The Environmental Behaviour of Polonium

Technical Reports SeriEs No. 484

The Environmental

Behaviour of Polonium

F. Carvalho, S. Fernandes, S. Fesenko, E. Holm, B. Howard,

P. Martin, M. Phaneuf, D. Porcelli, G. Pröhl, J. Twining

The Environmental Behaviour of Poloniumtechnical reportS series no. 484

THE ENVIRONMENTAL

BEHAVIOUR OF POLONIUM

AFGHANISTAN

ALBANIA

ALGERIA

ANGOLA

ANTIGUA AND BARBUDA

ARGENTINA

ARMENIA

AUSTRALIA

AUSTRIA

AZERBAIJAN

BAHAMAS

BAHRAIN

BANGLADESH

BARBADOS

BELARUS

BELGIUM

BELIZE

BENIN

BOLIVIA, PLURINATIONAL

STATE OF

BOSNIA AND HERZEGOVINA

BOTSWANA

BRAZIL

BRUNEI DARUSSALAM

BULGARIA

BURKINA FASO

BURUNDI

CAMBODIA

CAMEROON

CANADA

CENTRAL AFRICAN

REPUBLIC

CHAD

CHILE

CHINA

COLOMBIA

CONGO

COSTA RICA

CÔTE D’IVOIRE

CROATIA

CUBA

CYPRUS

CZECH REPUBLIC

DEMOCRATIC REPUBLIC

OF THE CONGO

DENMARK

DJIBOUTI

DOMINICA

DOMINICAN REPUBLIC

ECUADOR

EGYPT

EL SALVADOR

ERITREA

ESTONIA

ETHIOPIA

FIJI

FINLAND

FRANCE

GABON

GEORGIA

GERMANY

GHANA

GREECE

GUATEMALA

GUYANA

HAITI

HOLY SEE

HONDURAS

HUNGARY

ICELAND

INDIA

INDONESIA

IRAN, ISLAMIC REPUBLIC OF

IRAQ

IRELAND

ISRAEL

ITALY

JAMAICA

JAPAN

JORDAN

KAZAKHSTAN

KENYA

KOREA, REPUBLIC OF

KUWAIT

KYRGYZSTAN

LAO PEOPLE’S DEMOCRATIC

REPUBLIC

LATVIA

LEBANON

LESOTHO

LIBERIA

LIBYA

LIECHTENSTEIN

LITHUANIA

LUXEMBOURG

MADAGASCAR

MALAWI

MALAYSIA

MALI

MALTA

MARSHALL ISLANDS

MAURITANIA

MAURITIUS

MEXICO

MONACO

MONGOLIA

MONTENEGRO

MOROCCO

MOZAMBIQUE

MYANMAR

NAMIBIA

NEPAL

NETHERLANDS

NEW ZEALAND

NICARAGUA

NIGER

NIGERIA

NORWAY

OMAN

PAKISTAN

PALAU

PANAMA

PAPUA NEW GUINEA

PARAGUAY

PERU

PHILIPPINES

POLAND

PORTUGAL

QATAR

REPUBLIC OF MOLDOVA

ROMANIA

RUSSIAN FEDERATION

RWANDA

SAN MARINO

SAUDI ARABIA

SENEGAL

SERBIA

SEYCHELLES

SIERRA LEONE

SINGAPORE

SLOVAKIA

SLOVENIA

SOUTH AFRICA

SPAIN

SRI LANKA

SUDAN

SWAZILAND

SWEDEN

SWITZERLAND

SYRIAN ARAB REPUBLIC

TAJIKISTAN

THAILAND

THE FORMER YUGOSLAV

REPUBLIC OF MACEDONIA

TOGO

TRINIDAD AND TOBAGO

TUNISIA

TURKEY

TURKMENISTAN

UGANDA

UKRAINE

UNITED ARAB EMIRATES

UNITED KINGDOM OF

GREAT BRITAIN AND

NORTHERN IRELAND

UNITED REPUBLIC

OF TANZANIA

UNITED STATES OF AMERICA

URUGUAY

UZBEKISTAN

VANUATU

VENEZUELA, BOLIVARIAN

REPUBLIC OF

VIET NAM

YEMEN

ZAMBIA

ZIMBABWE

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’’.

TECHNICAL REPORTS SERIES No. 484

THE ENVIRONMENTAL

BEHAVIOUR OF POLONIUM

F. CARVALHO, S. FERNANDES, S. FESENKO,

E. HOLM, B. HOWARD, P. MARTIN, M. PHANEUF,

D. PORCELLI, G. PRÖHL, J. TWINING

INTERNATIONAL ATOMIC ENERGY AGENCY

VIENNA, 2017

All IAEA scientific and technical publications are protected by the terms of

the Universal Copyright Convention as adopted in 1952 (Berne) and as revised

in 1972 (Paris). The copyright has since been extended by the World Intellectual

Property Organization (Geneva) to include electronic and virtual intellectual

property. Permission to use whole or parts of texts contained in IAEA publications

in printed or electronic form must be obtained and is usually subject to royalty

agreements. Proposals for non-commercial reproductions and translations are

welcomed and considered on a case-by-case basis. Enquiries should be addressed

to the IAEA Publishing Section at:

Marketing and Sales Unit, Publishing Section

International Atomic Energy Agency

Vienna International Centre

PO Box 100

1400 Vienna, Austria

fax: +43 1 2600 29302

tel.: +43 1 2600 22417

email: [email protected]

http://www.iaea.org/books

Printed by the IAEA in Austria

November 2017

STI/DOC/010/484

IAEA Library Cataloguing in Publication Data

Names: International Atomic Energy Agency.

Title: The environmental behaviour of polonium / International Atomic Energy

Agency.

Description: Vienna : International Atomic Energy Agency, 2017. | Series: Technical

reports series (International Atomic Energy Agency), ISSN 0074–1914 ; no. 484 |

Includes bibliographical references.

Identifiers: IAEAL 17-01115 | ISBN 978–92–0–112116–5 (paperback : alk. paper)

Subjects: LCSH: Radioecology. | Radioactive pollution. | Polonium.

Classification: UDC 539.163 | STI/DOC/010/484

FOREWORD

Polonium-210 is the main contributor to internal doses due to ingestion

of radionuclides from the uranium and thorium decay series. With the expected

increase in uranium mining and other industries generating naturally occurring

radioactive material residues in the future, it is likely that the radiological impact

210

Po will increase in importance.

The IAEA attaches great importance to the dissemination of information

that can assist Member States with the implementation and improvement of

activities relating to radiation safety, including the management of radioactive

residues containing natural radionuclides. This publication outlines the behaviour

of polonium in air, water and soil, and in the human body. The primary objective

is to provide information that can be used in radiological assessments of

accidental releases and routine discharges of polonium to the environment. Case

studies and environmental applications of polonium isotopes are also presented.

The IAEA wishes to express its appreciation of the work of all the

contributors to the drafting and review of this publication. The IAEA officers

responsible for this publication were S. Fesenko and M. Phaneuf of the

IAEA Environment Laboratories.

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.

This publication does not address questions of responsibility, legal or otherwise, for acts

or omissions on the part of any person.

Guidance provided here, describing good practices, represents expert opinion but does

not constitute recommendations made on the basis of a consensus of Member States.

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 specic 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.

The IAEA has no responsibility for the persistence or accuracy of URLs for external or

third party Internet web sites referred to in this book and does not guarantee that any content

on such web sites is, or will remain, accurate or appropriate.

CONTENTS

CHAPTER 1. INTRODUCTION .............................. 1

1.1. Background ......................................... 1

1.2. Objective ........................................... 2

1.3. Scope .............................................. 2

1.4. Structure............................................ 3

References to Chapter 1 .................................... 3

CHAPTER 2. RADIOECOLOGY: BASIC DEFINITIONS AND

CONCEPTS ................................... 5

2.1. Basic parameters ..................................... 5

2.2. Classifications of plants and animals used for exposure

assessment models.................................... 7

2.2.1. Transfer to human foodstuffs ...................... 7

2.2.2. Transfer to wildlife: Reference organism concept ...... 9

References to Chapter 2 .................................... 13

CHAPTER 3. PROPERTIES OF POLONIUM.................... 15

3.1. Physical properties.................................... 15

3.1.1. Isotopes....................................... 15

3.1.2. Decay series systematics.......................... 15

3.2. Chemical properties................................... 22

3.2.1. Solution chemistry .............................. 23

3.2.2. Adsorption .................................... 25

3.2.3. Biovolatilization ................................ 27

3.3.

210

Po determination ................................... 27

References to Chapter 3 .................................... 31

CHAPTER 4. OCCURRENCE AND CYCLING OF

210

Pb AND

210

IN THE ENVIRONMENT

........................ 35

4.1. Occurrence of

210

Pb and

210

Po in rock and soil . . . . . . . . . . . . . . 35

4.2. Occurrence of

210

Pb and

210

Po in the atmosphere ............ 38

4.3. Occurrence of

210

Pb and

210

Po in terrestrial ecosystems ....... 40

4.3.1.

210

Pb and

210

Po in plants .......................... 40

4.3.2.

210

Pb and

210

Po in animals......................... 42

4.4. Occurrence of

210

Pb and

210

Po in marine ecosystems ......... 43

4.4.1.

210

Pb and

210

Po in marine water .................... 43

4.4.2.

210

Pb and

210

Po in marine organisms................. 44

4.5. Occurrence of

210

Pb and

210

Po in freshwater ecosystems ...... 44

4.5.1.

210

Pb and

210

Po in groundwater..................... 44

4.5.2.

