tolerate radiation stress, and

differences in plant community response can be

explained, in part, by the early work of Sparrow

[6.8]. He showed that characteristics such as large

chromosomes, normal (rather than diffuse) centro-

meres, small chromosome number, uninucleated

cells, diploid or haploid cells, sexual reproduction,

long intermitotic time and slow rates of meiosis are

associated with high radiosensitivity in plants, but

that sensitivity can be modified in time due to

seasonal processes (e.g. dormancy or the onset of

growth in spring; Table 6.1).

Scientific reviews (e.g. Ref. [6.3]) have

indicated that mammals are the most sensitive

organisms and that reproduction is a more sensitive

endpoint than mortality. For acute exposures of

Viruses

Molluscs

Protozoa

Bacteria

Moss, lichen, algae

Insects

Crustaceans

Reptiles

Amphibians

Fish

Higher plants

Birds

Mammals

1 10 100 1000 10 000

Acute lethal dose range (Gy)

FIG. 6.1. Acute dose ranges that result in 100% mortality in

various taxonomic groups. Humans are among the mos

ensitive mammals, and therefore among the most sensitive

organisms [6.5].

Moss–lichen

Grassland

Tropical rain forest

Old fields

Shrubs

Deciduous forest

Soil invertebrates

Rodents

Coniferous forest

0.01 0.1 1 10 100 1000 10 000

Minor effects

Intermediate to severe effects

Dose (Gy)

FIG. 6.2. Range of short term radiation doses (delivered

over 5–60 d) that produced effects in various plant commu-

nities, rodents and soil invertebrates. Minor effects include

chromosomal damage, changes in productivity, reproduc-

tion and physiology. Intermediate effects include changes

in species composition and diversity through selective

mortality. Severe effects (massive mortality) begin at the

upper range of intermediate effects [6.6, 6.7].

126

mammals, mortality generally occurs at doses above

3 Gy, while reproduction is affected at doses below

0.3 Gy. Chronic exposures alter the responses, with

mortality occurring at greater than 0.1 Gy/d and

reproduction effected at less than 0.01 Gy/d.

Among aquatic organisms, fish are the most

sensitive, with gametogenesis and embryo

development being the more sensitive stages.

Effects on animal populations can be reduced by

their mobility (in terms of moving from areas of

high exposure to areas of low exposure). Compara-

tively stationary soil invertebrates do not have such

abilities and can receive substantial doses relative to

the rest of the animal kingdom, particularly because

soil is a sink for most radioactive contamination.

The response of a plant or animal to radiation

depends on the dose received as well as its radiosen-

sitivity. The former is largely determined by its

habitat preference in relation to the evolving distri-

bution of radioactive contaminants as a function of

time, as well as the organism’s propensity to

accumulate radionuclides in its organs and tissues.

Owing to their particular use of the habitat, plants

and animals within a contaminated area may

receive radiation doses that can be substantially

higher than those of humans occupying the same

area (e.g. humans gain some shielding from housing

and may obtain food and water from less contami-

nated sources [6.3]).

Although all exposures to ionizing radiation

have the potential to damage biological tissue,

protraction of a given total absorbed dose in time

can, depending on the dose rate, result in a

reduction in response due to the intervention of

cellular and tissue repair processes. This has led to

the conventional, but somewhat artificial,

distinction between so called acute and chronic

radiation exposure regimes. In general, an acute

radiation exposure is one that usually occurs at a

high dose rate and in a short period of time relative

to that within which obvious effects occur. Chronic

exposures are taken to be continuous in time, often

over a significant portion of an organism’s lifespan,

or throughout some particular life stage (e.g.

embryonic development) and usually at a suffi-

ciently low dose rate that the cumulative dose does

not produce acute effects.

The earlier reviews noted above were

consistent in concluding that it is unlikely that there

will be any significant detrimental effects:

(a) To terrestrial and aquatic plant populations,

and aquatic animal populations at chronic

dose rates of less than 10 mGy/d; or

(b) To terrestrial animal populations at dose rates

of less than 1 mGy/d.

It should be emphasized, however, that these

dose rates were not intended for use as limits in any

system to provide for the protection of the

environment; they were simply the dose rates below

which the available evidence, admittedly limited in

the range of organisms and biological responses

investigated, indicated little likelihood of any

TABLE 6.1. PRINCIPAL NUCLEAR CHARACTERISTICS AND FACTORS

INFLUENCING THE SENSITIVITY OF PLANTS TO RADIATION [6.5, 6.8]

Factors increasing sensitivity Factors decreasing sensitivity

Large nucleus (high DNA content) Small nucleus (low DNA content)

Much heterochromatin Little heterochromatin

Large chromosomes Small chromosomes

Acrocentric chromosomes Metacentric chromosomes

Normal centromere Polycentric or diffuse centromere

Uninucleate cells Multinucleate cells

Low chromosome number High chromosome number

Diploid or haploid nuclei High polypolid

Sexual reproduction Asexual reproduction

Long intermitotic time Short intermitotic time

Long dormant period Short or no dormant period

Slow meiosis Fast meiosis

127

significant response. The above dose rates are with

reference to population level effects, not to impacts

on individual organisms.

More recent reviews of the effects of radiation

exposure on individual organisms carried out in the

framework of two EC projects, FASSET

(Framework for the Assessment of Environmental

Impact) and EPIC (Environmental Protection from

Ionising Contaminants in the Arctic), have

produced broadly consistent conclusions [6.9–6.11].

Although minor effects may be seen at lower dose

rates in sensitive cell systems or individuals of

sensitive species (e.g. haematological cell counts in

mammals, immune response in fish, growth in pines

and chromosome aberrations in many organisms),

the threshold dose rate for significant effects in

most studies is about 0.1 mGy/h (2.4 mGy/d).

Detrimental responses then increase progressively

with increasing dose rate and usually become clear

at greater than 1 mGy/h (24 mGy/d) given over a

large fraction of the lifespan. The significance of the

minor morbidity and cytogenetic effects on the

individual, or on populations more generally, seen

at dose rates of less than 2.4 mGy/d has yet to be

determined [6.11].

The recently compiled EPIC database covers

a very wide range of radiation dose rates (from

below 10

–5

Gy/d up to more than 1 Gy/d) to wild

flora and fauna observed in northern parts of the

Russian Federation and in the Chernobyl contami-

nated areas [6.10]. The general conclusion from the

EPIC database is that the threshold for determin-

istic radiation effects in wildlife lies somewhere in

the range of 0.5–1 mGy/d for chronic low linear

energy transfer radiation.

These broad conclusions concerning the

impact of radiation on plants and animals provide

an appropriate context within which to consider the

available information on the effects that have been

observed from the increased radiation exposures

following the accident at Chernobyl.

6.2. TEMPORAL DYNAMICS OF

RADIATION EXPOSURE FOLLOWING

THE CHERNOBYL ACCIDENT

It is critical to frame any discussion of

Chernobyl environmental effects within the specific

time period of interest. Effects observed now, nearly

20 years after the accident, are drastically different

from those that occurred during the first 20 days.

Three distinct phases of radiation exposure have

been identified in the area local to the accident

[6.4]. In the first 20 days, radiation exposures were

essentially acute because of the large quantities of

short lived radionuclides present in the passing

cloud of contamination (

Mo,

132

Te /

132

133

Xe,

131

and

140

Ba/

140

La). Most of these short lived, highly

radioactive nuclides were deposited on to plant and

ground surfaces, resulting in the accumulation of

large doses that measurably affected biota. High

exposures of the thyroids of vertebrate animals also

occurred during the first days and weeks following

the accident from the inhalation and ingestion of

radioactive iodine isotopes or their radioactive

precursors.

The measured exposure rates on the day of

the accident in the immediate vicinity of the

damaged reactor are shown in Fig. 6.3. These

exposure rates were mainly due to gamma radiation

from deposited radionuclides and range up to about

20 Gy/d. However, for surface tissues and small

biological targets (e.g. mature needles and growing

buds of pine trees), there was a considerable

additional dose rate from the beta radiation of the

deposited radionuclides. Taking into account the

high dose rates during the relatively short exposure

period from the short lived radioisotopes, this first

phase of 20–30 days can be generally characterized

as an acute exposure regime that had pronounced

effects on biota.

The second phase of radiation exposure

extended through the summer and autumn of 1986,

during which time the short lived radionuclides

decayed and longer lived radionuclides were

transported to different components of the

environment by physical, chemical and biological

100

COOLING POND

1 km

TOWN OF PRIPYAT

REACTOR

100

0.1

RIVER PRIPYAT

FIG. 6.3. Measured exposure rates in air on 26 April 1986

in the local area of the Chernobyl reactor. Units of isolines

are R/h (1 R/h is approximately 0.2 Gy/d) [6.12].

128

processes. Dominant transport processes included

rain induced transfer of radionuclides from plant

surfaces on to soil, and bioaccumulation through

plant tissues. Although the dose rates at the soil

surface declined to much less than 10% of the initial

values, due to radioactive decay of the short lived

isotopes (see Fig. 5.3), damaging total doses were

still accumulated. The modifying effect of radionu-

clide wash-off by rain on radiation damage of

conifers is shown in Fig. 6.4.

In general, approximately 80% of the total

radiation dose accumulated by plants and animals

was received within three months of the accident,

and over 95% of this was due to beta radiation [6.4].

This finding agrees with earlier studies on the

importance of beta radiation, relative to the gamma

component, to the total dose from radioactive

fallout; for example, when a 10 h old mixture of

fresh fission products was experimentally deposited

on cereal plants at differing stages of growth, at a

density of 7 GBq/m

, the ratio of resulting beta and

gamma dose rates, measured with thermolumi-

nescent dosimeters, varied from 1 to 130 [6.13].

Measurements made with thermoluminescent

dosimeters on the soil surface at sites within the

CEZ indicated that the ratio of beta dose to gamma

dose was about 26:1 (i.e. 96% of the total dose was

from beta radiation). For a gamma dose rate of

0.01 mGy/h at the soil surface, 15 days after the

accident, the total cumulative dose in the first

month from beta and gamma radiation was

estimated to be 0.5 ± 0.2 Gy, and 0.6 and 0.7 Gy at

the end of the second and third months, respectively

[6.14].

In the third (and continuing) phase of

radiation exposure, dose rates have been chronic,

less than 1% of the initial values, and derived

mainly from

137

Cs contamination. With time, the

decay of the short lived radionuclides and the

migration of much of the remaining

137

Cs into the

soil have meant that the contributions to the total

radiation exposure from beta and gamma

radiations have tended to become more

comparable. The balance does depend, however, on

the degree of bioaccumulation of

137

Cs in organisms

and the behaviour of the organism in relation to the

main source of external exposure (i.e. the soil).

Aside from the spatial heterogeneity in the dose

rate arising from the initial deposition, large

variations in the radiation exposure of different

organisms occurred at different times due to their

habitat niche (e.g. birds in the canopy versus

rodents on the ground). Immigration of animals

into the CEZ and reproduction of those plants and

animals that are present means that new animals

and plants are constantly being introduced into the

radioactively contaminated conditions that exist

around Chernobyl today. The current conditions

are presented in Section 6.8.

6.3. RADIATION EFFECTS ON PLANTS

Doses received by plants from the Chernobyl

fallout were influenced by the physical properties of

the various radionuclides (i.e. their half-lives,

radiation emissions, etc.), the physiological stage of

the plant species at the time of the accident and the

different species dependent propensities to take up

radionuclides into critical plant tissues. The

FIG. 6.4. A small conifer, the upper portion damaged by

the initial deposition of radiation on to the crown of the

lant, and the lower part of the plant damaged by the irra-

diation of surface deposited material subsequently washed

rom the plant’s crown, leaving the middle section of the

lant unaffected (photograph courtesy of T. Hinton, 1991).

129

occurrence of the accident in late April heightened

the damaging effects of the fallout because it

coincided with the period of accelerated growth and

reproduction in plants. The deposition of beta

emitting contamination on critical plant tissues

resulted in their receiving a significantly larger dose

than animals living in the same environment [6.13,

6.15]. Large apparent inconsistencies in dose–

response observations occurred when the beta

radiation component was not appropriately

accounted for [6.16].

Within the CEZ, deposition of total beta

activity and associated doses to plants were

sufficient (0.7–3.9 GBq/m

) to cause short term

sterility and reduction in the productivity of some

species [6.15]. By August 1986, crops that had been

sown prior to the accident began to emerge. Growth

and development problems were observed in plants

growing in fields with contamination densities of

0.1–2.6 GBq/m

and with estimated dose rates

initially received by plants reaching 300 mGy/d.

Spot necroses on leaves, withered tips of leaves and

inhibition of photosynthesis, transpiration and

metabolite synthesis were detected, as well as an

increased incidence of chromosome aberrations in

meristem cells [6.17]. The frequency of various

anomalies in winter wheat exceeded 40% in 1986–

1987, with some abnormalities apparent for several

years afterwards [6.18].

Coniferous trees were already known to be

among the more radiosensitive plants, and pine

forests 1.5–2 km west of the Chernobyl nuclear

power plant received a sufficient dose (>80 Gy) to

cause mortality [6.19] at dose rates that exceeded

20 Gy/d [6.12]. The first signs of radiation injury in

pine trees in close proximity to the reactor were

yellowing and needle death, which appeared within

two to three weeks. During the summer of 1986 the

area of radiation damage expanded in the north-

west direction up to 5 km; serious damage was

observed at a distance of 7 km. The colour of the

dead pine stands resulted in the forest being

referred to as the Red Forest.

Tikhomirov and Shcheglov [6.19] and

Arkhipov et al. [6.20] found that mortality rate,

reproduction anomalies, stand viability and re-

establishment of pine tree canopies were dependent

on absorbed dose. Acute irradiation of Pinus

silvestris at doses of 0.5 Gy caused detectable

cytogenetic damage; at more than 1 Gy, growth

rates were reduced and morphological damage

occurred; and at more than 2 Gy, the reproductive

abilities of trees were altered. Doses of less than 0.1

Gy did not cause any visible damage to the trees.

Table 6.2 shows the variation in activity concen-

tration and dose among pine trees within the CEZ.

The radiosensitivity of spruce trees was observed to

be greater than that of pines. At absorbed doses as

low as 0.7–1 Gy, spruce trees had malformed

needles, buds and shoot growth [6.22].

Of the absorbed dose to critical parts of trees,

90% was due to beta radiation from the deposited

radionuclides and 10% to gamma radiation. As

early as 1987, recovery processes were evident in

the surviving tree canopies and young forests were

re-established in the same place as the perished

trees by replanting in reclamation efforts [6.20]. In

the decimated pine stands, a sudden invasion of

pests occurred that later spread to adjoining areas.

The deceased pine stands have now been replaced

by grassland, with a slow invasion of self-seeding

deciduous trees. Four distinct zones of radiation

TABLE 6.2. RADIOACTIVE CONTAMINATION (kBq/kg) OF CONIFEROUS TREES AS A

FUNCTION OF DISTANCE FROM THE CHERNOBYL REACTOR (AZIMUTH 205–260°), WITH

CORRESPONDING ESTIMATES OF THE AIR DOSE RATE (mGy/h) IN OCTOBER 1987 AND THE

ACCUMULATED EXTERNAL DOSE (Gy) [6.21]

Distance from

the Chernobyl

nuclear power

plant (km)

Air dose

rate

(mGy/h)

External

dose

(Gy)

Activity concentration in needles (kBq/kg)

Cerium-144 Ruthenium-106 Zirconium-95 Niobium-95 Caesium-134 Caesium-137

2.0 2.2 126 13 400 4100 800 1500 1500 4100

4.0 0.10 5 150 60 8 15 17 72

16.0 3.5 × 10

–4

0.014 1.5 0.6 0.1 0.17 0.18 0.55

Dose rate and dose of gamma radiation at 1 m height from the soil surface.

130

induced damage to conifers were discernable

(Table 6.3).

6.4. RADIATION EFFECTS ON SOIL

INVERTEBRATES

Although between 60% and 90% of the initial

fallout was captured by the forest canopy and other

plants [6.19], within weeks to a few months the

processes of wash-off by rain and leaf fall moved the

majority of the contamination to the litter and soil

layers (see Section 3.4 for more details), where soil

and litter invertebrates were exposed to high

radiation levels for protracted time periods. The

potential for impact on soil invertebrates was partic-

ularly large, since the timing of the accident

coincided with their most radiosensitive life stages:

reproduction and moulting following their winter

dormancy.

Within two months of the accident, the

number of invertebrates in the litter layer of forests

3–7 km from the nuclear reactor was reduced by a

factor of 30 [6.14], and reproduction was strongly

affected (larvae and nymphs were absent). Doses of

approximately 30 Gy (estimated from thermolumi-

nescent dosimeters placed in the soil) had

catastrophic effects on the invertebrate community,

causing mortality of eggs and early life stages, as

well as reproductive failure in adults. Within a year,

reproduction of invertebrates in the forest litter

resumed, due, in part, to the migration of inverte-

brates from less contaminated sites. After 2.5 years,

the ratio of young to adult invertebrates in the litter

layer, as well as the total mass of invertebrates per

unit area, was no different from control sites;

however, species diversity remained markedly

lower [6.14].

The diversity of invertebrate species within

the soil facilitates an analysis of community level

effects (i.e. changes in species composition and

abundance); for example, only five species of inver-

tebrates were found in ten soil cores taken from

pine stands in July 1986 at a distance of 3 km from

the Chernobyl nuclear power plant, compared with

23 species at a control site 70 km away. The mean

density of litter fauna was reduced from 104

individuals per 225 cm

core at the control location

to 2.2 at the 3 km site. Six species were found in all

ten cores taken from the control site, whereas no

species was found in all ten cores from the 3 km

location [6.23]. The number of invertebrate species

found at the heavily contaminated sites was only

half that of controls in 1993, and complete species

diversity did not recover until 1995, almost ten years

after the accident [6.14].

Compared with invertebrates within the forest

litter layer, those residing in arable soil were not

affected so much. A fourfold reduction in

earthworm numbers was found in arable soils, but

no catastrophic mortality of any group of soil inver-

tebrates was observed. There was no reduction in

soil invertebrates below a 5 cm depth in the soil.

Radionuclides had not yet migrated into deeper soil

layers, and the overlying soil shielded the inverte-

brates from beta radiation, the main contributor

(94%) to the total dose. The dose to invertebrates in

forest litter was threefold to tenfold higher than that

to those residing in surface soil [6.14].

Although researchers are unclear if sterility of

invertebrates occurred in the heavily contaminated

TABLE 6.3. ZONES AND CORRESPONDING DAMAGE TO CONIFEROUS FORESTS IN THE AREA

AROUND THE CHERNOBYL NUCLEAR POWER PLANT [6.22]

Zone and classification

External gamma dose

(Gy)

Air dose rate

(mGy/h)

Internal dose to

needles (Gy)

Conifer death (4 km

): complete death of pines,

partial damage to deciduous trees

>80–100 >4 >100

Sublethal (38 km

): death of most growth points,

death of some coniferous trees, morphological changes

to deciduous trees

10–20 2–4 50–100

Medium damage (120 km

): suppressed reproductive

ability, dried needles, morphological changes

4–5 0.4–2 20–50

Minor damage: disturbances in growth,

reproduction and morphology of coniferous trees

0.5–1.2 <0.2 <10

Dose rate and dose of gamma radiation at 1 m height from the soil surface.

131

sites at Chernobyl [6.14], the 30 Gy cumulative dose

reported for Chernobyl field studies is within the

range of experimental doses used to control pest

insects by external irradiation. A recent review

indicated that most insect, mite and tick families

require a sterilization dose of less than 200 Gy

[6.24], although the sterilization dose for some

insects and related arthropods is much lower and

ranges widely among and within orders. As was

found for plants [6.8], radiosensitivity of insects is

related to the average interphase nuclear volume

[6.24].

6.5. RADIATION EFFECTS ON FARM

ANIMALS

Ruminants, both domesticated (cattle, goats,

sheep) and wild (elk, deer), generally receive high

doses in radioactively contaminated environments

because they consume large amounts of vegetation,

and many radionuclides accumulate in their bodies;

for example, each day a single cow consumes about

30% of the grass from an area of 150 m

. Ingestion

of radionuclides leads to exposure of the gut, the

thyroid and other body organs. Injuries to cattle are

a major fallout consequence for rural populations,

because of livestock loss but also because of the

associated social and psychological implications

[6.25, 6.26].

In the period shortly after the accident,

domestic livestock within the CEZ were exposed to

high levels of radioactive iodine (

131

I and

133

I, with

half-lives of 8 d and 21 h, respectively); this resulted

in significant internal and external doses from beta

and gamma radiation (Table 6.4). A thyroid dose of

76 Gy from the two isotopes of iodine is sufficient to

cause serious damage to the gland [6.27]. The soils

of Ukraine and Belarus are naturally low in stable

iodine, cobalt and manganese. In conditions of

endemic deficiency of stable iodine, the transfer of

radioactive iodine from blood to the thyroid gland

may be two to three times higher than normal

[6.15]. These conditions accentuated the conse-

quences of the accident.

Depressed thyroid function in cattle was

related to the dose received (69% reduction in

function with a thyroid dose of 50 Gy, and 82%

reduction in animals that received a dose of

280 Gy). The concentration of thyroid hormones in

the blood of animals was lower than the physio-

logical norm during the whole lactation period.

Radiation damage to the thyroid gland was

confirmed by histological studies (i.e. hyperplasia of

connective tissue and sometimes adipose tissue,

vascular hyperaemia and necrosis of epithelium).

Animals with practically no thyroid tissue were

observed in Ukraine. Disruptions of the hormonal

status in calves born to cows with irradiated thyroid

glands were especially pronounced [6.28]. Similar

effects were observed in cattle evacuated from the

Belarusian portion of the CEZ [6.26].

Although most livestock were evacuated from

the area after the accident, several hundred cattle

were maintained in the more contaminated areas

for a two to four month period. By the autumn of

1986, some of these animals had died; others

showed impaired immune responses, lowered body

temperatures and cardiovascular disorders.

Hypothyroidism lasted until 1989, and may have

been responsible for reproductive failures in

animals that had received a thyroid dose of more than

180 Gy [6.26]. The offspring of highly exposed cows

had reduced weight, reduced daily weight gains and

signs of dwarfism. Reproduction returned to normal in

the spring of 1989. Haematological parameters were

normal for animals kept in areas with

137

Cs contami-

nation of 0.2–1.4 MBq/m

(5–40 Ci/km

) [6.28].

TABLE 6.4. DOSES TO CATTLE THAT STAYED IN THE 30 km ZONE OF CHERNOBYL FROM 26

APRIL TO 3 MAY 1986 [6.21]

Distance from the

Chernobyl nuclear

power plant (km)

Surface activity

(10

Bq/m

)

Absorbed dose (Gy)

Thyroid Gastrointestinal tract Whole body internal

3 8.4 300 2.5 1.4

10 6.1 230 1.8 1.0

14 3.5 260 1.0 0.6

12 2.4 180 0.7 0.4

35 1.2 90 0.4 0.2

132

Chronic radiation damage was observed in

over 2000 sheep and 300 horses (3–8 years old)

removed from the highly contaminated Khoiniki

area of Belarus 1.5 years after the accident [6.26].

Doses were not estimated. In sheep, a depression of

general condition, emaciation, heavy breathing,

decrease of temperature and other abnormalities

were found. Leukopaenia, erythropaenia, thrombo-

cytopaenia and eosinophilia, increase in blood sugar

concentrations 1.5–2 times higher than normal, and

a significant decrease of thyroid hormone concen-

trations compared with normal levels were

observed. The offspring weight and fleece clip yield

of irradiated sheep were half as much as, or less

than, those of healthy individuals. In horses, the

damage resulted in depression of general condition,

oedema, leukopaenia, thrombocytopaenia,

eosinophilia and myelocytosis. Seventy per cent of

the animals had thyroid hormone concentrations in

blood serum that were lower than the detection

levels of the assay methods [6.26].

Numerous news reports of radiation induced

teratogenesis (birth defects) in cattle and pigs

occurred in regions where total doses did not

exceed 0.05 Gy/a. Scientific evidence indicates that

increased birth defects are not distinguishable from

background frequencies at such low doses [6.25].

Additionally, data for 1989 show that livestock birth

defects in the contaminated area of the Zhytomyr

region were no higher than in the uncontaminated

areas of the same region. Photographs of a six

footed calf were widely disseminated and the

abnormality was attributed to the accident. The calf,

however, was born in June 1986, and thus the

process of differentiation and organ formation

within the womb finished prior to the accident; this

much publicized observation of teratogenesis was

therefore caused by factors other than radiation

from the Chernobyl accident.

6.6. RADIATION EFFECTS ON OTHER

TERRESTRIAL ANIMALS

Four months after the accident, surveys and

autopsies of wildlife and abandoned domestic

animals that remained within the 10 km radius of

the Chernobyl nuclear power plant were conducted

[6.14]. Fifty species of birds were identified,

including some rare species; all appeared normal in

appearance and behaviour. No dead birds were

found. Swallows and house sparrows were found to

be producing progeny that also appeared normal.

Forty-five species of mammals from six orders were

observed, and no unusual appearances or

behaviours were noted.

Some wildlife and domestic animals were shot

and autopsied in August and September 1986. Dogs

and chickens showed signs of chronic radiation

syndrome (reduced body mass, reduced fat reserves,

increase in mass of lymph nodes, liver and spleen,

haematomas present in liver and spleen and

thickening of the lining of the lower intestine). No

eggs were found in the nests of chickens or in their

ovaries.

During the autumn of 1986, the number of

small rodents on highly contaminated research plots

decreased by a factor of two to ten. Estimates of

absorbed dose during the first five months after the

accident ranged from 12 to 110 Gy for gamma and

580 to 4500 Gy for beta radiation. The numbers of

animals were recovering by the spring of 1987,

mainly due to immigration from less affected areas.

In 1986 and 1987 the percentage of pre-implan-

tation deaths in rodents from the highly contami-

nated areas increased twofold to threefold

compared with controls. Resorption of embryos

also increased markedly in rodents from the

affected areas; however, the number of progeny per

female did not differ from controls [6.29].

6.7. RADIATION EFFECTS ON AQUATIC

ORGANISMS

Cooling water for the Chernobyl nuclear

power plant was obtained from the 21.7 km

human-

made reservoir located south-east of the plant site.

The cooling reservoir became heavily contaminated

following the accident (see Section 3.5 for details)

with over 6.5 ± 2.7 × 10

Bq of a mixture of radio-

nuclides in the water and sediments [6.30]. Aquatic

organisms were exposed to external exposure from

radionuclides in water and contaminated bottom

sediments and irradiation from contaminated

aquatic plants. Internal exposure occurred as

organisms took up radioactively contaminated food

and water or inadvertently consumed contaminated

sediments. The resultant doses to aquatic biota over

the first 60 days following the accident are depicted

in Fig. 6.5.

The maximum dose rates for aquatic

organisms (excluding fish) were reported in the first

two weeks after the accident, when short lived

isotopes (primarily

131

I) contributed 60–80% of the

dose. During the second week, the contribution of

133

short lived radionuclides to doses to aquatic

organisms decreased by a factor of two. Maximum

dose rates to fish were delayed (Fig. 6.5), due to the

time required for their food webs to become

contaminated with longer lived radionuclides

(mostly

134,137

Cs,

144

Ce/

144

Pr,

106

Ru/

106

Rh and

Sr/

Y). Differences in dose rates among fish species

occurred due to their trophic positions. Non-

predatory fish (carp, goldfish, bleak) reached

estimated peak dose rates of 3 mGy/d from internal

contamination in 1986, followed by significant

reductions in 1987. Dose rates in predatory fish

(perch), however, increased in 1987 and did not

start to decline until 1988 [6.21]. Accumulated doses

were highest for the first generation of fish born in

1986 and 1987. Bottom dwelling fish (goldfish, silver

bream, bream, carp) that received significant

exposure from the bottom sediments received

accumulated total doses of approximately 10 Gy.

In 1990 the reproductive capacity of young

silver carp was analysed [6.31]. The fish were in live

boxes within the cooling pond at the time of the

accident. By 1988 the fish reached sexual maturity.

Over the entire post-accident period they received a

dose of 7–8 Gy. Biochemical analyses of muscles,

liver and gonads indicated no difference from the

controls. The amount of fertilized spawn was 94%;

11% of the developing spawn were abnormal.

Female fertility was 40% higher than the controls,

but 8% of the irradiated sires were sterile. The level

of fluctuating asymmetry in offspring did not differ

from the controls, although the level of cytogenetic

damage (22.7%) significantly exceeded the controls

(5–7%). In contrast, Pechkurenkov [6.32] reported

that the number of cells with chromosome

aberrations in 1986–1987 in carp, flat bream and

silver carp was within the norm. It is worth noting

that the cooling pond was subject not only to

radioactive contamination but also to chemical

pollution.

Recent reviews of the chronic effects of

ionizing radiation on reproduction in fish, with the

Chernobyl data included (Table 6.5), have been

summarized.

6.8. GENETIC EFFECTS IN ANIMALS AND

PLANTS

Quality data concerning the incidence of

Chernobyl related induced mutations in plants and

animals are relatively sparse. An increased

mutation level was apparent in 1987 in the form of

various morphological abnormalities observed in

Canada flea-bane, common yarrow and mouse

millet plants. Examples of abnormalities include

unusual branching of stems, doubling of the number

of racemes, abnormal colour and size of leaves and

flowers, and development of ‘witches’ brooms’ in

pine trees. Similar effects in the 5 km radius circle

near the reactor also appeared in deciduous trees

(leaf gigantism, changes in leaf shape; see Fig. 6.6).

0.1

100

0 10 20 30 40 50 60

Days post-accident

Phytoplankton

Zooplankton

Fish

Macroalgae

Benthic

organisms

cGy/d

FIG. 6.5. The dynamics of absorbed dose rate (cGy/d) of

organisms within the Chernobyl cooling pond during the

irst 60 days following the accident. Data are model results

based on concentrations of radionuclides in the water

column and lake sediments [6.21].

FIG. 6.6. Typical morphological abnormality seen on

conifer trees. Such enhancement of vegetative growth and

igantism of some plant parts were not uncommon (photo-

raph courtesy of T. Hinton, 1991).

134

Morphological changes were observed at an initial

gamma close rate of 0.2–0.3 mGy/h. At 0.7–

1.3 mGy/h enhancement of vegetative reproduction

(heather) and gigantism of some plant species were

observed [6.19, 6.20, 6.34, 6.35].

Cytogenetic analysis of cells from the root

meristem of winter rye and wheat germ of the 1986

harvest demonstrated a dose dependence in the

number of aberrant cells. A significant excess over

the control level of aberrations was observed at an

absorbed dose of 3.1 Gy, inhibition of mitotic

activity occurred at 1.3 Gy and germination was

reduced at 12 Gy [6.36]. Analysis of three successive

generations of winter rye and wheat on the most

contaminated plots revealed that the rate of

aberrant cells in the intercalary meristem in the

second and third generations was higher than in the

first.

From 1986 to 1992 mutation dynamics were

studied in populations of Arabidopsis thaliana

Heynh. (L.) within the CEZ [6.37]. On all study

plots in the first two to three years after the

accident, Arabidopsis populations exhibited an

increased mutation burden. In later years, the level

of lethal mutations declined; nevertheless the

mutation rate in 1992 was still four to eight times

higher than the spontaneous level. The dose

dependence of the mutation rate was best approxi-

mated by a power function with a power index of

less than one.

Zainullin et al. [6.38] observed elevated levels

of sex linked recessive lethal mutations in natural

Drosophila melanogaster populations living under

conditions of increased background radiation due to

the Chernobyl accident. Mutation levels were

increased in 1986–1987 in flies inhabiting contami-

nated areas with initial exposure rates of 2 mGy/h

and more. In the subsequent two years mutation

frequencies gradually returned to normal.

Studies of adverse genetic effects in wild mice

have been reported by Shevchenko et al. [6.39] and

Pomerantseva et al. [6.40]. These involved mice

caught during 1986–1991 within a 30 km radius of

the Chernobyl reactor with different levels of

gamma radiation and in 1992–1993 at a site in the

Bryansk region of the Russian Federation. The

TABLE 6.5. CHRONIC EFFECTS OF IONIZING RADIATION ON REPRODUCTION IN FISH,

DERIVED FROM THE FASSET DATABASE [6.33]

Dose rate

(mGy/h)

Dose rate

(mGy/d)

Effects

0–99 0–2.4 Background dose group; normal cell types, normal damage and normal mortality observed

100–199 2.4–4.8 No data available

200–499 4.8–12 Reduced spermatogonia and sperm in tissues

500–999 12–24 Delayed spawning, reduction in testes mass

1000–1999 24–48 Mean lifetime fecundity decreased, early onset of infertility

2000–4999 48–120 Reduced number of viable offspring

Increased number of embryos with abnormalities

Increased number of smolts in which sex was undifferentiated

Increased brood size reported

Increased mortality of embryos

5000–9999 120–240 Reduction in number of fish surviving to one month of age

Increased vertebral abnormalities

>10 000 >240 Interbrood time tending to decrease with increasing dose rate

Significant reduction in neonatal survival

Sterility in adult fish

Destruction of germ cells within 50 days in medaka fish

High mortality of fry; germ cells not evident

Significant decrease in number of male salmon returning to spawn

After four years, female salmon had significantly reduced fecundity

135

estimated total doses of gamma and beta radiation

varied widely and reached 3–4 Gy per month in

1986–1987. One endpoint was dominant lethality,

measured by embryo mortality of the offspring of

wild male mice mated to unexposed female

laboratory mice. The dominant lethality rate was

elevated for a period of a few weeks following

capture for mice sampled at the most contaminated

site. At dose rates of about 2 mGy/h, two of 122

captured males produced no offspring and were

assumed to be sterile. The remainder showed a

period of temporary infertility and reduced testes

mass, which, however, recovered with time after

capture.