210

Pb and

210

Po in surface water and sediments ........ 45

4.5.3.

210

Pb and

210

Po in freshwater organisms.............. 46

4.6. Occurrence of

210

Pb and

210

Po in food (including seafood)

and drinking water

.................................... 46

4.7. Anthropogenic sources of

210

Po in the environment .......... 51

4.7.1. Industrial uses of human-made

210

Po ................ 51

4.7.2. Environmental release of naturally occurring

radioactive material.............................. 52

4.8. Summary ........................................... 54

References to Chapter 4 .................................... 55

CHAPTER 5.

210

Pb AND

210

Po IN ATMOSPHERIC SYSTEMS...... 61

5.1. Introduction to the atmospheric environment ............... 61

5.2. Sources of

210

Pb and

210

Po in the atmosphere ............... 63

5.2.1. Radon exhalation ............................... 63

5.2.2. Volcanic activity ................................ 65

5.2.3. Resuspended soil dust............................ 65

5.2.4. Biomass burning ................................ 66

5.2.5. Wind blown sea spray and emission of volatile

polonium compounds from water surfaces............ 66

5.2.6. Anthropogenic sources ........................... 66

5.3. Processes for the removal of

210

Pb and

210

Po from the

atmosphere

.......................................... 67

5.3.1. Wet deposition ................................. 67

5.3.2. Dry deposition ................................. 69

5.3.3. Deposition fluxes ............................... 69

5.4. Distribution of

210

Pb and

210

Po in the atmosphere............ 70

5.4.1. Concentrations at the land surface .................. 70

5.4.2. Vertical profiles................................. 72

References to Chapter 5 .................................... 73

CHAPTER 6.

210

Pb AND

210

Po IN TERRESTRIAL SYSTEMS ...... 79

6.1. Description of the terrestrial environment.................. 79

6.2. Major transfer processes and pathways.................... 80

6.3. Atmosphere–plant interactions .......................... 81

6.3.1. Interception.................................... 82

6.3.2. Physical and biological weathering ................. 83

6.3.3. Foliar translocation .............................. 83

6.4. Mobility of polonium in soils ........................... 84

6.4.1. Resuspension .................................. 84

6.4.2. Radon exhalation ............................... 85

6.4.3. Partitioning of the soil Po:Pb ratio into solution ....... 85

6.5.

210

Pb and

210

Po transfer from soil to plants ................. 87

6.5.1. Processes governing

210

Pb and

210

Po transfer to

plants

......................................... 87

6.5.2.

210

Pb and

210

Po transfer to mosses, plants and

lichens

........................................ 87

6.5.3.

210

Pb and

210

Po transfer to berries, mushrooms and

understory species

............................... 88

6.5.4.

210

Pb and

210

Po transfer to agricultural plants.......... 93

6.6

210

Pb and

210

Po transfer to animals ....................... 99

6.6.1. Processes governing distribution of

210

Pb and

210

in animals

..................................... 99

6.6.2. Quantification of

210

Pb and

210

Po transfers to

agricultural animals

.............................. 101

6.6.3. Transfer to terrestrial animals...................... 103

References to Chapter 6 .................................... 106

CHAPTER 7.

210

Pb AND

210

Po IN FRESHWATER AND

GROUNDWATER SYSTEMS

..................... 111

7.1. Description of the hydrological cycle ..................... 111

7.2. Groundwater ........................................ 112

7.2.1. Groundwater environment ........................ 112

7.2.2. Sources of

210

Pb and

210

Po in groundwater............ 113

7.2.3. Groundwater concentrations....................... 115

7.2.4. Particles and colloids in groundwater................ 115

7.2.5. In situ adsorption coefficients...................... 117

7.3. Surface water........................................ 121

7.3.1. Sources in surface fresh water ..................... 122

7.3.2. Levels of

210

Po in the freshwater environment ......... 123

7.3.3. Behaviour of

210

Po in standing water bodies .......... 123

7.3.4. Quantification of water and sediment quality.......... 126

7.3.5. Freshwater aquatic biota .......................... 127

References to Chapter 7 .................................... 132

CHAPTER 8.

210

Pb AND

210

Po IN MARINE SYSTEMS............ 137

8.1. Introduction to the marine environment ................... 137

8.2. Polonium sources in the marine environment ............... 137

8.3. Polonium in the oceans ................................ 138

8.4. Accumulation and turnover of

210

Pb and

210

in marine organisms

................................... 142

8.5. Polonium in marine biota living in several ocean regions...... 147

8.5.1. Intertidal zone .................................. 147

8.5.2. Coastal seas.................................... 153

8.5.3. Epipelagic zone................................. 157

8.5.4. Mesopelagic and bathypelagic zones ................ 160

8.5.5. Abyssal zone................................... 161

8.6. Polonium and lead concentration ratios in marine

organisms

........................................... 163

8.7. Transfer of

210

Pb and

210

Po in marine food chains............ 163

References to Chapter 8 .................................... 167

CHAPTER 9. POLONIUM IN HUMANS ....................... 175

9.1. Introduction ......................................... 175

9.2. Ingestion and inhalation................................ 177

9.3. Distribution in tissues ................................. 177

9.4. Retention and biological half-lives ....................... 181

References to Chapter 9 .................................... 184

CHAPTER 10. RADIOLOGICAL DOSE ASSESSMENT FOR

210

TO HUMANS AND OTHER BIOTA

............... 187

10.1. Models and data for estimating internal exposures to

humans

............................................. 187

10.2. Dose conversion factors for wildlife ...................... 188

10.2.1. General assumptions ............................ 188

10.2.2. Dose conversion factors for flora and fauna .......... 193

10.2.3. Non-homogeneous distribution .................... 200

10.2.4. Application of dose conversion coefficients .......... 201

References to Chapter 10 ................................... 202

APPENDIX I: GEOCHRONOLOGICAL APPLICATIONS OF

POLONIUM ................................... 205

APPENDIX II: RADIOCAESIUM,

210

Po AND

210

Pb IN WOLVES .... 209

APPENDIX III:

210

Po IN MARINE MAMMALS ................... 215

APPENDIX IV: POLONIUM RELEASES FROM VEGETATION

AND FOREST FIRES

........................... 227

APPENDIX V: POLONIUM IN TOBACCO AND DOSES TO

HUMANS..................................... 237

APPENDIX VI: ESTIMATION OF AEROSOL RESIDENCE TIMES

IN THE ATMOSPHERE

......................... 241

APPENDIX VII: USE OF

210

Po AS A RADIOTRACER IN

BIOGENIC PARTICULATE FLUX STUDIES

IN THE OCEANS

.............................. 247

ABBREVIATIONS ............................................ 253

CONTRIBUTORS TO DRAFTING AND REVIEW .................. 255

Chapter 1

INTRODUCTION

1.1. BACKGROUND

For many years, the IAEA has supported efforts to develop publications on

the environmental behaviour and use of radionuclides in response to the needs

of its Member States. Since

210

Po is the main contributor to internal doses due to

ingestion of radionuclides from the uranium and thorium decay series

[1.1, 1.2],

it was determined that a resource publication for polonium would be beneficial.

Polonium-210 is an alpha emitting radionuclide with no radioactive progeny and

produces only very-low-intensity gamma rays at very low abundance; thus, the

dose largely arises from internal exposure. The main reasons for its radiological

importance are its relatively high activity concentrations in certain foods and its

relatively high ingestion dose coefficient.

There are seven polonium isotopes naturally present in the environment:

210

Po,

214

Po and

218

Po of the uranium decay series;

212

Po and

216

Po of the thorium

decay series; and

211

Po and

215

Po of the actinium decay series. Polonium-210 has

a half-life of approximately 138

days, which is long enough to play a significant

role in many environmental processes. All of the other naturally occurring

isotopes have half-lives of only 3

minutes or less. The isotopes

214

Po and

218

are short lived progeny of

222

Rn, and are important in dose assessment. However,

they are best examined in relation to the environmental behaviour of radon and

other short lived radon progeny. Hence, the primary focus of this publication is

the environmental behaviour of

210

Po.

Radiation doses from

210

Po arise owing to natural occurrences of the

radionuclide as well as to human activities. As

210

Po is part of the uranium decay

series, it is naturally occurring and is found in varying amounts worldwide. The

anthropogenic sources include uranium mining and milling activities, as well

as the production of naturally occurring radioactive material (NORM), such as

phosphogypsum, and oil and gas scales. With uranium mining and other industries

generating NORM residues expected to grow in the future, it is likely that the

radiological impact of

210

Po will be of increasing importance. The challenges

faced have been addressed in a number of IAEA publications on radionuclide

transfer in the environment, which focus on radionuclide transfers in terrestrial,

freshwater and marine environments, and provide information on key transfer

processes, concepts and models important in radiological assessments for all

radionuclides (see Refs [1.3–1.8]).

CHAPTER 1

Since 2004, the IAEA has also organized a series of projects aimed at

improving environmental assessment and remediation. Through these projects,

the environmental behaviour of radionuclides has been presented in other IAEA

publications, which cover the following topics:

— Contaminated site characterization [1.9–1.11];

— Waste discharge, surveillance and management [1.12–1.14];

— Environmental remediation [1.15, 1.16];

— Remediation of uranium mill tailings [1.17];

— Remediation of dispersed contamination [1.18];

— Radiation protection and the management of radioactive waste in the oil

and gas industry

[1.19].