The frequencies of reciprocal translocations in

mouse spermatocytes were consistent with previous

studies. For all collected mice, a dose rate

dependent increased incidence of reciprocal trans-

locations (scored in spermatocytes at meiotic

metaphase I) was observed. The frequency of mice

harbouring recessive lethal mutations decreased

with time post-accident [6.40]. Radiation related

gene mutation is unlikely to have any adverse effect

on populations at the dose rates that prevail now.

Advances in the sophistication and associated

technologies of detecting molecular and

chromosomal damage have occurred since the early

genetic studies prior to the Chernobyl accident.

Such advances have allowed researchers on the

genetic consequences of the Chernobyl accident to

examine endpoints not previously considered. Most

prominent, and controversial, is the mutation

frequencies in repeat DNA sequences termed

‘minisatellite loci’ or expanded simple tandem

repeats (ESTRs). These are repeat DNA sequences

that are distributed throughout the germline and

that have a high background (spontaneous)

mutation rate. At present, ESTRs are considered to

have no function, although this is a matter of much

interest and discussion [6.41, 6.42]. Minisatellite

mutations have only rarely been associated with

recognizable genetic disease [6.43].

Although laboratory examination of

mutations in mouse ESTR loci shows clear evidence

of a mutational dose response [6.44, 6.45], no

convincing data on elevated levels of minisatellite

mutations in plants or animals residing in the

Chernobyl affected areas appear to have been

published so far in the peer reviewed scientific

literature. In general, quantitative interpretation of

the ESTR data is difficult because of conflicting

findings, their weak association with genetic disease,

dosimetric uncertainties and methodological

problems [6.42]. This is an area of science that

requires additional research.

6.9. SECONDARY IMPACTS AND CURRENT

CONDITIONS

Prior to the accident, much of the area around

Chernobyl was covered in 30–40 year old pine

stands that, from a successional standpoint,

represented mature, stable ecosystems. The high

dose rates from ionizing radiation during the first

few weeks following the accident altered the

balanced community by killing sensitive individuals,

altering reproduction rates, destroying some

resources (e.g. pine stands), making other resources

more available (e.g. soil water) and opening niches

for immigration of new individuals. All these

components, and many more, were interwoven in a

complex web of action and reaction that altered

populations and communities of organisms.

Exposure to ionizing radiation is an environ-

mental stress, in many ways similar to other

environmental stresses such as pollution by metals

or the destruction caused by forest fires. If such

stressors are sufficient, the community organization

is changed and generally reverts to an earlier

successional state. However, when the stress is

subsequently reduced and sufficient time passes,

recovery occurs and the ecosystem again regains

stability, advancing towards a more mature state.

The change in species diversity observed within the

soil invertebrate communities presented above is

perhaps the most obvious published example of

community level change and subsequent recovery

following the Chernobyl accident. The death of pine

stands close to the Chernobyl reactor and the

subsequent establishment of grasslands and

deciduous trees are striking visual examples.

Age and sex distributions, diversity and the

abundance and gross physiological conditions of

small mammal populations in the CEZ appear to be

similar to background locations in other parts of

Ukra

ine

[6.46–6.48]. Reports on the current genetic

conditions of rodents within the zone are contra-

dictory; for example, Shevchenko et al. [6.39] found

significant disorders in spermatogenesis, while

Baker et al. [6.46] found no reproductive inhibition

or germ line mutations.

Layered on top of the impacts of the radiation

exposure was the abrupt and drastic change that

occurred when humans were removed from the

CEZ. The town adjacent to the Chernobyl reactor,

136

Pripyat, was abandoned when over 50 000 people

were evacuated. Agricultural activity, forestry,

hunting and fishing within the CEZ were stopped

because of the radioactive contamination of the

products. Only activities designed to mitigate the

consequences of the accident were carried out, as

well as those supporting the living conditions of the

cleanup workers, including substantial road

construction.

For some years after the accident, the agricul-

tural fields still yielded domesticated produce, and

many animal species, especially rodents and wild

boars, consumed the abandoned cereal crops,

potatoes and grasses as an additional source of

forage. This advantage, along with the special

reserve regulations established in the CEZ (e.g. a

ban on hunting), tended to compensate for the

adverse biological effects of the radiation and

promoted an increase in the populations of wild

animals. Significant population increases of game

mammals (wild boar, roe deer, red deer, elk, wolves,

foxes, hares, beavers, etc. (Fig. 6.7)) and bird species

(black grouse, ducks, etc.) were observed soon after

the Chernobyl accident [6.49, 6.50].

More than 400 species of vertebrate animals,

including 67 ichthyoids, 11 amphibians, 7 reptiles,

251 birds and 73 mammals, inhabit the territory of

the evacuated town of Pripyat and its vicinity; more

than 50 of them belong to a list of those protected

according to national Ukrainian and European Red

Books. The CEZ has become a breeding area for

white tailed eagles, spotted eagles, eagle owls,

cranes and black storks (Fig. 6.8) [6.51].

In the Pripyat River floodplain a developed

system of artificial drainage channels now supports

about a hundred families of beavers. Recognizing

the value of the abandoned land around Chernobyl,

28 endangered Przhevalsky wild horses were

introduced in 1998. After six years their number had

doubled [6.51]. In both the Ukrainian and

Belarusian parts of the CEZ, State radioecological

reserves have been created with a regime of nature

protection.

As has been shown many times before, when

humans are removed, nature flourishes. This

phenomenon exists in US National Parks such as

Yellowstone and the Grand Tetons, and at large US

Department of Energy sites where the general

public has been excluded for over 50 years. Human

presence in any environment is a disturbance to the

natural biota. Normal activities of farming, hunting,

logging and road building, to name but a few,

fragment, pollute and generally stress the processes

and mechanisms of natural environments. The

removal of humans alleviates one of the more

persistent and ever growing stresses experienced by

natural ecosystems.

(

FIG. 6.7. Wild boar (a) and wolves (b) inhabiting the CE

are not afraid of people because of long term hunt prohibi-

tion (photographs courtesy of S. Gaschak, 2004).

FIG. 6.8. A white tailed eagle chick observed recently in the

CEZ. Before 1986 these rare predatory birds had rarely

been found in this area (photograph courtesy of S.

Gaschak, 2004).

137

On the other hand, the absence of forest

management, and the associated increase in forest

fires, have substantial impacts on natural commu-

nities. After the human population was evacuated,

both wood cutting and the construction of

mineralized fire prevention strips ceased. The

number of dead trees increased and created

conditions that enhanced the development of forest

diseases and pests (borers, bark beetles, etc.). The

amount of dead wood and brushwood has gradually

increased in the unmanaged forests. The

degradation of the forests resulted in enormous

forest fires during the dry summer period of 1992,

when the area of burnt forest amounted to 170 km

(i.e. about one sixth of the woodlands) [6.52].

Without a permanent residence of humans for

20 years, the ecosystems around the Chernobyl site

are now flourishing. The CEZ has become a wildlife

sanctuary [6.47], and it looks like the nature park it

has become.

6.10. CONCLUSIONS AND

RECOMMENDATIONS

6.10.1. Conclusions

(a) Radiation from radionuclides released by the

Chernobyl accident caused numerous acute

adverse effects in the biota located in the areas

of highest exposure (i.e. up to a distance of a

few tens of kilometres from the release point).

Beyond the CEZ, no acute radiation induced

effects on biota have been reported.

(b) The environmental response to the Chernobyl

accident was a complex interaction among

radiation dose, dose rate and its temporal and

spatial variations, and the radiosensitivities of

the different taxons. Both individual and

population effects caused by radiation

induced cell death have been observed in

plants and animals as follows:

(i) Increased mortality of coniferous plants,

soil invertebrates and mammals;

(ii) Reproductive losses in plants and

animals;

(iii) Chronic radiation syndrome in animals

(mammals, birds, etc.).

No adverse radiation induced effects have

been reported in plants and animals exposed

to a cumulative dose of less than 0.3 Gy during

the first month after the radionuclide fallout.

levels due to radionuclide decay and

migration, populations have been recovering

from the acute radiation effects. By the next

growing season after the accident, the

population viability of plants and animals

substantially recovered as a result of the

combined effects of reproduction and

immigration. A few years were needed for

recovery from the major radiation induced

adverse effects in plants and animals.

(d) The acute radiobiological effects observed in

the Chernobyl accident area are consistent

with radiobiological data obtained in experi-

mental studies or observed in natural

conditions in other areas affected by ionizing

radiation. Thus rapidly developing cell

systems, such as meristems of plants and insect

larvae, were predominantly affected by

radiation. At the organism level, young plants

and animals were found to be the most

sensitive to the acute effects of radiation.

(e) Genetic effects of radiation, in both somatic

and germ cells, were observed in plants and

animals in the CEZ during the first few years

after the accident. Both in the CEZ and

beyond, different cytogenetic anomalies

attributable to radiation continue to be

reported from experimental studies

performed on plants and animals. Whether the

observed cytogenetic anomalies have any

detrimental biological significance is not

known.

(f) The recovery of affected biota in the CEZ has

been confounded by the overriding response

to the removal of human activities (e.g.

termination of agricultural and industrial

activities and the accompanying environ-

mental pollution in the most affected area).

As a result, the populations of many plants

and animals have expanded, and the present

environmental conditions have had a positive

impact on the biota in the CEZ.

6.10.2. Recommendations for future research

(a) In order to develop a system of environmental

protection against radiation, the long term

impact of radiation on plant and animal

populations should be further investigated in

the CEZ; this is a globally unique area for

138

radioecological and radiobiological research

in an otherwise natural setting.

(b) In particular, multigenerational studies of

radiation effects on the genetic structure of

plant and animal populations might bring

fundamentally new scientific information.

methods for biota–dose reconstruction, for

example in the form of a unified dosimetric

protocol.

6.10.3. Recommendations for countermeasures

and remediation

(a) Protective actions for farm animals in the event

of a nuclear or radiological emergency should

be developed and internationally harmonized

based on modern radiobiological data,

including the experience gained in the CEZ.

(b) It is likely that any technology based

remediation actions aimed at improving the

radiological conditions for plants and animals

in the CEZ would have adverse impacts on

biota.

REFERENCES TO SECTION 6

[6.1] INTERNATIONAL ATOMIC ENERGY

AGENCY, Effects of Ionizing Radiation on

Aquatic Organisms and Ecosystems, Technical

Reports Series No. 172, IAEA, Vienna (1976).

[6.2] INTERNATIONAL ATOMIC ENERGY

AGENCY, Assessing the Impact of Deep Sea

Disposal of Low Level Radioactive Waste on

Living Marine Resources, Technical Reports

Series No. 288, IAEA, Vienna (1988).

[6.3] INTERNATIONAL ATOMIC ENERGY

AGENCY, Effects of Ionizing Radiation on Plants

and Animals at Levels Implied by Current

Radiation Protection Standards, Technical

Reports Series No. 332, IAEA, Vienna (1992).

[6.4] UNITED NATIONS, Sources, Effects and Risks

of Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(1996).

[6.5] WHICKER, F.W., SCHULTZ, V., Radioecology:

Nuclear Energy and the Environment, CRC Press,

Boca Raton, FL (1982).

[6.6] WHICKER, F.W., FRALEY, L., Effects of

ionizing radiation on terrestrial plant communi-

ties, Adv. Radiat. Biol. 4 (1974) 317–366.

[6.7] WHICKER, F.W., “Impacts on plant and animal

populations”, Health Impacts of Large Releases of

Radionuclides, CIBA Foundation Symposium

No. 203, Wiley, New York (1997) 74–93.

[6.8] SPARROW, A.H., Correlation of nuclear volume

and DNA content with higher plant tolerance to

chronic radiation, Science 134 (1961) 282–284.

[6.9] WOODHEAD, D., ZINGER, I., Radiation

Effects on Plants and Animals, Deliverable 4 to

the Project FASSET, Framework for the Assess-

ment of Environmental Impact, Contract No.

FIGE-CT-2000-00102, Swedish Radiation Protec-

tion Authority, Stockholm (2003).

[6.10] SAZYKINA, T.G., JAWORSKA, A., BROWN,

J.E. (Eds), Dose–Effect Relationships for

Reference (or Related) Arctic Biota, Deliverable

Report 5 for the EPIC Project, Contract No.

ICA2-CT-200-10032, Norwegian Radiation

Protection Authority, Østerås (2003).

[6.11] REAL, A., SUNDELL-BERGMAN, S.,

KNOWLES, J.F., WOODHEAD, D.S., ZINGER,

I., Effects of ionising radiation exposure on plants,

fish and mammals: Relevant data for environ-

mental radiation protection, J. Radiol. Prot. 24

(2004) A123–A137.

[6.12] UNITED NATIONS, Sources and Effects of

Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(2000).

[6.13] PRISTER, B.S., SHEVCHENKO, V.A.,

KALCHENKO, V.A., “Genetic effects of radio-

nuclides on agricultural crops”, Progress of

Modern Genetics, USSR Academy of Science,

Moscow (1982) 138–148 (in Russian).

[6.14] KRIVOLUTSKY, D., MARTUSHOV, V.,

RYABTSEV, I., “Influence of radioactive

contamination on fauna in the area of the

Chernobyl NPP during first years after the

accident (1986–1988)”, Bioindicators of Radioac-

tive Contamination, Nauka, Moscow (1999) 106–

122 (in Russian).

[6.15] PRISTER, B.S., LOSCHILOV, N.A., NEMETS,

O.F., POYARKOV, V.A., Fundamentals of Agri-

cultural Radiology, 2nd Edn, Urozhay, Kiev (1991)

(in Russian).

[6.16] GRODZINSKY, D.M., KOLOMIETS, O.D.,

KUTLAKHMEDOV, Y.A., Anthropogenic radio-

nuclide anomaly and plants, Lybid, Kiev (1991) (in

Russian).

[6.17] SHEVCHENKO, V.A., et al., Genetic conse-

quences of radioactive contamination of the envi-

ronment after the Chernobyl accident for

populations of plants, Radiat. Biol. Radioecol. 36

(1996) 531–545.

[6.18] GRODZINSKY, D.N., GUDKOV, I.N., “Radibio-

logical effects on plants in the contaminated terri-

tory”, The Chornobyl Exclusion Zone, Nat.

Academy of Science of Ukraine, Kiev (2001) 325–

377.

[6.19] TIKHOMIROV, F.A., SHCHEGLOV, A.I., Main

investigation results on the forest radioecology in

139

the Kyshtym and Chernobyl accidents zones, Sci.

Total Environ. 157 (1994) 45–47.

[6.20] ARKHIPOV, N.P., KUCHMA, N.D.,

ASKBRANT, S., PASTERNAK, P.S., MUSICA,

V.V., Acute and long-term effects of irradiation on

pine (Pinus silvestris) stands post-Chernobyl, Sci.

Total Environ. 157 (1994) 383–386.

[6.21] KRYSHEV, I., et al., Radioecological Conse-

quences of the Chernobyl Accident, Nuclear

Society, Moscow (1992).

[6.22] KOZUBOV, G., et al., Radiation exposure of the

coniferous forest in the area exposed to the

Chernobyl contamination, Komy Scientific Centre

of the Academy of Sciences, Syktyvkar (1990) (in

Russian).

[6.23] KRIVOLUTSKY, D., POKARZHEVSKY, A.,

Effect of radioactive fallout on soil animal popula-

tions in the 30 km zone of the Chernobyl atomic

power station, Sci. Total Environ. 112 (1992) 69–

77.

[6.24] BAKRI, A., HEATHER, N., HENDRICHS, J.,

FERRIS, I., Fifty years of radiation biology in

entomology: Lessons learned from IDIDAS, Ann.

Entomol. Soc. Am. 98 (2005) 1–12.

[6.25] PRISTER, B.S., Consequences of the Accident at

the Chernobyl NPP for Agriculture in Ukraine,

Center of Privatization and Economic Reform,

Kiev (1999) (in Russian).