1.2. OBJECTIVE

This publication provides information on the environmental behaviour

of polonium, which can be used in the radiological assessment of polonium

occurring naturally in the environment or present due to routine discharges and

accidental releases. This publication can support environmental investigations,

assessments, and remediation in areas contaminated by polonium, and presents

concepts, models and parameters describing the behaviour of polonium.

1.3. SCOPE

This publication explores the behaviour of polonium in atmospheric,

terrestrial, freshwater and marine environments. The primary focus is the

environmental behaviour of

210

Po, the only polonium isotope with a half-life long

enough to play a significant role in environmental processes but, at 138

days,

also short enough that ingrowth from its immediate progenitors

210

Bi and

210

can be important when studying polonium’s behaviour and for inferring kinetics

and mechanisms of the processes. Information on the progenitors of

210

Po — in

particular

210

Pb — is also presented to capture

210

Po behaviour in the environment.

In addition to presenting information on the environmental transfer of

radionuclides to humans and non-human biota, this publication provides further

information on the assessment of the impact of radioactive discharges (e.g. see

Refs [1.12, 1.20]. Applications of

210

Po as a tracer of environmental processes

are discussed, but industrial, military and other applications of

210

Po, such as

occupational dose assessment, are outside the scope of this publication.

INTRODUCTION

1.4. STRUCTURE

Chapter 2 introduces basic definitions and concepts for radioecology.

The physical, chemical and decay properties of polonium are provided in

Chapter

3, and Chapter 4 addresses sources and cycling of

210

Pb and

210

Po in

the environment. Chapters

5–8 explore the behaviour of

210

Pb and

210

Po in

the atmospheric, terrestrial, freshwater (including groundwater) and marine

environments, respectively. Chapter

9 describes the metabolism of polonium

in the human body. Radiological dose assessment models for humans and

wildlife relating to the occurrence of polonium in the environment are given in

Chapter

10. Appendices I–VII consider various case studies and environmental

applications of polonium.

REFERENCES TO CHAPTER 1

[1.1] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Sources and Effects of Ionizing Radiation, UNSCEAR 2000 Report to

the General Assembly, with Scientific Annexes, Vol. I: Sources, United Nations,

New York (2000) Annex B.

[1.2] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Sources and Effects of Ionizing Radiation, UNSCEAR 2008 Report to

the General Assembly, with Scientific Annexes, Vol. II: Effects, United Nations,

New York (2008) Annex E.

[1.3] INTERNATIONAL ATOMIC ENERGY AGENCY, The Environmental Behaviour of

Radium: Revised Edition, Technical Reports Series No.

476, IAEA, Vienna (2014).

[1.4] INTERNATIONAL ATOMIC ENERGY AGENCY, Measurement and Calculation of

Radon Releases from NORM Residues, Technical Reports Series No.

474, IAEA,

Vienna

(2013).

[1.5] INTERNATIONAL ATOMIC ENERGY AGENCY, Sediment Distribution Coefficients

and Concentration Factors for Biota in the Marine Environment, Technical Reports

Series No.

422, IAEA, Vienna (2004).

[1.6] INTERNATIONAL ATOMIC ENERGY AGENCY, Quantification of Radionuclide

Transfer in Terrestrial and Freshwater Environments for Radiological Assessments,

IAEA-TECDOC-1616, IAEA, Vienna

(2009).

[1.7] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer in Terrestrial and Freshwater Environments,

Technical Reports Series No.

472, IAEA, Vienna (2010).

[1.8] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer to Wildlife, Technical Reports Series

No.

479, IAEA, Vienna (2014).

CHAPTER 1

[1.9] INTERNATIONAL ATOMIC ENERGY AGENCY, Characterization of Radioactively

Contaminated Sites for Remediation Purposes, IAEA-TECDOC-1017, IAEA,

Vienna

(1998).

[1.10] INTERNATIONAL ATOMIC ENERGY AGENCY, Site Characterization Techniques

Used in Environmental Restoration Activities, IAEA-TECDOC-1148, IAEA,

Vienna

(2000).

[1.11] INTERNATIONAL ATOMIC ENERGY AGENCY, Strategy and Methodology for

Radioactive Waste Characterization, IAEA-TECDOC-1537, IAEA, Vienna (2007).

[1.12] INTERNATIONAL ATOMIC ENERGY AGENCY, Regulatory Control of Radioactive

Discharges to the Environment, IAEA Safety Standards Series No.

WS-G-2.3, IAEA,

Vienna

(2000).

[1.13] INTERNATIONAL ATOMIC ENERGY AGENCY, Management of Radioactive

Waste from the Mining and Milling of Ores, IAEA Safety Standards Series

No.

WS-G-1.2, IAEA, Vienna (2002).

[1.14] INTERNATIONAL ATOMIC ENERGY AGENCY, Monitoring and Surveillance of

Residues from the Mining and Milling of Uranium and Thorium, Safety Reports Series

No.

27, IAEA, Vienna (2002).

[1.15] INTERNATIONAL ATOMIC ENERGY AGENCY, Factors for Formulating Strategies

for Environmental Restoration, IAEA-TECDOC-1032, IAEA, Vienna

(1998).

[1.16] INTERNATIONAL ATOMIC ENERGY AGENCY, Technologies for Remediation of

Radioactively Contaminated Sites, IAEA-TECDOC-1086, IAEA, Vienna

(1999).

[1.17] INTERNATIONAL ATOMIC ENERGY AGENCY, The Long Term Stabilization of

Uranium Mill Tailings, IAEA-TECDOC-1403, IAEA, Vienna

(2004).

[1.18] INTERNATIONAL ATOMIC ENERGY AGENCY, Remediation of Sites with

Dispersed Radioactive Contamination, Technical Reports Series No.

424, IAEA,

Vienna

(2004).

[1.19] INTERNATIONAL ATOMIC ENERGY AGENCY, Radiation Protection and the

Management of Radioactive Waste in the Oil and Gas Industry, Safety Reports Series

No.

34, IAEA, Vienna (2003).

[1.20] INTERNATIONAL ATOMIC ENERGY AGENCY, Generic Models for Use in

Assessing the Impact of Discharges of Radioactive Substances to the Environment,

Safety Reports Series No.

19, IAEA, Vienna (2001).

Chapter 2

RADIOECOLOGY: BASIC DEFINITIONS AND CONCEPTS

Several generic parameters and concepts are used throughout this

publication to characterize the environmental behaviour of polonium in various

media. They are defined here, generally according to the recommendations of

the International Commission on Radiation Units and Measurements

[2.1], and

follow IAEA publications on radionuclide transfers in different environments

(see Refs

[2.2, 2.3]). This publication has the advantage that the relatively

large quantity of data for

210

Pb and

210

Po, covering a wide range of species and

ecosystems, is typically described using these concepts.

2.1. BASIC PARAMETERS

There are several parameters which are commonly used to describe

radionuclide partitioning between environmental compartments. These include

transfer to plants or animals from soil, water or within the food chain, namely

concentration ratios (CRs) or transfer coefficients (e.g. B

, F

and F

), and

between sediments or soil minerals and water, using distribution coefficients (K

These parameters are defined for equilibrium conditions, and so it is assumed

that an organism or particle is in equilibrium with its surroundings in terms of

its accumulation of the radionuclide of interest. The definitions of other, more

specific parameters are given in Ref.

[2.2]. The parameters described here are

also regularly applied within existing predictive models to estimate the transfer

of radionuclides in the environment and the resulting exposures for humans and

non-human biota.

For terrestrial plants or wildlife, the primary indicator used to characterize

radionuclide behaviour in the soil–plant system is the soil–plant CR, also often

referred to as the transfer factor (TF), and it describes the transfer of radionuclides

from the soil to the plant when uptake by plant roots is the only process affecting

the transfer. The CR is defined as the ratio of the activity concentration of

radionuclide in the plant or animal (Bq/kg, DW) and in soil (Bq/kg, DW), based

on the dry weight (DW):

( )

Activity concentration of radionuclide in edible tissue Bq/kg, DW

Activity concentration of radionuclide in soil Bq/kg, DW

(2.1)

CHAPTER 2

The soil can be measured to a defined depth (e.g. 10 cm or 20 cm) and can

refer to soil contamination per m

as an aggregated concentration ratio (CR

agg

particularly for materials deposited via atmospheric transfer (see Chapter 5 for

a more detailed discussion). The CRs for plants are usually given for the edible

parts of the plant.

A similar ratio is applied for aquatic organisms with activity concentration

in water (Bq/L) or sediment (Bq/kg), and often on a fresh weight (FW) basis. The

transfer of radionuclides from water to aquatic organisms includes the intake of

water, sediments and feed, so the assumption is often that all those processes are

aggregated.

The CR for any organism can sometimes be expressed as a bioavailable

ratio (CR

bioav

). This acknowledges that some of any environmental radionuclide

(

210

Po in this case) will be in a chemical or physical form not amenable for

biological uptake. In those cases, geochemical speciation modelling (e.g. based

on pH, Eh, dissolved organics or other water chemistry) or physical processing

(e.g. filtration to remove particulates) will be used to estimate the bioavailable

portion in the soil or water and this, in turn, can be used to estimate the CR

bioav

For estimating radionuclide transfer from feeds to domestic animal food

products, a transfer coefficient is widely used for both milk (F

) and meat (F

This coefficient is defined as the equilibrium ratio of the radionuclide activity

concentration in the milk/meat on a FW basis to the daily dietary radionuclide

intake:

( )

m/f

Activity concentration in milk/meat Bq/kg, FW

Daily radionuclide intake Bq/d

F =

(2.2)

CRs for human food chains and biota have some differences. For human

food chains, CRs are always ratios of the radionuclide activity concentrations

in the edible parts of the organisms to that in the surroundings. However, for

non-human species, the CR is calculated for the whole organism. This difference

reflects the purpose in the application of these parameters: in the first case, the

CRs are intended for assessments of radionuclide activity concentrations in food

that may be ingested by humans and are used to provide a radiological dose;

whereas the whole organism CRs are intended for dose assessments of the biota.

Considering abiotic compartments, K

is one of the basic parameters used

to characterize the mobility of radionuclides in the environment. The degree to

which

210

Po is bound to solids (soils in the terrestrial environment and sediments

in aquatic systems) is an important factor for determining the concentration of

a radionuclide in solution, and it directly influences the fraction of radionuclide

that can be incorporated by organisms (see above regarding bioavailability).

Dissolved radionuclide ions can bind to solid surfaces by a number of processes

RADIOECOLOGY

that are often classified under the broad term of ‘sorption’. Polonium is

particularly affected by sorption owing to its low environmental concentrations

and its variable oxidation state (see Chapter

3). K

for the partitioning of a

radionuclide between the particulate phase and the dissolved phase under

equilibrium conditions is similar to the aquatic CR described above and is

defined as:

( )

Activity concentration in solid phase Bq/kg, DW

Activity concentration in aqueous phase Bq/L

K =

(2.3)

All of the above approaches have some limitations owing to the inherent

complexity and variability of natural environmental conditions. In particular,

the assumption of equilibrium being established in ecosystems is debatable.

Hence, the application of CRs for radiological conditions undergoing substantial

temporal variations in radionuclide activity concentrations in the environmental

media, for example following any accidental release, might be inappropriate.

Dynamic models are widely used for quantifying transfers of

210

Po in

non-equilibrium situations (see Ref. [2.3]), although the data coverage to populate

these models is not always sufficient for robust assessments.

2.2. CLASSIFICATIONS OF PLANTS AND ANIMALS USED FOR

EXPOSURE ASSESSMENT MODELS

There are many data in the scientific literature on radionuclide transfer to

plants and animals. The data are given for many different species, for whole or

parts of plants and animals, as well as for different environmental conditions.

However, to suit the purpose of dose assessment using predictive models such

as PC-CREAM

[2.4] or the ERICA Tool [2.5], the data need to represent some

widely recognized, standardized conditions. Such classifications for both human

and wildlife food chains can be found in IAEA publications and are discussed

here (see Refs [2.2, 2.6, 2.7]).

2.2.1. Transfer to human foodstuffs

The IAEA plant classification system is based on just 14 plant groups

[2.2].

All plants are categorized as cereals, maize, rice, leafy vegetables, non-leafy

vegetables, leguminous vegetables, root crops, tubers, fruits, grasses (cultivated

species), leguminous fodder (cultivated species), pasture (species mixture —

natural or cultivated), herbs and other crops. This system has been proposed

CHAPTER 2

as a basis for estimating the transfer of radionuclides to plant foodstuffs in the

framework of the assessment of exposures to humans through ingestion.

Plant tissues are subdivided into ten compartments: berries, buds, fruits,

grains, heads, leaves, roots, seeds and pods, stems and shoots, and tubers. Not all

of these ten compartments are assigned to each plant group, but only where the

portion represents an edible part of a specific plant.

The transfer of radionuclides to plants greatly depends on soil properties.

Existing international soil classification systems try to capture important

information for plant cultivation. The soil classification system established by the

Food and Agriculture Organization of the United Nations (FAO) and the United

Nations Educational, Scientific and Cultural Organization (UNESCO) has

units (or categories) of soil and 125 subunits [2.8]. TF values are not available

for subunits defined on such a detailed basis: the differences between these units

in terms of radionuclide transfer are generally not substantial. Therefore, a much

simpler classification system based on texture and organic matter content was

suggested in Ref.

[2.3], while ensuring that a reasonable amount of data are

available for each category. Four soil groups — sand, loam, clay and organic

soil — are defined for radiological assessments and are mostly based on grain

size (clay

≤ 0.004 mm, sand ≥ 0.06–2 mm, gravel > 2 mm and loams being

intermediate). The soils are grouped according to the percentage of sand and clay

mineral and the organic matter content in the soil

[2.3]:

“For the mineral soils, three groups were created according to the sand

and clay percentages…: ‘Sand group’: sand fraction ≥65%; clay fraction

<18%; ‘Clay group’: clay fraction ≥35%; ‘Loam group’: rest of cases. A

soils was included in the ‘Organic group’ if the organic matter content was

≥20%. Finally, an ‘Unspecified soil group’ was created for soils without

characterization data, or for mineral soils with unknown sand and clay

contents.”

With regard to animals, there are only five categories (cattle, sheep, goats,

pigs and poultry) and three groups (meat, milk and eggs) that are considered

to provide transfer coefficients from daily radionuclide intake of animals and

animal products consumed by humans

[2.3]. The classification systems are

assumed adequate to cover the variability of environmental conditions and can

also be applied for the assessment of the transfer of polonium through the human

consumption of food. A detailed discussion of the data and concepts used to

define these categories of soil, plants and animals are given in Refs [2.2, 2.3],

and many empirical data on

210

Po (and

210

Pb, where appropriate) presented in this

publication use these parameters.

RADIOECOLOGY

2.2.2. Transfer to wildlife: Reference organism concept

Since it is impossible to consider all species of plants and animals in

radiological impact assessments, a classification based on a reference organisms

is applied. Larsson [2.9] defines reference organism as “a series of entities that

provide a basis for the estimation of radiation dose rate to a range of organisms

which are typical, or representative, of a contaminated environment”. Such

groups need to be selected on the basis of their representativeness for the selected

environments of interest and should also allow realistic assessments, illustrating

possible exposure pathways. This approach is used in the ERICA project, which

develops a framework for radiation protection of the environment (see Ref.

[2.5]).

A similar concept has been applied by the International Commission on

Radiological Protection (ICRP), which introduced a system of discrete and

clearly defined reference animals and plants for assessing radiation effects on

wildlife. According to the ICRP definition [2.10]:

“A Reference Animal or Plant is a hypothetical entity, with the assumed

basic characteristics of a specific type of animal or plant, as described to

the generality of the taxonomic level of Family, with defined anatomical,

physiological and life-history properties, that can be used for the purposes

of relating exposure to dose, and dose to effects, for that type of living

organism.”

The ICRP approach is based on the consideration of 12 more or less

globally representative reference animals and plants, covering different life stages

(e.g. fish egg, adult fish). This is a basis for systematically relating radionuclide

exposure to radiological dose, and then dose (or dose rate) to different types of

effect, for a number of organisms that are characteristic of different types of

natural environment [2.10].

In 2009, the IAEA, within the international programme Environmental

Modelling for Radiation Safety (EMRAS

II), initiated a working group to

develop an international handbook for estimating the transfer of radionuclides

to wildlife, similar to Ref.

[2.2] for estimating transfer to human foodstuffs.

The approach adopted for classification of wildlife was based on a reference

organism concept

[2.6, 2.7] and was consistent with the ICRP approach [2.11].

However, it was applied more generally, and it defines broader wildlife groups

(e.g. soil invertebrate, predatory fish and terrestrial mammal; see Table

2.1). In

some cases, a consideration of specific subcategories is included. This approach

is also used in this publication, along with the relevant data for

210

Pb and

210

collected within the EMRAS

II programme to harmonize environmental impact

assessments.

CHAPTER 2

TABLE 2.1. WILDLIFE GROUPS AND REFERENCE ORGANISMS

Broad group Subcategory Reference organism

Terrestrial environments

Amphibians —

Frog

Annelids —

Earthworm

Arachnids —

—

Arthropods

Carnivorous

Detritivorous

Herbivorous

—

Bee

Birds

Carnivorous

Herbivorous

Omnivorous

Duck

Ferns —

—

Fungi

Mycorrhizal

Parasitic

Saprophytic

—

Gastropods —

—

Grasses and herbs

Grasses

Herbs

Wild grass

—

Lichens/bryophytes —

—

Mammals

Carnivorous

Herbivorous

Marsupial

Omnivorous

Rangifer spp.

Rat or deer

Rat

Rat or deer

—

Rat

—

Reptiles

Carnivorous

Herbivorous

—

Shrubs —

—

Trees

Coniferous

Broadleaf

Pine tree

—

RADIOECOLOGY

TABLE 2.1. WILDLIFE GROUPS AND REFERENCE ORGANISMS (cont.)

Broad group Subcategory Reference organism

Freshwater environments

Algae —

—

Amphibians —

Frog

Birds

Carnivorous

Herbivorous

Omnivorous

Duck

Crustaceans —

—

Fish

Benthic feeding

Piscivorous

Forage

—

Salmonid

—

Insects —

—

Insect larvae

—

Mammals

Carnivorous

Herbivorous

Omnivorous

—

Molluscs

Bivalve

Gastropod

—

Phytoplankton —

—

Reptiles —

—

Vascular plants —

Wild grass

Zooplankton —

—

Marine environments

Birds

Carnivorous

Herbivorous

Omnivorous

Duck

Crustaceans

Large

Small

Crab

—

CHAPTER 2

TABLE 2.1. WILDLIFE GROUPS AND REFERENCE ORGANISMS (cont.)