[6.26] ILYAZOV, R.G., et al., Ecological and Radiobio-

logical Consequences of the Chernobyl Catas-

trophe for Stock-breeding and Ways of its

Overcoming (ILYAZOV, R.G., Ed.), Fan, Kazan

(2002) (in Russian).

[6.27] BELOV, A.D., KIRSHIN, V.A., Veterinary Radi-

obiology, Atomizdat, Moscow (1987) (in Russian).

[6.28] ASTASHEVA, N.P., et al., “Influence of radiation

released during the Chernobyl NPP accident on

clinical and physiological status of agricultural

animals”, Problems of Agricultural Radiology,

UIAR, Kiev (1991) 176–180 (in Russian).

[6.29] TASKAEV, A., TESTOV, B., “Number and repro-

duction of mouse-like rodents in the Chernobyl

accident area”, Bioindicators of Radioactive

Contamination, Nauka, Moscow (1999) 200–205

(in Russian).

[6.30] KRYSHEV, I., Radioactive contamination of

aquatic ecosystems following the Chernobyl

accident, J. Environ. Radioact. 227 (1995) 207–

219.

[6.31] RYABOV, I.N., Effect of radioactive contamina-

tion on hydrobionts within the thirty-kilometer

zone of the Chernobyl NPP, Radiobiologiya 32

(1992) 662–667 (in Russian).

[6.32] PECHKURENKOV, V.L., The influence of the

Chernobyl disaster on fish populations in a cooling

pond, Radiobiologiya 31 (1991) 704–708 (in

Russian).

[6.33] COPPLESTONE, D., ZINGER, I., JACKSON,

D., “The challenge of protecting non-human biota

from exposure to ionising radiation”, paper

presented at the Society for Radiological Protec-

tion 40th Anniversary Meeting: ALARP: Princi-

ples and Practices, St. Catherine’s College,

Oxford, 2003.

[6.34] OZUBOV, G.M., TASKAEV, A.I., Radiobio-

logical and Radioecological Investigations of

Woody Plants, Nauka, St. Petersburg (1994) (in

Russian).

[6.35] TIKHOMIROV, F.A., SHCHEGLOV, A.I.,

SIDOROV, V.P., Forests and forestry: Radiation

protection measures with special reference to the

Chernobyl accident zone, Sci. Total Environ. 137

(1993) 289–305.

[6.36]

GERASKIN, S.A., et al., Genetic consequences of

radioactive contamination by the Chernobyl

fallout to agricultural crops, J. Environ. Radioact.

66 (2003) 155–169.

[6.37] ABRAMOV, V.I., FEDORENKO, O.M.,

SHEVCHENKO, V.A., Genetic consequences of

radioactive contamination for populations of

Arabidopsis, Sci. Total Environ. 112 (1992) 19–28.

[6.38] ZAINULLIN, V.G., SHEVCHENKO, V.A.,

MYASNYANKINA, E.N., GENERALOVA,

M.V., RAKIN, A.O., The mutation frequency of

Drosophila melanogaster populations living under

conditions of increased background radiation due

to the Chernobyl accident, Sci. Total Environ. 112

(1992) 37–44.

[6.39] SHEVCHENKO, V.A., et al., Genetic disorders in

mice exposed to radiation in the vicinity of the

Chernobyl nuclear power station, Sci. Total

Environ. 112 (1992) 45–56.

[6.40] POMERANTSEVA, M.D., RAMAIYA, L.K.,

CHEKHOVICH, A.V., Genetic disorders in

house mouse germ cells after the Chernobyl catas-

trophe, Mutat. Res. 381 (1997) 97–103.

[6.41] INTERNATIONAL COMMISSION ON RADI-

OLOGICAL PROTECTION, Biological Effects

After Prenatal Irradiation (Embryo and Fetus),

Publication 90, Pergamon Press, Oxford and New

York (2004).

[6.42] COMMITTEE EXAMINING RADIATION

RISKS FROM INTERNAL EMITTERS, Report

of the Committee Examining Radiation Risks of

Internal Emitters, CERRIE, London (2004),

www.cerrie.org

[6.43] BRIDGES, B.A., Radiation and germline

mutations at repeat sequences: Are we in the

middle of a paradigm shift? Radiat. Res. 156

(2001) 631–641.

[6.44] FAN, Y.J., et al., Dose-response of a radiation

induction of a germline mutation at a hyper-

variable mouse minisatellite locus, Int. J. Radiat.

Biol. 68 (1995) 177–183.

[6.45] DUBROVA, Y.E., et al., State specificity, dose

response, and doubling dose for mouse minisatel-

lite germ-line mutation induced by acute radi-

ation, Proc. Natl. Acad. Sci. USA 95 (1998) 6251–

6265.

140

[6.46] BAKER, R.J., et al., Small mammals from the

most radioactive sites near the Chornobyl nuclear

power plant, J. Mammol. 77 (1996) 155–170.

[6.47] BAKER, R.J., CHESSER, R.K., The Chernobyl

nuclear disaster and subsequent creation of a

wildlife preserve, Environ. Toxicol. Chem. 19

(2000) 1231–1232.

[6.48] JACKSON, D., COPPLESTONE, D., STONE,

D.M., Effects of chronic radiation exposure on

small mammals in the Chernobyl exclusion zone,

Nucl. Energy 43 (2004) 281–287.

[6.49] GAICHENKO, V.A., KRYZANOVSKY, V.I.,

STOLBCHATY, V.N., Post-accident state of the

Chernobyl nuclear power plant alienated zone

faunal complexes, Radiat. Biol. Ecol., Special issue

(1994) 27–32 (in Russian).

[6.50] SUSCHENYA, L.M., et al. (Eds), The Animal

Kingdom in the Accident Zone of the Chernobyl

NPP, Navuka i Technica, Minsk (1995) (in

Russian).

[6.51] GASCHAK, S.P., et al., Fauna of Vertebrates in

the Chernobyl Zone of Ukraine, International

Chornobyl Center for Nuclear Safety, Radioactive

Waste and Radioecology, Slavutych (2002) (in

Ukrainian).

[6.52] FRANTSEVICH, L.I., et al., “Secondary ecolog-

ical changes associated with resettlement of

population”, The Chernobyl Disaster (BARJA-

KHTAR, V.G., Ed.), Naukova Dumka, Kiev

(1995) 335–340 (in Ukrainian).

141

7. ENVIRONMENTAL AND RADIOACTIVE WASTE

MANAGEMENT ASPECTS OF THE DISMANTLING

OF THE CHERNOBYL SHELTER

The destruction of the unit 4 reactor at the

Chernobyl nuclear power plant created radioactive

contamination and radioactive waste

in the unit,

the Chernobyl nuclear power plant site and the

surrounding area (further referred to as the CEZ).

The future development of the CEZ depends on the

strategy for the conversion of unit 4 into an ecologi-

cally safe system (i.e. the development of the NSC,

the dismantlement of the current shelter, removal of

FCM, and eventual decommissioning of the reactor

site).

In particular, the long term strategy for unit 4

involves implementation of the NSC concept to

cover the unstable shelter and the related

radioactive waste management activities at the

Chernobyl nuclear power plant site and the CEZ.

Currently units 1, 2 and 3 (1000 MW RBMK

reactors) are shut down awaiting decommissioning;

two additional reactors (units 5 and 6) that had been

near completion were abandoned in 1986 following

the accident.

This section addresses the current status of

unit 4 and the existing and future environmental

impact associated with it and the management of

the radioactive waste from the accident at the

Chernobyl nuclear power plant site and in the

CEZ.

7.1. CURRENT STATUS AND THE FUTURE

OF UNIT 4 AND THE SHELTER

7.1.1. Unit 4 of the Chernobyl nuclear power

plant after the accident

In the course of the 1986 accident, a small part

of the nuclear fuel (3.5% according to past

estimates [7.1] or 1.5% according to recent

estimates [7.2]) and a substantial fraction of volatile

radionuclides (see Section 3.1) were released from

the damaged unit 4. The remainder of the damaged

nuclear fuel, more than 95% of the fuel mass at the

moment of the accident, i.e. about 180 t, was left in

the remains of the reactor [7.1]. The uncertainty in

this estimate is discussed in Section 7.1.5.

The first measures taken after the accident to

control the fire and the radionuclide releases

consisted of dumping neutron absorbing

compounds and fire control material into the crater

formed by the destruction of the reactor [7.1] (see

Fig. 7.1). The total amount of material dumped on

the reactor was approximately 5000 t, including

about 40 t of boron compounds, 2400 t of lead,

1800 t of sand and clay and 600 t of dolomite, as well

as sodium phosphate and polymer liquids [7.1].

At the Chernobyl nuclear power plant site in

mid-May 1986 there were high levels of air radiation

dose rate and air activity concentration, caused by

relatively uniform contamination of the area with

finely dispersed nuclear fuel and aerosols of short

lived radionuclides, and also the presence of

dispersed nuclear fuel particles or fragments. These

fragments consisted of discrete and non-uniform

material from the reactor core, reactor construc-

tional material and graphite.

After the accident, the debris of the destroyed

reactor building was collected, along with fragments

of the reactor core, etc., and the soil surface layer.

Thousands of cubic metres of radioactive waste

generated by this work were disposed of in the

pioneer wall and the cascade wall. Construction of

walls around the damaged reactor reduced the

radiation dose rates by a factor of 10–20 [7.3]. The

completion of the pioneer wall and the cascade wall

and a significant reduction in radiation levels

allowed the shelter to be constructed.

The shelter, which was intended to provide the

environmental containment of the damaged

reactor, was erected within an extremely short

period of time, between May and November 1986,

under conditions of high radiation exposure of the

personnel. The steps taken to save time and cost

during the construction, and the high dose rates

inside the structure, resulted in a lack of reliable and

comprehensive data on the stability of the damaged

older structures, a need for remote control

The radioactive waste in the CEZ does not

include the waste associated with the decommissioning of

Chernobyl nuclear power plant units 1, 2 and 3.

142

concreting and the impossibility of carrying out

welding in some specific situations.

7.1.2. Current status of the damaged unit 4 and

the shelter

The shelter [7.4] was constructed using steel

beams and plates as structural elements. Its

foundation rests at some points on the original

structural elements of unit 4, whose structural

integrity, following the accident, is not well known.

At other points it rests on debris remaining from the

accident. Thus the ability of the shelter structure to

withstand natural events such as earthquakes and

tornados is known only with large uncertainties. In

addition to uncertainties on the structural stability

at the time of its construction, structural elements of

the shelter have degraded as a result of moisture

induced corrosion during the nearly 20 years since

the accident.

The shelter has approximately 1000 m

openings in its surface. These openings allow

approximately 2000 m

/a of precipitation to

percolate through the radioactively contaminated

debris and eventually to pool in rooms in the lower

levels of unit 4 (see Fig. 7.2) [7.5]. Condensation

within unit 4 of approximately 1650 m

/a of water

and the residues from periodic spraying of 180 m

of liquid dust suppressant contribute to the

quantities of water percolating through the unit 4

debris and collecting in its basement. The collected

water is contaminated with

137

Cs,

Sr and

transuranic elements, resulting in average concen-

trations of 1.6 × 10

Bq/m

137

Cs, 2.0 × 10

Bq/m

Sr, 1.5 × 10

Bq/m

of plutonium and 6 mg/L of

uranium. About 2100 m

/a of the collected water

evaporates, and about 1300 m

/a leaks through the

foundation into the soil beneath unit 4 [7.6]. The

existing Chernobyl nuclear power plant radioactive

waste management system is not capable of treating

liquid radioactive waste that contains transuranic

elements.

The inside conditions of unit 4 (Fig. 7.3) are

hazardous and present significant risks to workers

and the environment. General area radiation dose

rates range from 2 µSv/h to 0.1 Sv/h inside the

FIG. 7.1. The destroyed reactor after the accident in 1986.

143

shelter [7.5]. Individual occupational radiation

exposures during current operations at unit 4 are

controlled so that they do not exceed the dose limit

of 20 mSv/a [7.7].

Unit 4 is ventilated during current activities

through a monitored exhaust above the reactor

room. The unfiltered exhaust air is normally below

the permitted limits for atmospheric discharge, and

a filtration system exists for use should the exhaust

air levels approach the permitted discharge limits.

The ventilation system is zoned so that air flows

from outside the shelter through spaces with

increasing levels of contamination.

Unit 4 and the associated cascade walls have

an accumulation of FCM, including large core

fragments that could conceivably lead to criticality

under flooded conditions. Such a criticality accident

is considered unlikely; however, if criticality should

occur, it might lead to the exposure of some workers

inside unit 4 to an external dose of only a few

millisieverts, because workers tend to avoid the

spaces with criticality risk. It has been estimated

that, in such a case, there would be no significant

consequences inside and outside the CEZ [7.5, 7.8,

7.9].

A number of activities have been performed

in recent years to stabilize and improve the

conditions of the shelter. These include: repair of

the unit 3/4 ventilation stack foundation and

bracing; reinforcement of the B1 and B2 beams

(Fig. 7.4); improvement of the physical protection

and access control system; design of an integrated

automated control system (control of building

structure conditions, seismic control, nuclear safety

control and radiation control); modernization of the

dust suppression system; and additional structural

stabilization. Computerized control systems were

installed in the shelter [7.9] to monitor gamma

radiation, neutron flux, temperature, heat flux,

concentrations of hydrogen, carbon oxide and air

FIG. 7.2. Opening in the shelter allowing infiltration of

atmospheric water.

FIG. 7.3. Reactor room of unit 4 after the accident.

FIG. 7.4. Major structural components of the Chernoby

helter: (A) pipe roofing; (B) southern panels; (C)

outhern hockey sticks; (D) B1/B2 beams; (E) mammoth

beam; (F) octopus beam.

144

moisture, mechanical stability of structures, etc. This

has been achieved with significant support from

Ukraine and donor countries

The magnitude and importance of possible

future radioactive releases from the shelter (in the

event of its collapse) significantly depend on the

radiological and physicochemical properties of the

radioactive material, including dust that may arise

from the area inside the shelter. Now, nearly 20

years after the accident, dust has penetrated

concrete walls, floors and ceilings, and is in the air in

the form of aerosols. Thus in a number of shelter

premises the fuel-containing dust has become the

main source of radiation hazard. Research [7.5,

7.10] shows that the typical size of these particles

(activity median aerodynamic diameter) is from 1 to

10 µm. Hence most of the material is expected to be

respirable, which increases its potential inhalation

hazard. The potential for inhalation hazard is

increased by the winds that may be generated if the

shelter roof were to collapse.

Should the shelter collapse, it would also

complicate continuing accident recovery efforts,

and the resulting radioactive dust cloud would have

adverse environmental impacts. Further analysis of

the environmental release is sensitive to the source

term assumed in the dust cloud that would be

generated as a result of the collapse. Different

studies give different possible radioactive dust

releases to the environment, ranging from about

500 to 2000 kg of particulate dust, which could

contain from 8 to 50 kg of finely dispersed nuclear

fuel. Regardless of the source term assumption,

almost all material that might be raised into the

atmosphere by a shelter collapse is expected to be

deposited within the CEZ [7.11, 7.12].

Another concern related to the FCM is its

possible transport out of the shelter into

groundwater through the accumulated water. The

potential for FCM to dissolve in the accumulated

water was confirmed when bright yellow stains and

faded pieces of FCM were found on the surface of

solidified fuel–lava streams in unit 4 [7.3];

subsequent analysis proved the presence of soluble

uranium compounds. Until recently, this FCM was

considered to be a glassy mass that was very

insoluble. The possibility of leaching of radionu-

clides from the FCM and of mobile radionuclides

such as

Sr migrating and reaching the Pripyat

River was expected to be very low [7.9]. The

expected significance of this phenomenon is not

known, and therefore monitoring of the evolving

groundwater situation at and around the shelter is

important.

Additional studies of the water table showed

that it has risen by up to 1.5 m in a few years to

about 4 m from the ground level, and may still be

rising. This effect is considered to have occurred

mainly as a result of the construction of a 3.5 km

long and 35 m deep wall around unit 4 that aimed to

protect the Kiev reservoir from potential contami-

nation through groundwater [7.9].

The main potential hazard associated with the

shelter is a possible collapse of its top structures and

the release of radioactive dust into the

environment; therefore a dust suppression system

was installed beneath the shelter roof that periodi-

cally sprays dust suppression solutions and fixatives.

The system has operated since January 1990, and

more than 1000 t of dust suppressant has been

sprayed during this period.

7.1.3. Long term strategy for the shelter and the

new safe confinement

In order to avoid a collapse of the shelter,

some measures have been implemented and

additional measures are planned to strengthen

unstable parts of the shelter and to extend their

stability from 15 to 40 years [7.13]. In addition, the

NSC is planned to be built as a cover over the

existing shelter as a longer term solution (see

Fig. 7.5). The Ukrainian Government supports the

concept of a multifunctional facility with at least 100

years service life. This facility aims to reduce the

probability of shelter collapse, reduce the conse-

quences of a shelter collapse, improve nuclear

safety, improve worker and environmental safety,

and convert unit 4 into an environmentally safe site.

The construction of the NSC is expected to allow for

the dismantlement of the current shelter, removal of

FCM from unit 4 and the eventual decommissioning

of the reactor.

The specific operational aspects related to the

construction and operation of the NSC, including

maintenance in the long term, have not yet been

identified. It is important to note that the NSC

Contributors to the work of the Chernobyl

Shelter Fund include Austria, Belgium, Canada,

Denmark, the European Commission, Finland, France,

Germany, Greece, Ireland, Italy, Kuwait, Luxembourg,

the Netherlands, Norway, Poland, Spain, Sweden,

Switzerland, Ukraine, the UK and the USA. Additional

donors to the Fund include Iceland, Israel, Korea,

Portugal, Slovakia and Slovenia.

145

design is based on the current plans for removal of

the FCM that depend on the availability of a final

geological disposal facility about 50 years from now

[7.13]. This extended dormancy period could result

in the dispersal of the special human resources

needed to remove and dispose of the FCM safely.

Accordingly, there are good reasons for removing

the FCM and structural material as soon as possible

after the construction of the NSC.

7.1.4. Environmental aspects

7.1.4.1. Current status of the shelter

At present, the environmental contamination

around the Chernobyl nuclear power plant site is

due to the initial radioactive contamination of the

area from the accidental release of 1986, the routine

releases of radionuclides through the ventilation

system of the shelter and the engineering and other

activities carried out in the CEZ. The main dose

contributing radionuclides within the CEZ around

the Chernobyl nuclear power plant site are

137

Cs,

Sr,

241

Am and

239,240

Pu (see also Section 3); the

distribution of these radionuclides is shown in

Fig. 7.6 [7.2].

7.1.4.2. Impact on air

Currently, radioactive aerosol releases into

the atmosphere from the shelter are considered to

result from two main sources: controlled releases

from the central hall of unit 4 into the environment

through the exhaust ventilation system and

ventilation stack No. 2, and uncontrolled releases

through the leaks in the roof and walls. Ventilation

stack No. 2 releases 4–10 GBq/a, which is many

times lower than the regulatory limit of 90 GBq/a

[7.9]. The uncontrolled releases depend on the

locations and areas of the openings in the external

structures and the air transfer rate through them,

which depends on many factors, such as temper-

ature, barometric pressure, humidity and wind

speed and direction.

As a result, the air in the immediate vicinity of

the shelter contains finely dispersed fuel particles

with concentrations of up to 40 mBq/m

137

Cs at

distances less than 1 km and 2 mBq/m

at about

3 km from the shelter. The aerosol particles have

radioactive compositions similar to those of the

fuel; the primary beta emitters are

Sr and

137

Cs,

while the alpha emitters are mostly plutonium and

241

Am. Inhalation doses to individuals outside the

shelter result from a combination of the ongoing

shelter releases and resuspended material from the

initial accident. If a person (worker) were to spend

an entire year adjacent to the shelter, a recent

inhalation dose assessment indicates that releases

would result in an annual dose of about 0.5 mSv,

which would decrease to about 0.0002–0.0005 mSv

beyond a distance of 10 km [7.14]. Inhalation doses

from the ongoing releases outside the CEZ are

significantly less than the dose limits for the

population [7.7].

7.1.4.3. Impact on surface water

The average concentrations of radionuclides

in surface water bodies are declining. In the Pripyat

River in 2003, for example, concentrations were

observed to be 0.05 (max. 0.12) Bq/L for

137

Cs and

0.15 (max. 0.35) Bq/L for

Sr [7.15]. The main

sources of radionuclides in the rivers in the CEZ

during ordinary and high water seasons continue to

be runoff from the watersheds situated outside the

immediate Chernobyl nuclear power plant area,

infiltrating waters from the Chernobyl nuclear

power plant cooling pond and old water

reclamation systems in the heavily contaminated

territories. During winter and low water seasons, the

radionuclide fluxes from regional groundwater

contribute the majority of the radionuclide

migration to the Pripyat River from this area.

However, the values of radionuclide flux from all

groundwater into surface water are still relatively

low, and the contribution of groundwater

contamination plumes from the temporary

FIG. 7.5. Planned NSC.

146

(a)

kBq/m

20 000

7500

4000

2000

750

400

200

· State border

· Railway

· Road

· Towns and villages

· Rivers and lakes

· Exclusion zone border

Scale 1:200 000

0 2 4

(b)

kBq/m

20 000

7500

4000

2000

750

400

200

· State border

· Railway

· Road

· Towns and villages

· Rivers and lakes

· Exclusion zone border

Scale 1:200 000

0 2 4

IG. 7.6. Surface contamination by radioactive fallout within the CEZ [7.2]. (a) Caesium-137 in soils of the CEZ in 199

(kBq/m

); (b)

Sr in soils of the CEZ in 1997 (kBq/m

); (c)

241

Am in soils of the CEZ in 2000 (kBq/m

); (d)

239,240

Pu in soils

of the CEZ in 2000 (kBq/m

147

(c)

kBq/m

· State border

· Railway

· Road

· Towns and villages

· Rivers and lakes

· Exclusion zone border

Scale 1:200 000

0 2 4

1000.0

400.0

200.0

100.0

40.0

20.0

10.0

4.0

1.0

0.4

0.1

0.0

(d)

kBq/m

· State border

· Railway

· Road

· Towns and villages

· Rivers and lakes

· Exclusion zone border

Scale 1:200 000

1000.0

400.0

200.0

100.0

40.0

20.0

10.0

4.0

1.0

0.4

0.1

0.0

148

radioactive waste facilities and the shelter area has

been identified as about 3–10% [7.15] of the annual

migration of radionuclides into the Pripyat–Dnieper

River system from the CEZ (see also Section 3.5).

7.1.4.4. Impact on groundwater

Surface contamination around the Chernobyl

nuclear power plant site is the cause of groundwater

contamination, with groundwater levels of 100–

1000 Bq/m

for

Sr and about 10–100 Bq/m

for

137

Cs. Radionuclide contamination of the

groundwater at the shelter site is much higher. In

recent studies, the primary source term for radio-

nuclide contamination of the groundwater is

considered to be water accumulating inside the

underground rooms of unit 4 (as a result of precipi-

tation), groundwater accumulated near the pioneer

wall (because of absence of a drainage system) and

other water infiltrating from the nuclear power

plant site.

In some places,

137

Cs in groundwater in the

subsurface horizons near the shelter reaches 100 Bq/L

and even 3000–5000 Bq/L. However, in the majority

of the shelter area,

137

Cs concentrations in

groundwater are more or less similar and vary from

1 to 10 Bq/L. Typical concentrations of

Sr in

groundwater around the shelter site are in the range

from 2 to 160 Bq/L, with maximum concentrations

observed during the past five years ranging from

1000 to 3000 Bq/L. Estimated concentrations of

transuranic elements in the groundwater of this area

also vary over a wide range, from 0.003 to 3–6 Bq/L

for

238

Pu and

239,241

Pu, and from 0.001 to 8–10 Bq/L

for

241

Am [7.16, 7.17].

7.1.4.5. Impacts of shelter collapse without the new

safe confinement

Owing to concerns about the long term

stability of the shelter, estimates have been made of

the probability of its collapse. Depending on the

mechanisms considered, the probability ranges

from about 0.001 to 0.1/a [7.5, 7.18], and therefore

an analysis (summarized from Ref. [7.6]) of the

potential impacts of a shelter collapse has been

performed for scenarios without and with the NSC

in place.

(a) Impact on air

Collapse of the shelter could raise a large

cloud of fine dust (up to 500–2000 kg) containing 8–

50 kg of nuclear fuel particles with an activity of

about 1.6 × 10

Bq. This could lead to an additional

annual inhalation dose of up to 0.4 Sv near the

shelter. The estimated annual doses outside the

CEZ could reach 2 mSv [7.6], which would exceed

the established dose limits for the public in Ukraine

[7.7].

Within the boundaries of the CEZ, the

depositions of radionuclides from such a collapse

would be, in all cases, a small fraction of the existing

contamination levels caused by the original

Chernobyl accident. Typical results are shown in

Fig. 7.7 [7.8]. The highest relative increase in soil

contamination would occur if the wind were to blow

the plume from a shelter collapse to the south-west

towards the area that received the least impact from

the original accident. In this case, the additional

deposition might add about 10% to the existing soil

contamination levels. Outside the CEZ boundaries,

at 50 km from the shelter, additional surface

contamination with

137

Cs,

Sr and

238,239,240

Pu due to

a shelter collapse would contribute from a few to

10%.

(b) Impact on surface water

In the event of a shelter collapse, additional

radioactive material could also be deposited in and

near the rivers.

As shown in Fig. 7.7, radionuclide

depositions into the Pripyat River could be as high

as 1.1 × 10

Bq of

Sr, 2.4 × 10

Bq of

137

Cs,

1.6 × 10

Bq of

238

Pu, 4.0 × 10

Bq of

239,240

Pu and

5.0 × 10

Bq of

241

Am. Estimates of the maximum

possible concentrations of these radionuclides in

the Dnieper reservoirs show that the peak concen-

tration of

Sr can be expected in the Kiev

reservoir on the 41st day after the accident

happens, and would be about 700 Bq/m

. The

maximum concentration of

Sr in the Kakhovka

reservoir would be about 200 Bq/m

or less. This

confirms that the normative values of

Sr in

potable water (2000 Bq/m

[7.7]) would not be

exceeded if an accident leading to the maximum

impacts were to occur at the shelter.

The maximum possible concentrations of

137

that can be expected in the Kiev and Kanev

reservoir water, even in the worst simulated

scenarios, are three to ten times lower than the

limits for potable water. Such a collapse would not

affect the concentrations of

238

Pu,

239,240

Pu and

241

Am in the Pripyat and Dnieper Rivers [7.6].

149

(a)

kBq/m

(b)

kBq/m

FIG. 7.7. Strontium-90 soil density of Chernobyl fallout (a) upstream of Yanov bridge, 1999, and (b) predicted from a shelte

collapse [7.6]. The distance from the shelter is given on both axes.

150

Releases from a shelter collapse could lead to

some increased exposure of people living

downstream of the most impacted area in the CEZ

and who consume water and fish from the

reservoirs. Radiation doses to individuals are

discussed in Ref. [7.6], in which the highest values

are predicted to be for professional fishermen and

typical consumers.

Rainwater infiltration and condensation

within the existing shelter have also been studied

[7.6]. The radiological significance of the large pool

of water in the basement of the shelter was

confirmed. The leakage of the heavily contaminated

water from this pool through the concrete walls and

floor of the room is a main source of the contami-

nation of the vadose zone and groundwater beneath

the shelter. Under existing conditions, a positive

water balance exists and water collects in the

basement rooms.

The results of an assessment of groundwater

contamination without the NSC show that a

concentration of

Sr in the groundwater of about

4×10

Bq/m

is expected to occur at distances less

than 100 m from the shelter, and would decline to

100 Bq/m

at 600 m from the shelter. The contami-

nation is predicted to reach the Pripyat River in 800

years. However, the infiltration fluxes of

Sr from

the shelter even without the NSC are not expected

to cause significant impacts on the Pripyat River.

7.1.4.6. Impacts of shelter collapse within the new

safe confinement

(a) Impact on air

Placement of the NSC over the shelter is

expected to reduce the release of dust on to the site

resulting from a collapse, thus reducing the

magnitude of inhalation doses. The dust would

largely settle within the NSC and not be released to

the environment except through normal ventilation

pathways. The amount of transported dust would

depend on the ventilation and the confinement

capability designed into the NSC. The doses are

expected to be reduced by factors of seven to 70

compared with the estimated doses in the event of a

shelter collapse without the NSC, depending on the

capacity of the NSC ventilation system [7.6]. This

leads to an expected decrease of outdoor workers’

exposure by a factor of two in comparison with the

scenario of a shelter collapse without the NSC.

However, some workers might be inside the NSC at

the time of the collapse; doses to these workers

might be increased because of containment of the

dust.

For the small number of individuals who have

chosen to reside within the CEZ, inhalation doses

are expected to be reduced by factors of 50 to 500,

to no more than 1 or 2 mSv [7.6]. Even assuming the

worst (95th percentile) meteorological conditions

and also assuming that the dust cloud passes over

one of the larger cities, such as Slavutych, an

increase of latent fatal cancer risk projected for the

population in the event of a collapse with the NSC is

not expected.

Very minor additions to soil contamination

would be caused by the discharge and deposition of

airborne radionuclides should the shelter collapse

inside the NSC. Within the boundaries of the CEZ,

the radionuclide deposition would be in all cases a

small fraction of the existing levels caused by the

original Chernobyl accident. The highest relative

increase would occur if the wind were to blow the

plume from a shelter collapse to the south-west

towards the area that received the least impact from

the original accident; in this case, the additional

deposition might add less than 0.2% to the existing

soil contamination levels.

(b) Impact on surface water

Emplacement of the NSC would ensure that,

in the event of a collapse, additional deposition of

radionuclides on surface water would be minimal.

The depositions illustrated in Fig. 7.7 would be

reduced by factors of 50 to 500 [7.6], and the

resulting concentrations in downstream waters

would not exceed the Ukrainian norms.

The dynamics of radionuclide migration into

the groundwater with the presence of the NSC was

evaluated assuming that the water level is reduced

to zero in the basement one and a half years after

the NSC is constructed. After NSC construction, the

precipitation fluxes are expected to be minimized

and evaporation fluxes will be higher than fluxes

from dust suppression and condensation. This

means that the water level in the basement is

expected to diminish due to seepage through the

walls, and the room would therefore be empty in

less than two years.

151

7.1.5. Issues and areas for improvement

7.1.5.1. Influence of the source term uncertainty on

environmental decisions

There is still considerable uncertainty

regarding the amount of nuclear fuel remaining in

unit 4. One estimate [7.1, 7.19] is that inside unit 4

there is approximately 95% of the 190 t of nuclear

fuel (as uranium) that was in the reactor at the time

of the accident. Another estimate [7.12] is that there

is 60% of the original core plus the fuel in the decay

pool and in the central room remaining in the

facility (212 t total, less 80 t of FCM, less 6 t of

blown out fuel = 126 t remaining). The estimated

radioactivity inside the shelter in 1995 was approxi-

mately 7 × 10

Bq [7.3]. Despite these and other

studies, to date there is no comprehensive

information regarding the amount and distribution

of fuel inside the shelter. This lack of knowledge is

an important factor in the evaluation of the safety

and environmental consequences of unit 4 and the

shelter evolution, as well as for the selection of

adequate solutions for the long term management

of the associated radioactive waste.

7.1.5.2. Characterization of fuel-containing material

The physical state of the FCM appears to be

changing with time. It appears that the FCM has

begun to oxidize and may be decomposing into fine

particulate matter with an unknown oxidation rate,

particle size and behaviour. Another related

important uncertainty is on the dust distribution in

the shelter, and more specifically in the NSC

atmosphere over long term operation of the facility.

As estimates of the environmental impacts (e.g.

transport and inhalation calculations) of long term

shelter development are sensitive to the

assumptions about these source term parameters, it

is necessary that these parameters be further inves-

tigated. This would contribute to increased

confidence in the safety assessment results and the

selection of appropriate protective measures for

workers, the public and the environment.

7.1.5.3. Removal of fuel-containing material

concurrent with development of a geological

disposal facility

The stabilization of the shelter and

construction of the NSC is expected to generate

significant amounts of long lived radioactive waste,

some of which would contain FCM. However, there

are no plans for removal of the FCM until a

geological disposal facility is constructed and

commissioned. A long term management strategy

for the FCM and long lived radioactive waste

therefore needs to be developed to ensure safe

management of this waste.

It can be concluded that there is no technical

reason to delay removal of the FCM until a

geological disposal facility is available. Removal of

FCM could commence following dismantlement of

the unstable structures of the shelter, continuing

with radioactive waste predisposal management and

temporary storage on the Chernobyl nuclear power

plant site until the geological disposal facility

becomes available. Due to the high content of long

lived radionuclides there is also no significant

worker dose benefit to be obtained from waiting for

the availability of a geological disposal facility.

Whether retrieved now or after 50 years, remote

retrieval and radioactive waste management

techniques will be required to remove the FCM and

restore the unit 4 site.