Broad group Subcategory Reference organism

Fish

Benthic feeding

Piscivorous

Forage

Flatfish

Salmonid

—

Insects —

—

Macroalgae —

Brown seaweed

Mammals

Carnivorous

Herbivorous

Planktivorous

—

Molluscs

Bivalve

Cephalopod

Gastropod

—

Phytoplankton —

—

Polychaete worms —

—

Reptiles —

—

Sea anemones/true corals —

—

Vascular plants —

—

Zooplankton —

—

Source: Tables 2–4 of Ref. [2.6]. The list is based on an on-line database available at

www.wildlifetransferdatabase.org

—: data not available.

Herb refers to any non-woody plant which does not fall into one of the other categories.

Does not include Rangifer spp. (reindeer and caribou).

All marsupials, regardless of feeding strategy.

Fish feeding on benthic dwelling organisms.

Fish consuming smaller fish, amphibians or birds.

Fish feeding on primary producers and pelagic invertebrates and zooplankton.

Insect larvae are included as the aquatic life phase is important for many species which are

terrestrial as an adult.

Squid, octopus and cuttlefish.

RADIOECOLOGY

REFERENCES TO CHAPTER 2

[2.1] INTERNATIONAL COMMISSION ON RADIATION UNITS AND

MEASUREMENTS, Quantities, Units and Terms in Radioecology, ICRU Report

65,

ICRU, Bethesda, MD (2001).

[2.2] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer in Terrestrial and Freshwater Environments,

Technical Reports Series No.

472, IAEA, Vienna (2010).

[2.3] INTERNATIONAL ATOMIC ENERGY AGENCY, Quantification of Radionuclide

Transfer in Terrestrial and Freshwater Environments for Radiological Assessments,

IAEA-TECDOC-1616, IAEA, Vienna

(2009).

[2.4] SMITH, J.G., SIMMONDS, J.R. (Eds), The Methodology for Assessing the

Radiological Consequences of Routine Releases of Radionuclides to the Environment

Used in PC-CREAM

08, HPA-RPD-058, Health Protection Agency, Chilton,

Didcot

(2009).

[2.5] BROWN, J.E., et al., The ERICA Tool, J. Environ. Radioact. 99 (2008) 1371–1383.

[2.6] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer to Wildlife, Technical Reports Series

No.

479, IAEA, Vienna (2014).

[2.7] HOWARD B.J., et al., The IAEA handbook on radionuclide transfer to wildlife,

J. Environ. Radioact. 121 (2013) 55–74.

[2.8] FOOD AND AGRICULTURE ORGANIZATION OF THE UNITED NATIONS,

UNITED NATIONS EDUCATIONAL, SCIENTIFIC AND CULTURAL

ORGANIZATION, Soil Map of the World, 1:5 000 000, UNESCO, Paris (1994).

[2.9] LARSSON, C.-M., The FASSET Framework for assessment of environmental impact

of ionising radiation in European ecosystems: An overview, J. Radiol. Prot.

(2004)

A1–A12.

[2.10] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION, The

2007 Recommendations of the International Commission on Radiological Protection,

Publication

103, Elsevier, Amsterdam (2007).

[2.11] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Environmental Protection: The Concept and Use of Reference Animals and Plants,

Publication

108, Elsevier, Amsterdam (2008).

Chapter 3

PROPERTIES OF POLONIUM

This chapter summarizes the physical and chemical properties of polonium

and acts as a reference for the subsequent chapters. It includes decay data

and series systematics, solution chemistry and adsorption, and methods of

determining

210

Po activity concentrations in samples.

3.1. PHYSICAL PROPERTIES

3.1.1. Isotopes

There are 42 known isotopes of polonium in the National Nuclear Data

Center Chart of Nuclides

and none is stable. There are seven naturally occurring

radioactive isotopes of polonium, which are part of the natural decay series of

long lived radionuclides (see Table 3.1 and Fig. 3.1):

— The

238

U series:

210

Po,

214

Po and

218

Po.

— The

235

U series:

211

Po and

215

Po.

— The

232

Th series:

212

Po and

216

Po.

At 138 days, however, only

210

Po has a half-life long enough to play a

significant role in environmental processes. The decay scheme for

210

Po, along

with those of

210

Bi and

210

Pb, which precede

210

Po in the

238

U decay series, is

shown in Fig. 3.2. There are several anthropogenic isotopes with a half-life

longer than a day, which are generally produced by irradiation of bismuth

(see Table 3.1). The longest lived,

209

Po (T

1/2

= 102 a), has been used for

polonium chemistry experiments and as a yield tracer of polonium during sample

processing and radiochemical analyses.

3.1.2. Decay series systematics

The distribution of

210

Po in the environment is to be considered in the

context of

238

U decay series systematics. The concentration of each radionuclide

is controlled by the concentration of its parent, its own half-life and the amount

See www.nndc.bnl.gov/chart

CHAPTER 3

TABLE 3.1. POLONIUM ISOTOPES

Isotope Half-life

Decay constant

−1

)

Major decay modes

(>1%)

Specific activity

(Bq/g)

Energy of decay

(MeV)

(with % of total decays)

Emission energy

(MeV)

Naturally

occurring

210

Po 138.376 d 1.829 73 α 1.66 × 10

5.407 (>99.9%) 5.304

211

Po 0.516 s 4.24 × 10

α 3.83 × 10

7.594 (98.9%)

7.024 (0.54%)

6.695 (0.52%)

7.450

6.891

6.568

212

Po 0.300 μs 7.29 × 10

α 6.56 × 10

8.954 (100%) 8.785

214

Po 162 μs 1.35 × 10

α 1.20 × 10

7.833 (99.99%)

7.034 (0.01%)

7.687

6.903

215

Po 1.781 ms 1.228 × 10

α 1.09 × 10

7.526 (99.93%)

7.087 (0.06%)

7.386

6.955

216

Po 0.148 s 1.48 × 10

α 1.31× 10

6.906 (>99.99%) 6.778

218

Po 3.07 m 1.19 × 10

α 1.04 × 10

6.115 (99.98%) 6.002

Anthropogenic

1/2

> 1 d)

206

Po 8.8 d 29 α (5.45%)

ε (94.55%)

2.7 × 10

5.327 (5.45%)

1.846 (94.55%)

5.224

208

Po 2.898 a 0.239 2 α 2.19 × 10

5.215 (>99.99%) 5.115

209

Po 115 a 6.03 × 10

−3

α 5.50 × 10

4.716 (0.55%)

4.977 (79.2%)

4.979 (19.8%)

4.622

4.883

4.885

Source: Data are from the Decay Data Evaluation Project (www.nucleide.org/DDEP_WG/DDEPdata.htm), where available, otherwise (for

206

and

208

Po) from the National Nuclear Data Center Chart of Nuclides (www.nndc.bnl.gov/chart).

Note: Transformations with intensities less than 0.01% have not been included in this table. Intensities are given as a percentage of total decay.

Where more complete data are required, the original sources should be consulted.

PROPERTIES

Note: The only long lived polonium isotope,

210

Po, is circled.

FIG. 3.1. The naturally occurring

238

235

U and

232

Th decay series, all of which contain

polonium isotopes (Z = 84).

CHAPTER 3

of time since fractionation between the isotope and its parent occurred. For the

decay series elements within samples that have been closed over long timescales,

the activity of each isotope (decay rate) is the same as that of its parent (i.e. it is

in secular equilibrium with its parent). For the

238

U decay series, which contains

210

Po, this can be represented by:

(

238

U) = (

226

Ra) = (

222

Rn) = … = (

210

Pb) = (

210

Po) (3.1)

where the parentheses denote activity, which is equal to the number of atoms

multiplied by the decay constant.

There are two important points to note:

— In secular equilibrium, the activity concentration (and, from that, the molar

concentration) of a daughter radionuclide is controlled by that of the parent

radionuclide, and so, ultimately, all radionuclides in the series are controlled

by the longest lived parent (here

238

U).

— While the activities are equal, the molar abundances are inversely

proportional to the decay constants, so that the molar concentrations of

very short lived nuclides are very low.

The molar ratio of

210

Po:

238

U is 8.5 × 10

−11

, while the molar ratio of

210

Po:

226

Ra is 2.4 × 10

−4

. Therefore, even where uranium or radium is greatly

enriched, polonium has an extremely low molar concentration — even when it

exhibits high radioactivity. This has important implications for understanding its

behaviour in the environment.

When any series is not in secular equilibrium due to chemical or physical

fractionation, the activity of the nuclides will tend to evolve back towards secular

equilibrium over a timescale determined by the half-life of the daughter isotope.

The section of the

238

U decay series that is most relevant to understanding

210

FIG. 3.2. The decay scheme for

210

Bi,

210

Pb and

210

Po.

PROPERTIES

is below

226

Ra. Including only those isotopes with half-lives longer than one day,

this part of the series is:

1602 a 3.82 d 22.3 a 5.01 d 138.4 d

226 222 210 210 210 206

Ra Rn Pb Bi Po Pb¾¾¾® ¾¾¾® ¾¾¾® ¾¾¾® ¾¾¾¾®

(3.2)

In secular equilibrium, the activities of all these radionuclides are the same.

However, the following environmental processes can separate them:

(a) The isotope

222

Rn, a noble gas, can migrate. It is generally unreactive and

forms no chemical bonds within environmental materials. It is sometimes

released from solids by a process known as alpha recoil (see below) and,

despite its relatively low aqueous solubility, it can enter any surrounding

water (e.g. pore water, groundwater and lake water) or vapour phase

(e.g. pore spaces, bubbles and the atmosphere), thereby separating from its

226

Ra parent.