7.2. MANAGEMENT OF RADIOACTIVE

WASTE FROM THE ACCIDENT

In the course of remediation activities, both at

the Chernobyl nuclear power plant site and in its

vicinity, large volumes of radioactive waste were

generated and placed in temporary near surface

waste storage facilities located in the CEZ (Fig. 7.8)

at distances of 0.5–15 km from the nuclear power

plant site. Sites for temporary waste storage of the

FIG. 7.8. Temporary radioactive waste disposal facilities in

the territory of the CEZ.

152

trench and landfill type were created from 1986 to

1987, intended for radioactive waste generated after

the accident as a result of the cleanup of

contaminated areas, to avoid dust spread, reduce

radiation levels and provide better working

conditions at unit 4 and its surroundings. These

facilities were established without proper design

documentation, engineered barriers or

hydrogeological investigations, which are required

by contemporary waste safety requirements.

During the years following the accident,

economic and human resources were expanded to

provide a systematic analysis and an acceptable

strategy for the management of existing radioactive

waste. However, as reported in some Ukrainian

studies [7.20], to date a broadly accepted strategy

for radioactive waste management at the Chernobyl

nuclear power plant site and the CEZ, and

especially for high level and long lived waste, has

not been developed. Some of the reasons are the

large number of and areas covered by the

radioactive waste storage and disposal facilities, of

which only half are well studied and inventoried.

This results in large uncertainties in the documented

radioactive waste inventories (volume, activity,

etc.).

The existing radioactive waste from the

accident and potential radioactive waste to be

generated during the NSC construction, shelter

dismantling, FCM removal and decommissioning of

unit 4 can be categorized as:

(a) Radioactive waste from the shelter and the

nuclear power plant site that will be created by

construction of infrastructure and the NSC;

(b) Accident generated transuranic waste that has

been mixed with radioactive waste from

operations at Chernobyl nuclear power plant

units 1, 2 and 3;

waste facilities located throughout the CEZ;

(d) Radioactive waste in existing radioactive

waste disposal facilities.

The safety and environmental issues related to

each of these categories of radioactive waste are

presented in this section. The radioactive waste

expected to be generated during the decommis-

sioning of the Chernobyl nuclear power plant units

1, 2 and 3 represents an additional category that is

not the subject of this report.

The current Ukrainian legislation applies a

categorization of radioactive waste in accordance

with its specific activity and radiotoxicity, specified

in Table 7.1 [7.21].

For waste contaminated with unspecified

mixtures of radionuclides emitting gamma

radiation, use of the classification ‘low’, ‘interme-

diate’ and ‘high’ activity is allowed using the air

dose rate at a distance of 0.1 m, as specified in Table

7.2 [7.21].

Current radioactive waste management

practice in Ukraine does not fully comply with the

above classification; measures are therefore being

taken to bring it into conformity with the new

regulations [7.22].

TABLE 7.1. UKRAINIAN SOLID RADIOACTIVE WASTE CATEGORIZATION [7.21]

Range of specific activity (kBq/kg)

Group 1

Group 2

Group 3

Group 4

Low activity 10

–1

–10

Intermediate activity 10

–10

High activity >10

>10

Group 1: transuranic alpha radionuclides; group 2: alpha radionuclides (excepting transuranium); group 3: beta and gamma

radionuclides (excluding those in group 4); group 4:

Cl,

Cf,

Mn,

Fe,

Ni,

93m

Nb,

Tc,

109

Cd,

135

Cs,

147

Pm,

151

Sm,

171

Tm and

204

Tl.

TABLE 7.2. CLASSIFICATION OF RADIO-

ACTIVE WASTE WITH UNKNOWN SPECIFIC

ACTIVITY USING THE DOSE RATE AT 0.1 m

DISTANCE [7.21]

Dose rate (µGy/h)

Low activity 1–100

Intermediate activity 100–10 000

High activity >10 000

153

7.2.1. Current status of radioactive waste from

the accident

7.2.1.1. Radioactive waste associated with the shelter

The shelter is considered to be “the destroyed

unit 4 after a radiological accident” and “a near

surface storage facility for unconditioned

radioactive waste at a stage of stabilization and

reconstruction” [7.22, 7.23]. The amount and type of

waste, debris and other radioactive material inside

the shelter is presented in Table 7.3.

In addition, soil that was heavily contaminated

by the deposition of fuel fragments and with radio-

nuclides and debris from the accident (metal pieces,

concrete rubble, etc.) was also collected and stored

in the vicinity of unit 4:

(a) Three pioneer walls (west, north and south of

the shelter), where contaminated soil,

concrete and containers are stored and which

contain an estimated 1700–4900 m

of high

level waste

and up to 72 000 m

of low and

intermediate level waste [7.25, 7.26].

(b) The cascade wall north of the shelter, where

core fragments, metal, concrete, core pit

equipment and accident cover material are

stored (16 600 m

of high level waste, 117 t of

reactor core elements and 53 400 m

of low

and intermediate level waste) [7.25].

concrete, gravel, sand, clay and contaminated

soil are stored that contain 7000 m

of high

level waste and 286 000 m

of low and inter-

mediate level waste [7.27]. Other studies show

that fuel, graphite, etc., are located in the

contaminated soil [7.26].

The radioactive waste inside the pioneer and

cascade walls was later covered with concrete. This

material is considered to be high level waste that is

not acceptable to be disposed of in near surface

disposal facilities. Since it cannot be retrieved easily

for conditioning, the radioactive waste recovered

from these walls is to be part of a global strategy for

the decommissioning of unit 4.

TABLE 7.3. ESTIMATED INVENTORY IN THE SHELTER [7.25]

Type of radioactive waste and

criteria of assessment

Category of

radioactive waste

Amount

FCM Fresh fuel assemblies, spent fuel

assemblies, lava type material, fuel

fragments, radioactive dust

High level About 190–200 t, 700 t

of graphite

Solid radioactive waste with less

than 1% nuclear fuel (mass)

Fragment of the core with a dose rate

at 10 cm of more than 10 mSv/h

Liquid radioactive waste Changing inventory based on

precipitation (e.g. pulp, oils,

suspensions with soluble uranium

salts)

Low level (up to

3.7 ¥ 10

Bq/L)

2500–5000 m

Intermediate

level (more than

3.7 ¥ 10

Bq/L)

500–1000 m

Solid radioactive waste Metal equipment and building

material, for example concrete, dust,

non-metal material (organic, plastic)

High level 38 000 m

(building

material), 22 240 t

(metal constructions)

Low and

intermediate

level

300 000 m

(building

material and dust),

5000 m

(non-metal)

High level waste falls into two subcategories: low

temperature waste with a heating rate of less than

2kW/m

and heat generating waste with a heating rate

higher than 2 kW/m

[7.24].

154

It is estimated that the current and expected

radioactive waste from unit 4 can be categorized as

short lived low and intermediate level waste (soil

from the construction of the NSC, construction

material, concrete, metal constructions, etc.) and

high level waste (e.g. FCM) according to Ukrainian

legislation [7.28, 7.36].

7.2.1.2. Mixing of accident related waste with

operational radioactive waste

During 1986–1993, some low and intermediate

level radioactive waste and high level waste with

transuranic elements were stored together with

some operational radioactive waste from units 1, 2

and 3 in an above ground storage facility (see

Fig. 7.9) at the Chernobyl nuclear power plant site.

This waste amounts to about 2500 m

, with a

total radioactivity of about 131 TBq [7.19], and is

stored unconditioned. Once filled, the storage

facility was backfilled with concrete grout and

covered with a concrete roof to reduce radiation

levels and water infiltration. Thus the retrieval of

the radioactive waste stored in this facility cannot

be easily achieved and will require particular care.

Plans for such retrieval are currently under study.

At present, this facility is being extended and is

intended to be used for the disposal of radioactive

waste produced during the decommissioning of

units 1, 2 and 3.

7.2.1.3. Temporary radioactive waste storage

facilities

The largest volumes of radioactive waste

generated by unit 4 remediation activities are

located in the CEZ (see Fig. 7.8). Sites for

temporary storage of radioactive waste, of the

trench and landfill type, were constructed shortly

after the accident at distances of 0.5–15 km from

the nuclear power plant site. They were created

from 1986 to 1987 and intended for radioactive

waste generated after the accident as a result of the

cleanup of contaminated areas to avoid dust

spread, reduce radiation levels and provide better

working conditions at unit 4. These facilities were

established without design documentation,

engineered barriers or hydrogeological

investigations.

The total area of temporary radioactive waste

facilities is about 8 km

, with the total volume of

disposed radioactive waste estimated to be over

. The main inventories of activity are concen-

trated in the Stroibaza and Ryzhy Les temporary

radioactive waste facilities along the western trace

of the Chernobyl fallout (see Fig. 7.8). The specific

activity of the radioactive waste in the temporary

radioactive waste facility at Ryzhy Les is

Bq/kg of

Sr and

137

Cs and 10

–10

Bq/kg of

plutonium isotopes (total).

Most of the facilities are structured in the form

of trenches 1.5–2.5 m deep in the local sandy soil.

The radioactive material (soil, litter, wood and

building debris) is overlain by a layer of alluvial

sand 0.2–0.5 m thick. The majority of the temporary

radioactive waste facilities consist of trenches in

various types of geological setting, in which waste

was stacked and covered with a layer of soil from

the nearby environment. These facilities are

therefore very variable with regard to their

potential for release, which depends on the total

radioactivity stored, the waste form (in particular

timber), the retention capacity of the substratum

along migration pathways and the location of the

sites in hydrogeological settings. At least half of

these temporary radioactive waste facilities have

been studied (see Table 7.4) [7.19, 7.29].

There are also many other temporary

radioactive waste facilities, estimated to comprise

about 800 trench facilities each with waste

disposal volumes in the range of 8 × 10

2×10

[7.29, 7.30]. The inventories of these

facilities are known for about half of them. The

facilities are not under regulatory control.

Estimates made for a few sites show that their

radioactive contents can be high (10–1000 TBq),

sometimes of an order of magnitude comparable

with the total radioactivity present in soil in the

CEZ (about 7000 TBq) [7.30].

FIG

. 7.9. Existing above ground storage facility for solid

radioactive waste at the Chernobyl nuclear power plant site.

155

7.2.1.4. Radioactive waste disposal facilities

The main radioactive waste disposal facilities

for accident waste are the Buriakovka, Podlesny

and Kompleksny sites, which are under regulatory

control. These three near surface disposal sites were

established after the accident to dispose of

radioactive waste from remediation actions carried

out during the first year following the accident.

These sites were chosen and designed for the

disposal of higher level accident waste than the

radioactive waste located in the temporary

radioactive waste facilities [7.19].

Buriakovka, built in 1987, is the only disposal

facility currently in operation in the CEZ. It

comprises 30 trenches covered with a 1 m clay layer

and is located on 23.8 ha. Up to 652 800 m

radioactive waste has been disposed of. After in situ

compaction, this was reduced to 530 000 m

, with a

total radioactivity of 2.5 × 10

Bq of solid short

lived low and intermediate level waste. It consists of

metal, soil, sand, concrete and wood contaminated

with

Sr,

137

Cs,

134

Cs,

238,239,240

Pu,

154,155

Eu and

241

Am. Radioactive waste with dose rates at 10 cm

from the surface in the range of 0.003–10 mGy/h

was accepted in this facility.

The Podlesny vault type disposal facility was

commissioned in December 1986 and closed in

1988. The facility was designed for the disposal of

high level waste with a dose rate 10 cm from the

surface in the range of 0.05–2.5 Gy/h. Material with

dose rates above this was also disposed of in the

facility. The total radioactive waste volume of

11 000 m

of building material, metal debris, sand,

soil, concrete and wood was placed in two vaults.

The disposal facility was covered with concrete at its

TABLE 7.4. STATUS OF TEMPORARY RADIOACTIVE WASTE FACILITIES [7.19, 7.29]

Size

(ha)

Number of

trenches

Number of

landfills

Radioactive waste type

Radioactive

waste volume

(10

)

Total activity

(Bq)

Sites with well known inventories

Neftebaza 53 221 4 Soil, plants, metal, concrete

and bricks

104 4 ¥ 10

Peschannoe

Plato

78 2 82 Short lived

low and

intermediate level waste

of soil, rubble and concrete

57 7 ¥ 10

Partially investigated sites

Stantzia

Yanov

128 Known: more

than 36

— Soil, plants, metal,

concrete and bricks

30 >4 ¥ 10

Ryzhy Les 227 Estimated at

more than 61

Estimated at

more than 8

Mainly soil, some construction

and domestic material

500 Up to 4 ¥ 10

Staraya

Stroibaza

130 More

than 100

— Soil, metal, concrete

and wood

171 1 ¥ 10

Novaya

Stroibaza

122 — — Soil, plants, metal,

concrete and bricks

150 2 ¥ 10

Pripyat 70 — — Contaminated vehicles,

machinery, wood and

construction waste

16 3 ¥ 10

(1990)

Chistogalovka 6 — — Material from demolition

of buildings, soil, wood and

work clothes

160 4 ¥ 10

Kopachi 125 — — Construction waste from

demolition

110 3 ¥ 10

According to Ukrainian legislation, short lived waste is radioactive waste whose release from regulatory control is

achieved earlier than 300 years after disposal; long lived waste is radioactive waste whose release from regulatory

control is achieved later than 300 years after disposal [7.21].

156

closure. In 1990 the estimated total radioactivity of

the disposed waste was 2600 TBq. In 2002 a re-

evaluation of the facility status showed reasons to

believe that the total activity of waste disposed at

this site may be higher than initially estimated, and

a need for a re-estimation of the current inventory

was identified. Due to the uncertainties in the

inventory, it is assumed that various types of waste

were disposed of, including FCM.

The Kompleksny vault type facility was based

on reconstructed facilities of the unfinished units 5

and 6 at the Chernobyl nuclear power plant site.

Kompleksny was in operation from October 1986

until 1988 and was designed for low and interme-

diate level waste corresponding to dose rates up to

0.01 Gy/h at 10 cm from the surface of the waste

container. More than 26 200 m

of solid waste with a

total activity of 4 × 10

Bq was disposed of in 18 000

containers and later covered with sand and clay.

This waste is mainly sand, concrete, metal,

construction material and bricks. Due to the high

level of groundwater at different periods of the

year, the facility is flooded 0.5–0.7 m above its

bottom. Significant uncertainties exist associated

with the radionuclide inventory because of the lack

of data about the radioactive waste disposed of at

the site.

At present, a new near surface facility, the

Vektor complex, for low and intermediate level

radioactive waste processing, storage and disposal, is

under development. This complex will include [7.19]:

(a) An engineering facility for the processing of

all types of solid radioactive waste (capacity of

3500 m

/a);

(b) A disposal facility for short lived solid

radioactive waste (55 000 m

total capacity);

radioactive material;

(d) A storage facility for FCM;

(e) Intermediate storage for high level

conditioned radioactive waste to be prepared

for final disposal at the deep geological

disposal facility.

7.2.2. Radioactive waste management strategy

At present, no further dismantlement and

cleanup of unit 4 is planned. However, estimates of

the radioactive waste generation and subsequent

management options have been performed for the

construction of the NSC and the dismantlement

phase of the unstable structures of the shelter. The

preparation phase is expected to generate about

390 t of solid radioactive waste and about 280 m

liquid [7.6]. It also requires the removal of

100 000 m

of contaminated soil around unit 4,

which may still contain fuel fragments. Preliminary

studies for the dismantlement of the shelter super-

structure predict that about 1200 t of steel, with an

estimated volume of radioactive waste of 1800 m

[7.14], mainly metal and large concrete pieces, will

be removed. This waste is planned to be sorted

based upon its radiation level. High level waste,

which is expected to be only a small part, is planned

to be placed in containers and stored within the

NSC.

According to Ukrainian legislation [7.31], all

industrial radioactive waste is classified according to

the scheme shown in Fig. 7.10: high level and long

lived waste must be disposed of in a deep geological

disposal facility; low and intermediate level and

short lived radioactive waste in a near surface

disposal facility. Following these criteria, a strategy

for the management of radioactive waste from the

1986 accident needs to be developed, and in

particular for the management of high level and

long lived radioactive waste.

The planned options for low level waste are to

sort the waste according to its physical character-

istics (soil, concrete, metal, etc.) and, possibly, to

decontaminate it and/or condition it for beneficial

reuse (reuse of soil for NSC foundations, melting of

metal pieces) or send it for disposal in a new

extension of the Buriakovka disposal facility [7.19]

or the Vektor disposal site.

The long lived waste is planned to be placed in

interim storage. Different storage options are being

considered at the Chernobyl nuclear power plant or

the Vektor site, and a decision has not yet been made.