(b) The radionuclides

226

Ra,

222

Rn and

210

Pb, which are produced by alpha

decay, and the very short lived radionuclides

218

Po and

214

Pb, which are

between

222

Rn and

210

Pb, are propelled in the direction opposite to that of

the alpha particle and with sufficient energy to travel approximately 20

in mineral structures. They can therefore escape to the surrounding water

or air or come to rest in an adjacent phase

[3.1, 3.2]. The direction of recoil

is random for each decay, and the rate of ejection is determined only by

the fraction of parent nuclides within recoil distance of mineral surfaces or

channels to the surface. This is sometimes referred to as alpha recoil.

wildfires (see Appendix IV) and industrial processes, and then condense or

scavenge during subsequent cooling.

(d) There can be contrasting adsorption behaviour of lead, polonium and

radium on particles in the atmosphere, and on surfaces in soils and aquifers,

and into colloids or particulates in surface waters.

(e) Contrasting uptake, excretion and biodistribution of lead, polonium and

radium can occur in biota.

The systematics of

210

Po under these varying circumstances are shown

in Fig. 3.3. In each case, following disequilibrium, the activity of the daughter

nuclide grows into secular equilibrium with the parent over a timescale relating to

its half-life. To understand the evolution of

210

Po and the resulting radiation, the

evolution of

210

Pb and

210

Bi also needs to be considered. When

210

Po is isolated

from the other isotopes in the decay series, such as owing to the migration of

222

Rn and subsequent decay to

210

Po, it will decay unsupported. If it is then not

CHAPTER 3

Note: (A)

210

Po is unsupported by decay of the parent

210

Bi (and so by decay of

210

Pb), for

example when

210

Po is separated from

210

Pb and

210

Bi by precipitation or volatilization.

(B) There is

210

Pb present, and the evolution of

210

Po is shown for several different

starting activities. The excess or deficiency of

210

Po will decay according to the

half-life of

210

Po. Such situations are found in both biological systems and abiotic

environments, where

210

Po can be fractionated from

210

Pb to generate either an excess

or a deficiency of

210

Po. The inset shows the ingrowth of

210

Bi, which occurs over

a much shorter time period, and so the amount of initial

210

Bi has little effect on the

overall evolution of

210

Po. (C) The ingrowth of

210

Pb is shown in the presence of

226

Ra,

for example within barite (BaSO

), which strongly incorporates radium. The different

timescale for this circumstance should be noted. Over this time period,

226

Ra is almost

constant. The activity of

210

Po will closely follow that of

210

Pb.

FIG. 3.3. The closed-system evolution of

210

Po under various circumstances.

PROPERTIES

subject to further additions, and losses only by decay, its activity will evolve over

time according to:

( ) ( )

210

210 210

Po Po e

tl-

(3.3)

where the activity (

210

Po) decays from the initial value (

210

Po)

according to the

decay constant

210

. The evolution is shown in Fig. 3.3, where the activity will

drop to 1% of the initial value in 6.6 half-lives or about 2.5 years.

When

210

Pb and

210

Po are separated from the rest of the decay series, the

activity of

210

Pb will evolve according to an equation analogous to Eq. (3.3),

although it will decay with a half-life of 22.3 years. Its daughter,

210

Bi, will grow

according to:

( ) ( )

( )

210

210 210

Bi Pb

Bi Pb e e

(3.4)

Over the time period in which the activity of

210

Bi approaches that of

210

(i.e. to within 1% in 33 days), the activity of

210

Pb drops by only 0.3% and so can

be considered constant:

( ) ( )

( )( )

210 210

210

210 210 210 210

210 210

210

210 210 210 210

210 210

210

210 210 210 210

Bi Po

Bi Pb Po Pb

210 210 210

Bi Po

Pb Bi Po Bi

Bi Po

Pb Po Bi Po

Po Pb e P

l ll l

llll

l ll l

éù

êú

=+ +

êú

ëû

( )

210

tl-

(3.5)

Substituting values for the decay constants gives:

( ) ( ) ( )

210

210 210 210

1.0179e

Po Pb 0.0376e Po e

1.0555e

éù

êú

=+ +

êú

ëû

(3.6)

Regardless of the initial activity of

210

Po relative to that of

210

Pb, the activity

210

Po will evolve to within 1% of that of

210

Pb in 2.5 years. Over this time, the

CHAPTER 3

activity of

210

Pb drops by only 7.5% (i.e. the term drops to 0.925, and so for much

of the time can be approximated to 1).

3.2. CHEMICAL PROPERTIES

While in some environments the activity concentration of

210

Po can be

of concern, the molar concentrations in the environment are always extremely

low. In the long term, the distribution of

210

Po is controlled by that of the parent,

210

Pb. The behaviour of

210

Po in aqueous systems is generally dominated by

adsorption onto surfaces, although incorporation into colloids, biovolatilization

and precipitation in sulphides can be important in some circumstances.

There have been several reviews on polonium chemistry and its behaviour

in the environment

[3.3, 3.4]. However, the data have been limited and some

reasons for this include the following:

(a) The very high specific activity of

210

Po (1.66 × 10

Bq/g; see Table 3.1)

causes practical difficulties in the safe handling of even small amounts of

natural polonium. The specific activities of artificially produced

208

Po and

209

Po are lower but still high.

(b) The very high specific activity also readily damages crystal lattices and

materials in experiments, so it is difficult to obtain undamaged materials

(e.g. precipitated minerals) for characterization.

210

Po is very scarce. One gram of

238

U is generally

accompanied by 75 pg of

210

Po. Hence, in crustal rocks with an average

concentration of 2.7 ppm uranium, there is only 0.20 ng of

210

Po per

tonne. Production of

208

Po and

209

Po requires bombardment of bismuth by

neutrons in a reactor.

(d) In the environment,

210

Po is present at such low concentrations that it clearly

does not form separate phases (see Sections

3.2.1–3.2.3 and Chapter 6 for

more details). For example, groundwater typically has less than 40 mBq/L,

with a maximum value of 19 Bq/L (see Section 4.5.1), which corresponds to

typical groundwater concentrations of less than 0.001 fM and a maximum

of only 0.45 fM.

(e) In many circumstances,

210

Po will decay or approach secular equilibrium

with its parent,

210

Pb. Its distribution is therefore strictly controlled by

210

Pb.

PROPERTIES

3.2.1. Solution chemistry

Polonium is in Group 16 of the periodic table, which is known as the

chalcogens and includes oxygen, sulphur, selenium and tellurium (see Fig. 3.4).

It is generally considered to be a metalloid, with physical and chemical

characteristics that are intermediate between those of metals and non-metals

(along with boron, silicon, germanium, arsenic, antimony, tellurium and astatine),

although it can behave more like the metals bismuth and lead (a review of the

early analytical chemistry can be found in Refs [3.5–3.7]). Based on analogy with

selenium and tellurium, Figgins [3.6] predicts that polonium would have stable

oxidation states of −2, +2, +4 and +6, with the +4 oxidation state being the stable

state in solution under oxic freshwater conditions; although Po

has also been

predicted to be most stable in sea water and under reducing conditions [3.8, 3.9].

Polonium hydrolyses, forming PoO(OH)

, PoO(OH)

and PoO

in slightly acidic

to neutral pH regions and PoO

−3

in alkaline solutions [3.10].

There are limited solubility data available for polonium, but it is generally

assumed to be highly insoluble. Work was conducted on the precipitation

of polonium from laboratory solutions in the 1950s, in particular during the

development of separation procedures, and polonium was found to precipitate

with various compounds of antimony, bismuth, tellurium and rare earth

elements [3.5]. The solubility product for Po(OH)

is 10

−37

[3.6]; hence, in

natural waters, equilibrium activities are highly sensitive to pH [3.11]. In acidic

solutions, trace polonium is precipitated by hydrogen sulphide (H

S), along

with other insoluble sulphides. Under these conditions, polonium is also highly

insoluble, sinwith ce the solubility product for PoS at around 5 × 10

−29

[3.6].

The concentration of dissolved polonium can therefore be strictly controlled by

sulphur recycling, where the dissolution and reprecipitation of sulphur minerals

occurs [3.12]. However, away from very dynamic conditions, adsorption onto

particles, mineral surfaces and colloids provides the stronger control on dissolved

polonium concentrations.

Polonium is present in precipitates containing radium owing to ingrowth

after mineral formation. Radium is commonly precipitated in barite (BaSO

Since polonium is not readily accommodated within a crystal structure like

radium, it might be released during recystallization, although laboratory

experiments indicate that recrystallization does not expel a significant fraction

of polonium [3.13]. In a study of the bacterial mobilization of polonium from

waste gypsum, LaRock et al. [3.14] demonstrate that sulphate reducing bacteria

are effective at mediating polonium release from gypsum, provided the sulphide

levels resulting from their metabolism does not rise above 10 μmol/L, in which

case polonium is evidently co-precipitated as a metal sulphide.

CHAPTER 3

FIG. 3.4. The location in the periodic table of polonium and the elements with isotopes within the

238

U decay series.

PROPERTIES

Polonium forms colloids in solution [3.15], and while this may be a

significant factor when there are high concentrations of polonium, or in pure

solutions in laboratory experiments (see Ref. [3.10]), it is more likely that in

natural environments polonium is adsorbed onto, or incorporated into, colloids

formed by more abundant elements.

3.2.2. Adsorption

In most natural environments, polonium is highly surface reactive, and

readily adsorbs onto mineral surfaces, particulates and colloids. However, there

are limited data quantifying this. Models for the description of radionuclide

sorption are still mostly based on empirical solid–liquid distribution coefficient

) values of bulk materials (see Sections 2.1 and 6.4.3 for more information on

). This approach is the simplest sorption model available and K

is the ratio of

the concentration of radionuclide in a specified solution to the concentration of

the radionuclide sorbed on a specified solid when the solution and solid are at

equilibrium, with units L/kg. However, these values cannot be easily extrapolated

to materials with different compositions.

There are only limited experiments to determine K

for polonium on

homogenous materials. Laboratory experiments on montmorillonite clay found

measured adsorption values of around 1.5 × 10

L/kg for

210

Po, which were

independent of both ionic strength (up to 0.2 eq/L) and pH (6 and 7) [3.10].

However, polonium that was adsorbed from solution could not be readily

desorbed, with a several orders of magnitude greater proportion of polonium

remaining adsorbed. It was suggested that polonium forms stable covalent surface

complexes. It was also found that

210

Po generated by radioactivity of sorbed

210

can be more readily desorbed, although a distribution value much higher than the

adsorption experiments, of 2 × 10

L/kg, was still obtained. For other materials,

batch partitioning experiments found average values of the following [3.6]:

— 8.4 × 10

L/kg for bentonite at pH10.1;

— 3.7 × 10

L/kg for a tuff (mainly plagioclase, smectite and the zeolite

clinoptilolite) at pH9.4;

— 2.5 × 10

L/kg

at pH10.1 for a granodiorite (largely quartz, plagioclase and

K-feldspar).

These values were obtained by ultrafiltering the solution, and somewhat

lower values were obtained by using 0.45 μm filters. Presumably, a significant

fraction of polonium was not in true solution but rather on colloid sized

particles, and this should be considered when comparing the results of different

studies. Manganese (IV) oxides have often been inferred to significantly

CHAPTER 3

adsorb or incorporate substantial amounts of polonium in field studies (see

Refs [3.16–3.18]). Pyrolusite (MnO

) has been qualitatively shown to effectively

adsorb polonium in column experiments [3.19]. However, adsorption coefficients

onto manganese minerals has not been quantified in the laboratory.

Some results for partitioning experiments with soils have been published.

An early study [3.20] still provides the basis for most estimates of polonium

adsorption in soil, and is the basis for recent recommended values [3.21, 3.22].

Polonium partitioning has also been determined in aquatic systems,

including rivers, estuaries, coastal seas and the open ocean. In these environments,

210

Po consistently displays K

values of approximately 10

, and usually slightly

higher than the K

of the parent radionuclide,

210

Pb [3.23, 3.24]. Values for a

range of natural environments, determined from measurements of filtered waters

and particulates from filters, are shown in Table 3.2.

TABLE 3.2. AQUATIC K

IN SITU DETERMINATIONS

Environment

210

Po K

210

Pb K

Ref.

River 5 (1.0 ± 0.3) × 10

(3 ± 1) × 10

[3.23]

Estuary 7 (3.3 ± 1.5) × 10

(2.4 ± 1.9) × 10

[3.24]

Coastal sea, shoreline 17 (2.8 ± 0.9) × 10

(1.7 ± 1.0) × 10

[3.24]

Surface, open ocean 4 (1–4) × 10

(1–4) × 10

[3.23]

In rivers,

210

Po is mostly associated with suspended particulate matter and

bottom sediments. In estuaries,

210

Po and lead partitioning may be controlled by

the concentration of dissolved organic matter and other ligands, the nature and

concentration of particulate matter, and by changes in salinity [3.23]. In coastal

sea water,

210

Po is mostly associated with biological particles (bacteria and

microplankton), while

210

Pb is mostly associated with the inorganic particulate

fraction (clay minerals). The enhanced incorporation of

210

Po in particulate

organic material, and also in the first levels of trophic marine chains, has been

linked to metabolic processes and

210

Po binding to proteins [3.25–3.28]. The

same occurs in open ocean waters and the degree of planktonic uptake in the

euphotic zone of oceans may have a strong influence on

210

Po flux in oceanic

water columns [3.29, 3.30].

Overall, the K

values that are available can be used as guidelines, but there

is little systematic data exploring the effects of various conditions on polonium

partitioning. Furthermore, K

values are generally values for the relative

PROPERTIES

concentrations between phases and not partition coefficients for equilibrium

adsorption. Further work is required to determine when polonium incorporation

is reversible, and how it varies with polonium concentration.

3.2.3. Biovolatilization

Hussain et al. [3.31] find that organo-polonium compounds may comprise a

substantial proportion of environmental polonium, based on experiments in which

up to 50% of polonium is lost from Florida groundwater with high polonium

concentrations by bubbling nitrogen. Furthermore, this effect is substantially

reduced when oxidizing agents are added, suggesting that volatile organic species

have been destroyed. By analogy with the behaviour of tellurium, Hussain et

al. [3.31] suggest that dimethyl polonium was present. A similar experiment

with groundwater from Nevada did not find significant polonium volatilization,

indicating that the extent of organic polonium complexation may vary [3.18].

In addition to the implications for volatile loss of polonium to the atmosphere,

these results suggest that there may be organic ligands that can control polonium

adsorption behaviour in soil and groundwater.

Momoshima et al. [3.32] demonstrate that polonium is lost by volatilization

in culture experiments in which microorganism growth has been enhanced

in seawater and coastal sediments, and suggest that methylation of polonium

is responsible. In a further, they show that when methylcobalamin, which is

involved in organic methylation, is added, polonium is also lost. Similar losses

are found from similar culture experiments using fresh waters and increase with

enhanced bacterial growth [3.33]. Volatilization has also been demonstrated in

different pure bacterial cultures [3.34].

3.3.

210

Po DETERMINATION

Only a short summary of

210

Po determination in environmental samples

is presented here, since there are several studies on the topic, including

a comprehensive review

[3.35] and a procedure for analysis of water

samples

[3.36–3.38]. Owing to the volatility of polonium, high temperature

sample preparation and digestion techniques, such as dry ashing and fusion with

molten salts, cannot be used. To avoid polonium losses, drying should be by

freeze drying or oven drying at a relatively low temperature (≤80°C).

Wet ashing is commonly used for sample dissolution, sometimes in

combination with microwave digestion. Soils, sediments and other solid

samples, such as filtered materials, are usually prepared using HCl, HF, HNO

and HClO

in varying proportions, with ashing in open vessels, pressure vessels

CHAPTER 3

or microwave digestion systems. Biological samples are generally treated by

first drying them thoroughly at temperatures of up to 80°C or freeze drying,

followed by treatment with HNO

(sometimes also with HCl and H

) to

destroy organic matter. Polonium is pre-concentrated from water samples by

a wide variety of techniques, with the most common involving one of three

procedures: evaporation; co-precipitation, typically on Fe(OH)

or MnO

; or

chelation, commonly with ammonium pyrrolidine dithiocarbamate (APDC)

species

[3.35, 3.39].

210

Po is an alpha emitting radionuclide with a relatively short half-life

and produces only very-low-intensity gamma rays and no radioactive progeny,

the only practical way to determine its presence is to detect its alpha particle.

The most commonly used method is alpha particle spectrometry, which

utilizes a silicon surface barrier or passivated implanted planar silicon (PIPS)

detectors, on the account of their excellent energy resolution, compact size, low

background, excellent stability and low sensitivity to gamma radiation. Alpha

spectrometry can also be performed using gridded ion chambers, which have a

high efficiency and allow the use of large sources. Alternatively, alpha particle

counting may be performed using a ZnS scintillator screen and photomultiplier

tube, gas proportional counter or liquid scintillation counter. This approach has

the disadvantage that a radioisotopic tracer cannot be used, although some liquid

scintillation systems do enable separation of the

209

Po peak.

Whichever detection method is used,

210

Po has first to be chemically

separated from both the sample matrix and from other alpha emitting

radionuclides present in the sample. Polonium separation methods include

those based on solvent extraction, ion exchange chromatography and extraction

chromatography

[3.35]. In addition, polonium spontaneously deposits from mild

acid solutions onto metal surfaces, in particular silver, and this is very often

utilized as a combined chemical separation and source preparation step

[3.40].

The interference of iron during spontaneous deposition is suppressed by the

addition of ascorbic acid or hydroxylamine hydrochloride (to reduce Fe

). Citric

acid may also be added to suppress the effects of other ions present. Metals other

than silver have been used for plating of polonium, including copper and nickel.

However, analyses using these metals are more prone to interference as some

210

Bi or

210

Pb may also be deposited, resulting in an overestimation of

210

Po when

there is a delay between deposition and counting

[3.41, 3.42].

An advantage of the alpha spectrometry detection methods is that an

isotopic yield tracer can be used to allow for losses during sample preparation,

chemical separation and source preparation steps. Historically,

208

Po has been

the preferred tracer, probably on the account of its greater availability. The

advantage of

209

Po over

208

Po is that its E

is further separated from that of

210

(see Table

3.1), allowing better peak resolution. In the case of non-spectrometric

PROPERTIES

methods, such as total alpha counting using a ZnS scintillator screen or liquid

scintillation counting, the recovery for

210

Po during the separation and source

preparation process needs to be estimated.

Contamination of any alpha spectrometry detector by the tracer (

208

Po or

209

Po) or by

210

Po is a potentially serious problem. This contamination can be

minimized by a period of exposure of the source to air before counting, possibly

due to the formation of a surface oxide film on the silver

[3.5, 3.43]. A delay

period of at least two days has been suggested

[3.39].

A polonium spectrum with

209

Po used as the yield tracer is shown in

Fig.