After construction of the NSC, decommissioning of

Below Exemption

Level

Low and

Intermediate

Level Waste

(LILW)

High Level Waste

(HLW)

High level waste

Long

lived

Short

lived

Interim

storage

Geological

disposal

Near surface

disposal

Low and

intermediate

level waste

Below exemption

level

Recycle or

dispose

without radiological

restrictions

FIG. 7.10. Planned management of radioactive waste at the

Chernobyl nuclear power plant site [7.19].

157

the shelter facilities is envisaged, including shelter

dismantlement and further removal of FCM. High

level radioactive waste will be partially processed in

place and stored at a temporary storage site until a

deep geological disposal site is ready. At present, this

strategy is considered as the preferred option for high

level radioactive waste and FCM [7.19]. To

implement this strategy, it is planned to organize a

system for processing and temporary storage of the

high level and long lived radioactive waste at the

Vektor facility complex that is now being developed.

When the Vektor facility is in full operation, it may

be possible to begin the work of removing FCM and

other radioactive waste from the shelter under cover

of the NSC.

Such a strategic approach is foreseen in the

comprehensive programme on radioactive waste

management that was approved by the Ukrainian

Government [7.25]. Prior to elaboration of such a

programme, special field studies and geological

investigations must be carried out in the CEZ and

its surrounding area, and in particular in the areas

with crystalline rocks that have a depth of more

than 500 m. According to Ref. [7.25], it is considered

reasonable to begin an investigation for exploring

the most appropriate geological site in this area in

2006. Following such planning, the construction of a

deep geological disposal facility might be completed

before 2035–2040.

Management of future liquid radioactive

waste from the shelter is planned to be performed at

the new liquid radioactive waste treatment plant at

the Chernobyl nuclear power plant site. However,

the management of liquid waste containing

transuranic elements remains an issue to be

resolved.

In addition, in the strategy for management of

radioactive waste from the accident, account should

be taken of the management of other storage sites

containing about 2000 pieces of contaminated

equipment (transport vehicles, helicopters, tanks,

etc.) that were used in the first months after the

accident and for which the final disposition has not

been determined.

7.2.3. Environmental aspects

Safety concerns in relation to most of the

temporary radioactive waste facilities in the CEZ

need to be viewed within the context that most of

these facilities are located in very contaminated

areas, with surface levels of

Sr in the range of 400–

20 000 kBq/m

, 700–20 000 kBq/m

for

137

Cs and 40–

1000 kBq/m

for

239,241

Pu. In this same territory, the

temporary radioactive waste facilities occupy a

relatively small volume covered with several metres

of soil and other geological material.

The major concern is the risk of increased

contamination of groundwater and the possibility, in

the future, that such contamination reaches major

water sources used as water supplies. The measure-

ments reported in the French–German initiative

[7.32] (see Table 7.5) clearly show that some

temporary radioactive waste facilities have a

significant influence on groundwater. In particular,

flooded and partially flooded trenches are giving

rise to enhanced migration, due to the absence of

engineered safety features. More favourable

settings, such as the Buriakovka site, lessen the

radionuclide release variations and maintain

concentrations in groundwater at comparatively low

levels.

For a part of the year, some temporary

radioactive waste facilities are very near or in the

groundwater table, which can affect radionuclide

dispersion. It is noticeable that water table levels at

the trenches and landfills vary from about 1 to 7 m

depth and vary depending on the season. Part of the

facilities at Stantzia Yanov and Neftebaza are

constantly flooded. Flooding is also an important

concern at the Kompleksny disposal site, where the

waste containers are flooded from 0.5 to 0.7 m from

TABLE 7.5. CONTAMINATION OF GROUNDWATER NEAR SELECTED TEMPORARY RADIO-

ACTIVE WASTE STORAGE FACILITIES IN 1994–1995 [7.16] AND IN 1999 [7.23]

Strontium-90

(Bq/L)

Caesium-137

(Bq/L)

Plutonium-239, 240

(Bq/L)

1994–1995 1999 1994–1995 1999 1994–1999

Ryzhy Les 100–120 000 100–230 0.1–100 0.1–2.5 0.4–0.6

Stroibaza 3–200 30–50 1–20 0.02–0.004 No data

Peschannoe Plato 3–10 2–40 0.7–3 0.02–0.1 No data

158

the bottom of the disposal facility [7.19]. The degree

of contamination is monitored in these disposal sites

using a monitoring system, established in 1986–

1989, that needs upgrading.

Monitoring results of groundwater contami-

nation around the temporary storage facilities

indicated concentrations of

Sr in the range of

100–100 000 Bq/m

[7.19, 7.33]. The highest levels of

contamination are detected at the northern part of

the Chernobyl nuclear power plant site,

groundwater from which also runs into the Pripyat

River. Therefore, actual and potential impacts from

radionuclides exist for those radioactive waste

facilities located immediately next to the riverside

in alluvial soils and which might be at regular risk of

flooding during high water periods [7.20, 7.32].

These types of disposal facility have been studied

during the past five years and continue to be studied

as a basis for their step by step removal and

relocation to the properly established disposal

facilities.

As mentioned above, the rate of radionuclide

migration with groundwater is much lower than the

hydraulic transport of the water itself. This means

that, due to retardation factors and geochemical

processes, the majority of radionuclides being

released from the body of the temporary

radioactive waste facilities are accumulating in the

geological media. Taking into account the

adsorption capacity of the soils and geological

media surrounding the temporary radioactive waste

facilities, several studies have shown that a

significant fraction of

Sr is still associated with the

fuel matrix, which delays its release to the pore

water in the soil for many years. As a result, radio-

nuclide concentrations in groundwater, even for

such mobile radionuclides as

Sr, are very low.

Plutonium isotopes (and

241

Am associated with

them) have not yet been adequately studied;

however, it is well known that their migration

beyond the temporary radioactive waste sites is

negligible (Fig. 7.11).

Studies of the vertical and lateral transfer

rates of radionuclides demonstrated that, for the

local soil, there is a low risk of radionuclide contam-

ination in groundwater and therefore a proportion-

ately low risk of significant contamination of the

Pripyat River in the future, as discussed also in

Section 3.5 (see Fig. 3.58). It has been shown that

the leading edge of contaminated groundwater from

most of the significant temporary radioactive waste

facilities may reach regional surface water within

100 or more years, making this issue of minimal

importance in terms of radiological impact for

populations living downstream of the Pripyat River

system [7.17, 7.34]. However, for the CEZ,

groundwater is still an important potential source

for radionuclide migration in the environment, and

therefore the waste facilities have to be under

regular monitoring and institutional control.

The long term strategy for the temporary

facilities is related to the management of the

Trench No. 22-T

Strontium-90 in moisture

sampler (Bq/L)

Strontium-90 in well (Bq/L)

X (m)

Elevation (m.a.s.l)

Trench No. 22

outline

Multilevel

well profile

Tracer

test site

FIG. 7.11. Spatial distribution of

Sr (Bq/L) in the groundwater near surface trench No. 22 of the Ryzhy Les facility in Octobe

1998 [7.34].

159

associated radiological risks. The ultimate goal

should be that waste is disposed of or left in

temporary radioactive waste facilities that ensure

sufficient confinement of

137

Cs and

Sr to allow

their decay without having generated significant

impacts on potential critical groups. For those

temporary radioactive waste facilities located near

the banks of the Pripyat River that could be

inundated by floodwaters, the preferred strategy is

to remove and relocate the waste into the properly

established disposal facilities.

For the temporary radioactive waste facilities

in the CEZ whose inventory is not known, and for

which the potential for future contamination of

surrounding groundwater and surface water is

uncertain, safety assessments need to be performed,

taking into account radioactive decay and natural

attenuation. There is clearly a need to assess, with

an increased level of confidence, the migration of

contamination plumes and their interface with

major water resources and supply areas (aquifers,

rivers, reservoirs, local supplies for the nuclear

power plant and the CEZ). Such assessments need

to consider all sources of release likely to affect

these water resources.

The safety assessment results will guide

decisions on appropriate remediation or institu-

tional control measures at the temporary sites.

Operational waste acceptance criteria (e.g. activity

concentrations) also need to be established to

ensure that potential exposures from various

scenarios remain acceptable, on the hypothesis that

resettlement in the CEZ occurs after some

hundreds of years. It is obvious that institutional

control at such disposal sites will need to be

maintained for a period of a few hundred years to

allow

137

Cs and

Sr activities to decay to insignif-

icant levels. This will require significant resources

for monitoring, implementation of recovery actions

and probably major restrictions on resettlement.

However, long term institutional controls should

not be considered as an alternative to recovery

actions to improve overall safety in the CEZ.

7.2.4. Issues and areas of improvement

7.2.4.1. Radioactive waste management programme

for the exclusion zone and the Chernobyl

nuclear power plant

A comprehensive programme for radioactive

waste management has not yet been established for

further cleanup of contaminated areas or temporary

radioactive waste facilities at the Chernobyl nuclear

power plant and within the CEZ. As mentioned

earlier, the ongoing strategy is to monitor the

temporary waste sites with the highest radiological

risk to the environment, so as to assess whether

cleanup or environmental protection actions are

needed. In addition, options for the long term

processing, storage and disposal of long lived and

high level waste from the Chernobyl nuclear power

plant and the CEZ, as well as management of liquid

waste contaminated with transuranic elements, are

to be selected and the necessary facilities

developed. Development of such a programme

could ensure the consistent and coordinated long

term management of all types of accident waste and

hence provide protection of workers, the public and

the environment.

7.2.4.2. Decommissioning of unit 4

Two main factors need to be addressed within

the strategy for dismantlement of the shelter and

decommissioning of unit 4: the safety implications

of the management of associated radioactive waste

(in particular of the high level waste) and the safety

implications of delaying recovery operations. The

strategy for the management of radioactive waste

that cannot be disposed of in near surface facilities

needs to be developed. Specifically, there is a need

for new waste management facilities (e.g. storage of

long lived waste, geological disposal), with consider-

ation given to the capacity of these facilities and

also the possibility of using the existing facilities for

the decommissioning of the Chernobyl nuclear

power plant. Particular attention needs to be given

to the establishment of an adequate infrastructure

and facilities for the management of long lived

waste (in particular large quantities of soil,

transuranic liquid waste and contaminated metal)

and high level waste (i.e. FCM) and their

subsequent disposal.

7.2.4.3. Waste acceptance criteria

The waste management programme being

implemented does include criteria for the

categorization of accident radioactive waste, which

are needed for the selection of an appropriate

management option for individual radioactive waste

streams. Adoption of criteria for waste

management, based on

137

Cs and alpha specific

radioactivity levels in the waste, is under

160

development. Although such criteria are more

adapted to appraising the potential for waste to be

accepted in near surface facilities, the question of

estimating the acceptable specific activities for

existing waste, especially in temporary radioactive

waste facilities, remains difficult to solve. The

development of waste acceptance criteria is

important in order to ensure protection of workers

and the environment, as well as the public, in the

long term.

7.2.4.4. Long term safety assessment of existing

radioactive waste storage sites

There is a need to identify the remaining

temporary radioactive waste storage facilities and to

appropriately mark them to prevent inadvertent

intrusion. The long term impact of these facilities on

the environment also needs to be evaluated in order

to estimate the need, where necessary, for imple-

mentation of upgrading or remedial actions.

Taking into account the large number of

facilities, there is also a need to prioritize the needs

for safety assessment. These assessments should

evaluate safety in the present conditions and with

consideration of possible future resettlement.

Consideration should be given to the need to

restrict the number of sites that are flooded or will,

in the future, need extensive control over some

hundreds of years.

In order to select those facilities with higher

radiological risk it is important to improve

methods for assessing the radioactive content of

the waste in the temporary facilities, especially of

long lived radionuclides. For pragmatic reasons,

this assessment should be based on a limited

number of parameters and measures. In this way,

uncertainties that affect present estimates of the

potential impact of individual facilities on the

environment will be reduced, and a consistent

assessment that takes into account all existing and

potential sources of contamination in the CEZ will

become feasible.

7.2.4.5. Potential recovery of temporary waste

storage facilities located in the Chernobyl

exclusion zone

Work is under way on the development of a

strategy for the management of temporary waste

storage facilities; this envisages three options for the

different facilities, depending on their status and

radiological hazard to the environment [7.19, 7.29]:

(a) Retrievability of waste and disposal in the

short term in order to minimize environ-

mental consequences and improve the safety

of workers; for example, industrial sites, the

shelter, the flooded temporary storage

facilities and the Kompleksny disposal facility.

(b) Possible temporary storage of waste under

institutional control in accordance with

radiation protection requirements with a view

to future disposal; for example, the Podlesny

disposal facility and contaminated equipment

from the activities aimed at mitigation of the

Chernobyl accident.

studied in order to decide on adequate inter-

vention measures; for example, temporary

radioactive waste facilities and soil from the

construction of the NCF.

7.3. FUTURE OF THE CHERNOBYL

EXCLUSION ZONE

The long term development of the CEZ is an

important and complex task that must consider

various technical, economic, social and other

factors; various options have been considered for

the evolution of this zone. According to Likhtarev

et al. [7.35], after 2015 about 55% of the territory

around the Chernobyl nuclear power plant could be

considered for release from radiological limitations

according to Ukrainian legislation. However, the

final decision on permitting people to return to this

zone must take into account the inhomogeneous

character of the contaminated land, specific

features of radionuclide migration and accumu-

lation in different portions of the local landscape,

and the routine habits of the population living in

this region (hunting, fishing, berry picking,

mushroom gathering, etc.).

The overall plan for the development of the

CEZ is to recover the affected areas of the CEZ,

redefine the CEZ and make the non-affected areas

available for resettlement by the public. This will

require well defined administrative controls as to

the nature of activities that may be performed in the

resettled areas, prohibition of growing of food crops

and cattle grazing and the use of only clean feed for

cattle. Accordingly, these resettled areas are best

suited for an industrial s

ite rather than for a

residential area.

For the reasons given above, the activities

focused on decontamination and dismantlement of the

161

shelter and on radioactive waste management in this

territory are expected to continue, which requires

optimal management of this area. The new concept

foresees division of the CEZ into different sections:

(a) The industrial zone is planned to include the

most contaminated areas, where the

Chernobyl nuclear power plant, facilities for

processing radioactive waste and main

radioactive waste storage areas are situated.

Primarily industrial activities are envisaged to

be carried out here, specifically the

construction of the NSC facilities. To provide

the infrastructure for NSC construction, new

roads, shipping yards, railways and other

support structures are planned. The town of

Chernobyl has been considered as an option

for such infrastructure development [7.6]. If

the CEZ is selected as the site for construction

of the geological repository for high activity

and long lived radioactive waste, a significant

amount of drilling and mining work will have

to be performed, which will also require

specific development of the engineering infra-

structure.

(b) The sanitary restricted zone is considered to

be a buffer area between industrial and nature

reserve areas.

located where most industrial and human

activities are prohibited, with the aim of

preservation of the basic natural landscapes

and biodiversity of the region.

The rehabilitation of the CEZ is expected to

create optimal conditions for industrial activity and

environmental protection for a long period of time;

for example, the NSC is expected to be operational

for at least 100 years. Different types of radioactive

storage facility must provide safe storage for 300 or

more years. A possible activity in this area may be

construction of the main geological disposal facility

for radioactive waste. A national engineering centre

may be established for processing all categories of

radioactive material and waste to be delivered to

the geological repository from different parts of

Ukraine.

Continued monitoring and support studies in

the CEZ are needed to form a basis for the review

and optimization of the management strategy in the

contaminated territories, and also for developing

basic and practical knowledge about the dynamics

and evolution of radionuclide migration, the need

for additional engineering barriers and the imple-

mentation of environmental remediation technol-

ogies.

In summary, the future of the CEZ for the

next hundred years and more is envisaged to be

associated with the following activities:

(i) Construction and operation of the NSC and

relevant engineering infrastructure;

(ii) De-fuelling, decommissioning and

dismantling of units 1, 2 and 3 of the

Chernobyl nuclear power plant and shelter;

(iii) Construction of facilities for the processing

and management of radioactive waste, in

particular a deep geological repository for

high activity and long lived radioactive

material;

(iv) Development of nature reserves in the area

that remains closed to habitation;

(v) Maintenance of environmental monitoring

and research activities.

7.4. CONCLUSIONS AND

RECOMMENDATIONS

7.4.1. Conclusions

It can be concluded that the existing uncer-

tainties associated with the stability of the shelter

structures, the radioactive inventory, the insufficient

confinement, the evolving characteristics of the

FCM and the conditions inside and around the

shelter (e.g. groundwater conditions) create

uncertain safety conditions from the point of view

of protection of workers, the public and the

environment in the future. Therefore, continuation

of the stabilization measures at the shelter and

construction of the NSC are expected to improve

safety and prevent or mitigate accident scenarios

that would be expected to have consequences

outside the CEZ.

It is also required that prompt solutions be

found for the safe predisposal and disposal

management of the radioactive waste to be generated

during this period, in particular for the management of

long lived and high level waste. Planning and

evaluation of safety for the decommissioning of unit 4

after the construction of the NSC is needed in order to

develop appropriate measures and to allocate

necessary resources for the conversion of the shelter

into a safe environmental syste

162

The decommissioning of unit 4 will generate

significant amounts of radioactive waste with a wide

range of characteristics that will need to be safely

managed as part of the decommissioning and waste

management activities at the Chernobyl nuclear

power plant and the CEZ. A comprehensive

strategy for the management of all waste streams is

needed to ensure adequate infrastructure and

capabilities for the processing, storage and disposal

of this waste. Such a strategy will also need to take

into account the future development of

underground and on-surface storage and disposal

facilities, some of which are flooded.

At present, studies show that the known waste

facilities do not present an unacceptable hazard to

the public; however, an assessment of their long

term impact on the public and the environment is

needed. This should be done taking into account the

remaining sources of radioactive contamination in

the CEZ, and particularly those facilities that are

flooded and represent higher risks.

For the less known and less studied waste

facilities it will be necessary to reduce the uncer-

tainties associated with the waste inventories and

facility characteristics, assess their long term safety,

monitor the dynamics of radionuclide migration

into the environment and, where necessary,

implement remediation measures. This is important

for the successful implementation of waste

management activities in the CEZ and the

conversion of the zone into a safe environmental

system.

7.4.2. Recommendations

Recognizing the ongoing effort on improving

safety and addressing the aforementioned uncer-

tainties in the existing input data, the following

main recommendations are made regarding the

dismantling of the shelter and the management of

the radioactive waste generated as a result of the

accident.

(a) Since individual safety and environmental

assessments have been performed only for

individual facilities at and around the

Chernobyl nuclear power plant, a compre-

hensive safety and environmental impact

assessment, in accordance with international

standards and recommendations, that

encompasses all activities inside the entire

CEZ, should be performed.

(b) During the preparation and construction of

the NSC and soil removal, special monitoring

wells are expected to be destroyed. Therefore,

it is important to maintain and improve the

environmental monitoring strategies,

methods, equipment and staff qualification

needed for the adequate performance of

monitoring of the conditions at the Chernobyl

nuclear power plant site and the CEZ.

about 50 years does not seem to be a viable

option, due to the need for long term

maintenance of structure stability and

integrity, resources and knowledge. This long

term strategy raises concerns related to the

potential loss of the most experienced

personnel at the Chernobyl nuclear power

plant and the maintenance of a stable

workforce necessary for the safe operation of

the NSC. It is reasonable, therefore, to begin

retrieving FCM soon after dismantling the

unstable structures of the shelter rather than

waiting for the availability of a geological

disposal facility.

(d) Development of an integrated radioactive

waste management programme for the

shelter, the Chernobyl nuclear power plant

site and the CEZ is needed to ensure

application of consistent management

approaches and sufficient facility capacity for

all waste types. Specific emphasis needs to be

given to the characterization and classification

of waste (in particular waste with transuranic

elements) from all remediation and decom-

missioning activities, as well as to the estab-

lishment of sufficient infrastructure for the

safe long term management of long lived and

high level waste. Therefore, development of

an a

ppropriate waste management infra-

structure is needed in order to ensure

sufficient waste storage capacity; at present,

the rate and continuity of remediation

activities at the Chernobyl nuclear power

plant site and in the CEZ are being limited.

(e) A coherent and comprehensive strategy for

the rehabilitation of the CEZ is needed, with

particular focus on improving the safety of the

existing waste storage and disposal facilities.

This will require development of a prioriti-

zation approach for remediation of the sites,

based on safety assessment results, aimed at

making decisions about those sites at which

waste will be retrieved and disposed of and

163

those sites at which the waste will be allowed

to decay in situ.

REFERENCES TO SECTION 7

[7.1] UNITED NATIONS, Sources and Effects of

Ionizing Radiation (Report to the General

Assembly), Scientific Committee on the Effects of

Atomic Radiation (UNSCEAR), UN, New York

(2000).

[7.2] KASHPAROV, V.A., et al., Territory contamina-

tion with the radionuclides representing the fuel

component of Chernobyl fallout, Sci. Total

Environ. 317 (2003) 105–119.

[7.3] BOROVOY, A., BOGATOV, S., PASUKHIN,

E., Current status of the shelter and its impact on

the environment, Radiokhimiya 41 (1999) 368–378

(in Russian).

[7.4] NOVOKSHCHENOV, V., The Chernobyl

problem, Civ. Eng. 72 (2002) 74–83.

[7.5] BOROVOY, A. (Ed.), Object Shelter Safety

Analysis Report, Chernobyl nuclear power plant

(2002) (in Russian).

[7.6] STATE SPECIALIZED ENTERPRISE ‘CHER-

NOBYL NUCLEAR POWER PLANT’, Envi-

ronmental Impact Assessment, New Safe

Confinement Conceptual Design: Chernobyl

Nuclear Power Plant — Unit 4, State Specialized

Enterprise ‘Chernobyl Nuclear Power Plant’, Kiev

(2003).

[7.7] Radiation Safety Norms of Ukraine, Rep. NRSU-

97, Ministry of Health of Ukraine, Kiev (1998).

[7.8] GMAL, B., MOSER, E.F., PRETZSCH, G.,

QUADE, U., Criticality Behaviour of the Fuel-

containing Masses inside the Object “Shelter” of

the Chernobyl NPP, Unit 4, GRS Rep. A – 2414,

Gesellschaft für Anlagen- und Reaktorsicherheit,

Cologne (1997).

[7.9] OECD NUCLEAR ENERGY AGENCY, Cher-

nobyl: Assessment of Radiological and Health

Impacts, 2002 Update of Chernobyl: Ten Years

On, OECD, Paris (2002).

[7.10] BOGATOV, S., BOROVOY, A., Assessment of

Inventory and Determination of Features of Dust

Contained in the Shelter, Preprint of ISTC Shelter

Report, Chernobyl (2000) (in Russian).

[7.11] PRETZSCH, G., “Radiological consequences of a

hypothetical roof breakdown accident of the

Chernobyl sarcophagus”, One Decade after Cher-

nobyl: Summing up the Consequences of the

Accident, IAEA-TECDOC-964, Vol. 2, IAEA,

Vienna (1997) 591–597.

[7.12] SCIENCE AND TECHNOLOGY CENTER IN

UKRAINE, Comprehensive Risk Assessment of

the Consequences of the Chornobyl Accident,

Science and Technology Center in Ukraine, Kiev

(1998).

[7.13] STATE SPECIALIZED ENTERPRISE ‘CHER-

NOBYL NUCLEAR POWER PLANT’, SIP

Project, Shelter Transformation Safety, Digest

Prepared by the State Specialized Enterprise

‘Chernobyl Nuclear Power Plant’ Information

Department, Slavutych (2004).

[7.14] SCHMIEMAN, E.A., et al., Conceptual design of

the Chernobyl new safe confinement — An

overview, Can. Nucl. Soc. Bull. 25 2 (2004) 9–16.

[7.15] INTERNATIONAL ATOMIC ENERGY

AGENCY, Radiological Conditions in the

Dnieper River Basin, IAEA, Vienna (2006).

[7.16] BUGAI, D.A., WATERS, R.D., DZHEPO, S.P.,

SKALSKY, A.S., Risks from radionuclide

migration to groundwater in the Chernobyl 30-km

zone, Health Phys. 71 (1996) 9–18.

[7.17] SHESTOPALOV, V.M. (Ed.), Chernobyl

Disaster and Groundwater, Balkema, Amsterdam

(2002).

[7.18] VARGO, G.I., et al., The Chernobyl Accident: A

Comprehensive Risk Assessment, Battelle,

Columbus, OH (2000).

[7.19] Integrated Radioactive Waste Programme at the

Stage of Shutdown of the Chernobyl NPP and

Transformation of the Shelter into an Environ-

mentally Safe System, Ministry of Fuel and

Energy, Administration of the Exclusion Zone,

State Committee of Nuclear Regulation, Ministry

of Environment and Natural Resources, and

Ministry of Health of Ukraine (2004) (in

Ukrainian).

[7.20] BONDARENKO, O.O., DROZD, I.P.,

LOBACH, G.O., TOKAREVSKIY, V.V.,

SHIBETSKIY, Y.O., Present problem of radio-

active waste management at the Chernobyl

Exclusion Zone, Bull. Ecol. State Chernobyl

Exclusion Zone 23 (2004) 36–40 (in Ukrainian).

[7.21] Basic Sanitary Rules of Radiation Safety of

Ukraine, Ministry of Health of Ukraine, Kiev

(2000).

[7.22] National Report Developed in Compliance with

the Joint Convention on the Safety of Spent Fuel

Management and on the Safety of Radioactive

Waste Management, Kiev (2003).

[7.23] DEREVETS, V., et al., Radiation conditions in

the Chernobyl exclusion zone in 1999, Bull. Ecol.

State Chernobyl Exclusion Zone 17 (2000) 5–19

(in Ukrainian).

[7.24] INTERNATIONAL ATOMIC ENERGY

AGENCY, Classification of Radioactive Waste,

Safety Series No. 111-G-1.1, IAEA, Vienna

(1994).

[7.25] Concept of Radioactive Waste Management at the

Shelter, Approved by the State Commission on

Issues on Complex Solution of Problems of

Chernobyl NPP on 15 November 1999 (1999) (in

Ukrainian).

[7.26] EUROPEAN COMMISSION, Characterisation

of Radioactive Waste Located at “Shelter” Indus-

trial Site, Rep. EUR 19844, Office for Official

164

Publications of the European Communities,

Luxembourg (2001).

[7.27] PANASUK, N., et al., Results of the assessment of

contaminated soil and underground water at the

shelter site, J. Chernobyl Problems 7 (2002) 97.

[7.28] BOROVOY, A., Object Shelter Safety Analysis

Report (2001).

[7.29] SAVERSKY, S., BUGAY, D., ANTROPOV, V.,

Experience and perspectives of radioactive waste

management at the exclusion zone, Bull. Ecol.

State Chernobyl Exclusion Zone 17 (2001) 26–36

(in Ukrainian).

[7.30] INTERNATIONAL ATOMIC ENERGY

AGENCY, Present and Future Environmental

Impact of the Chernobyl Accident, IAEA-

TECDOC-1240, IAEA, Vienna (2001).

[7.31] Law of Ukraine on Radioactive Waste Manage-

ment of 30 June 1995, Government Courier

(1995).

[7.32] FRENCH–GERMAN INITIATIVE FOR

CHERNOBYL, Final Report on Project No. 2,

Radioecological Consequences of the Chernobyl

Accident, Sub-Project Waste Dumps and Waste

Strategies Management, Deliverable 18, Institute

for Radiological Protection and Nuclear Safety,

Paris, Gesellschaft für Anlagen- und Reaktor-

sicherheit, Cologne, International Chernobyl

Laboratory, Slavutych (2004).

[7.33] SHESTOPALOV, V.M., Some results of the pilot

studies carried out at the Chernobyl exclusion

zone related to the assessment of the radioactive

waste isolation at the deep geological repository,

Bull. Ecol. State Chernobyl Exclusion Zone 21

(2004) 33–36 (in Ukrainian).

[7.34] ANTROPOV, V.M., et al., Review and Analysis of

Solid Long-lived and High Level Radioactive

Waste arising at the Chernobyl Nuclear Power

Plant and the Restricted Zone, DG Environment

Project No. B7-5350/99/51983/MAR/C2, NNC,

Knutsford (2001).

[7.35] LIKHTAREV, I.A., KOVGAN, L.M., BOND-

ARENKO, O.O., If there is future for the

exclusion zone and the population of relocated

territories? (Opinion of a radiologist), Bull. Ecol.

State Chernobyl Exclusion Zone 15 (2000) 44–49

(in Ukrainian).

[7.36] Statement on Policy of Regulation of Nuclear and

Radiation Safety of the Shelter, Approved by the

Ministry of Environmental Protection and

Nuclear Safety, Approval No. 49 of 8 April 1998,

Chernobyl nuclear power plant (1998) (in

Ukrainian).

165

CONTRIBUTORS TO DRAFTING AND REVIEW

Alexakhin, R. Russian Institute of Agricultural Radiology and Agroecology,

Russian Federation

Anspaugh, L. University of Utah, United States of America

Balonov, M. International Atomic Energy Agency

Batandjieva, B. International Atomic Energy Agency

Besnus, F. Institut de radioprotection et de sûreté nucléaire, France

Biesold, H. Gesellschaft für Anlagen- und Reaktorsicherheit, Germany

Bogdevich, I. Belarusian Research Institute of Soil Science and Agrochemistry, Belarus

Byron, D. International Atomic Energy Agency

Carr, Z. World Health Organization

Deville-Cavelin, G. Institut de radioprotection et de sûreté nucléaire, France

Ferris, I. Food and Agriculture Organization of the United Nations

Fesenko, S. Russian Institute of Agricultural Radiology and Agroecology,

Russian Federation

Gentner, N. United Nations Scientific Committee on the Effects of Atomic Radiation

Golikov, V. Institute of Radiation Hygiene of the Ministry of Public Health,

Russian Federation

Gora, A. International Radioecology Laboratory, Ukraine

Hendry, J. International Atomic Energy Agency

Hinton, T. University of Georgia, United States of America

Howard, B. Centre for Ecology and Hydrology, United Kingdom

Kashparov, V. Ukrainian Institute of Agricultural Radiology, Ukraine

Kirchner, G. Institut für Angewandten Strahlenschutz, Germany

LaGuardia, T. TLG Services, Inc., United States of America

Linsley, G. Consultant, United Kingdom

Louvat, D. International Atomic Energy Agency

Moberg, L. Swedish Radiation Protection Authority, Sweden

Napier, B. Pacific Northwest National Laboratory, United States of America

Prister, B. Ukrainian Institute of Agricultural Radiology, Ukraine

Proskura, M. Ministry for Emergencies and Affairs of Population Protection from the

Consequences of the Chernobyl Catastrophe, Ukraine

Reisenweaver, D. International Atomic Energy Agency

166

Schmieman, E. Pacific Northwest National Laboratory, United States of America

Shaw, G. Imperial College of Science, Technology and Medicine, United Kingdom

Shestopalov, V. National Academy of Sciences, Ukraine

Smith, J. Centre for Ecology and Hydrology, United Kingdom

Strand, P. International Union of Radioecology, Norway

Tsaturov, Y. Russian Federal Service for Hydrometeorology and Environmental

Monitoring, Russian Federation

Vojtsekhovich, O. Hydrometeorological Scientific Research Institute, Ukraine

Woodhead, D. Consultant, United Kingdom

Consultants Meetings

Vienna, Austria: 30 June–4 July 2003, 15–19 December 2003, 26–30 January 2004, 14–18 June 2004, 18–22 October 2004,

29 November–3 December 2004, 31 January–4 February 2005

INTERNATIONAL ATOMIC ENERGY AGENCY

VIENNA

ISBN

92–0–114705–8

ISSN 1020–6566

The explosion on 26 April 1986 at the Chernobyl nuclear

power plant and the consequent reactor fire resulted in

an unprecedented release of radioactive material from

a nuclear reactor and adverse consequences for the

public and the environment. Although the accident

occurred nearly two decades ago, controversy still

surrounds the real impact of the disaster. Therefore

the IAEA, in cooperation with the Food and Agriculture

Organization of the United Nations, the United

Nations Development Programme, the United Nations

Environment Programme, the United Nations Office for

the Coordination of Humanitarian Affairs, the United

Nations Scientific Committee on the Effects of Atomic

Radiation, the World Health Organization and the World

Bank, as well as the competent authorities of Belarus,

the Russian Federation and Ukraine, established the

Chernobyl Forum in 2003. The mission of the Forum

was to generate “authoritative consensual statements”

on the environmental consequences and health effects

attributable to radiation exposure arising from the

accident as well as to provide advice on environmental

remediation and special health care programmes, and

to suggest areas in which further research is required.

This report presents the findings and recommendations

of the Chernobyl Forum concerning the environmental

effects of the Chernobyl accident.

Report of the

Chernobyl Forum Expert Group ‘Environment’

Environmental Consequences

of the Chernobyl Accident

nd their Remediation:

wenty Years of Experience

RADIOLOGICAL

AS SES SM E NT

R E P O R T S

S E R I E S

Environmental Consequences of the Chernobyl Accident and their Remediation: Twenty Years of Experience

9.9 mm

180 pages

P1239_covI+IV.indd 1 2006-03-30 14:41:37