3.5. The spectrum is a simple one, consisting essentially of two singlet

peaks. The spectrum obtained using

208

Po as a tracer is similar, but with a smaller

separation between the two peaks. As described in Ref. [3.35]:

“Once the count rates in the tracer and

210

Po peaks have been obtained

and tailing allowed for, the contributions due to detector background and

procedure blank need to be subtracted. The ratio of the net count rates in

the two peaks is then used to calculate the activity concentration of

210

in the sample on the date of chemical separation of Po, taking into account

the activity concentration of the tracer solution, the masses of the sample

and the tracer solution used, the decay of

210

Po between separation and

counting, the decay of the tracer between its calibration date and counting,

and the

α emission probabilities in the measured areas in the α-spectrum”.

Finally, the results need to be corrected for decay and for ingrowth from its

progenitors (especially

210

Pb) between the sampling date and the date of chemical

FIG. 3.5. A typical polonium alpha particle spectrum with

209

Po as the yield tracer.

CHAPTER 3

separation of polonium from the sample for analysis. This calculation needs to be

carried out separately from the decay correction to separation date because, in the

former, the

210

Po is supported by

210

Pb in the sample, whereas in the latter it is

unsupported (see Fig.

3.6).

In order to make the correction for

210

Po to the sample collection date, the

210

Pb activity concentration needs to be determined. The details of the ingrowth

and decay calculation are dependent on the procedure followed

[3.35]:

“One common approach is that when Po is chemically separated, the solution

containing the

210

Pb is retained and kept for an ingrowth period (usually

several months). A subsequent determination of

210

Po from this solution is

used to calculate the activity concentration of

210

Pb in the sample”.

An alternative approach is to use a separate determination of

210

Pb in the

sample by another method, such as gamma spectrometry or beta counting. In

general, the time delay between sample collection and

210

Po determination should

be kept as short as practicable to avoid errors due to this decay and ingrowth

correction, particularly for sample types for which the

210

Po:

210

Pb activity ratio

may be low (e.g. for rainwater) [3.37].

FIG. 3.6. Decay and ingrowth corrections for

210

Po determination, taking into account the

time intervals t

(separation date − sampling date) and t

(count date − separation date).

PROPERTIES

REFERENCES TO CHAPTER 3

[3.1] KIGOSHI, K., Alpha-recoil thorium-234: Dissolution into water and the uranium-234/

uranium-238 disequilibrium in nature, Science

173 (1971) 47–48.

[3.2] FLEISCHER, R.L., RAABE, O.G., Recoiling alpha-emitting nuclei: Mechanisms for

uranium-series disequilibrium, Geochim. Cosmochim. Acta

42 (1978) 973–978.

[3.3] COPPIN, F., ROUSSEL-DEBET, S., Comportement du

210

Po en milieu terrestre: revue

bibliographique, Radioprot. 39 (2004) 39–58.

[3.4] PERSSON, B.R.R., HOLM, E., Polonium-210 and lead-210 in the terrestrial

environment: A historical review, J. Environ. Radioact.

102 (2011) 420–429.

[3.5] BAGNALL, K.W., The chemistry of polonium, Q. Rev. Chem. Soc. 11 (1957) 30–48.

[3.6] FIGGINS, P.E., The Radiochemistry of Polonium, NAS-NS 3037, National Academy

of Sciences, Washington, DC (1961).

[3.7] BAGNALL, K.W., The chemistry of polonium, Radiochim. Acta 32 (1983) 153–161.

[3.8] BROOKINS, D.G., Eh-pH Diagrams for Geochemistry, Springer-Verlag, Berlin (1988).

[3.9] SWARZENSKI, P.W., McKEE, B.A., SØRESEN, K., TODD, J.F.,

210

Pb and

210

Po,

manganese and iron cycling acoss the O

S interface of a permanently anoxic fjord:

Framvaren, Norway, Mar. Chem.

67 (1999) 199–217.

[3.10] ULRICH, H.J., DEGUELDRE, C., The sorption of

210

Pb,

210

Bi, and

210

Po on

montmorillonit: A study with emphasis on reversibility aspects and on the effect of the

radioactive decay of adsorbed nuclides, Radiochim. Acta

62 (1993) 81–90.

[3.11] UPCHURCH, S.B., OURAL, C.R., FOSS, D.W., BROOKER, H.R., Radiochemistry

of Uranium-Series Isotopes in Groundwater, Technical Report

PB-98-126253/XAB,

Univ. South Florida, Tampa, FL

(1991).

[3.12] SEILER, R.L., STILLINGS, L.L., CUTLER, N., SALONEN, L., OUTOLA, I.,

Biogeochemical factors affecting the presence of

210

Po in groundwater, Appl.

Geochem.

26 (2011) 526–539.

[3.13] BERWANGER, D.J., DEJONG, G.I., WILES, D.R., The release of polonium-210

upon the recrystallization of Ba(Ra)SO

, Water, Air, Soil Pollut. 27 (1986) 77–84.

[3.14] LaROCK, P., HYUN, J.-H., BOUTELLE, S., BURNETT, W.C., HULL, C.D., Bacterial

mobilization of polonium, Geochim. Cosmochim. Acta

60 (1996) 4321–4328.

[3.15] MORROW, P.E., DELLA ROSA, R.J., CASARETT, L.J., MILLER, G.J.,

Investigations of the colloidal properties of polonium-210 solutions by using molecular

filters, Radiat. Res. Suppl.

5 (1964) 1–15.

[3.16] BACON, M.P., BREWER, P.G., SPENCER, D.W., MURRAY, J.W., GODDARD, J.,

Lead-210, polonium-210, manganese and iron in the Cariaco Trench, Deep Sea Res.

Part

A 27 (1980) 119–135.

[3.17] BALISTRIERI, L.S., MURRAY, J.W., PAUL, B., The geochemical cycling of stable

Pb,

210

Pb, and

210

Po in seasonally anoxic Lake Sammamish, Washington, USA,

Geochim. Cosmochim. Acta

59 (1995) 4845–4861.

[3.18] SEILER, R.L.,

210

Po in Nevada groundwater and its relation to gross alpha radioactivity,

Ground Water

49 (2011) 160–171.

[3.19] MARKOSE, P.M., Adsorption of natural radionuclides on pyrolusite, Water, Air, Soil

Pollut.

35 (1987) 381–386.

CHAPTER 3

[3.20] HANSEN, W.R., WATTERS, R.L., Unsupported

210

PoO

in soil: Soil adsorption and

characterization of soil solution species, Soil Sci.

112 (1971) 145–155.

[3.21] VANDENHOVE, H., GIL-GARCIA, C., RIGOL, A., VIDAL, M., New best estimates

for radionuclide solid–liquid distribution coefficients in soils — Part

2: Naturally

occurring radionuclides, J. Environ. Radioact.

100 (2009) 697–703.

[3.22] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer in Terrestrial and Freshwater Environments,

Technical Reports Series No.

472, IAEA, Vienna (2010).

[3.23] CARVALHO, F.P., Distribution, cycling and mean residence time of

226

Ra,

210

Pb and

210

Po in the Tagus estuary, Sci. Total Environ. 196 (1997) 151–161.

[3.24] CARVALHO, F.P., OLIVEIRA, J.M., ALBERTO, G., Factors affecting

210

Po and

210

activity concentrations in mussels and implications for environmental bio-monitoring

programmes, J. Environ. Radioact.

102 (2011) 128–137.

[3.25] DURAND, J.P., et al.,

210

Po binding to metallothioneins and ferritin in the liver of

teleost marine fish, Mar. Ecol. Prog. Ser.

177 (1999) 189–196.

[3.26] STEWART, G.M., FISHER, N.S., Experimental studies on the accumulation of

polonium-210 by marine phytoplankton, Limnol. Oceanogr. 48 (2003) 1193–1201.

[3.27] WILDGUST, M.A., McDONALD, P., WHITE, K.N., Temporal changes of

210

Po in

temperate coastal waters, Sci. Total Environ.

214 (1998) 1–10.

[3.28] WILDGUST, M.A., McDONALD, P., WHITE, K.N., Assimilation of

210

Po by the

mussel Mytilus edulis from the alga Isochrysis galbana, Mar. Biol.

136 (2000) 49–53.

[3.29] STEWART, G., et al., Exploring the connection between

210

Po and organic matter in

the northwestern Mediterranean, Deep Sea Res. Part

I 54 (2007) 415–427.

[3.30] JEFFREE, R.A., SZYMCZAK, R., Enhancing effect of marine oligotrophy on

environmental concentrations of particle-reactive trace elements, Environ. Sci.

Technol.

34 (2000) 1966–1969.

[3.31] HUSSAIN, N., FERDELMAN, T.G., CHURCH, T.M., LUTHER G.W., III,

Bio-volatilization of polonium: Results from laboratory analyses, Aquat. Geochem.

(1995)

175–188.

[3.32] MOMOSHIMA, N., SONG, L.-X., OSAKI, S., MAEDA, Y., Formation and emission

of volatile polonium compound by microbial activity and polonium methylation with

methylcobalamin, Environ. Sci. Tech.

35 (2001) 2956–2960.

[3.33] MOMOSHIMA, N., SONG, L.-X., OSAKI, S., MAEDA, Y., Biologically induced Po

emission from fresh water, J. Environ. Radioact.

63 (2002) 187–197.

[3.34] MOMOSHIMA, N., FUKUDA, A., ISHIDA, A., YOSHINAGA, C., Impact of

microorganism on polonium volatilization, J. Radioanal. Nucl. Chem.

272

(2007)

413–417.

[3.35] MATTHEWS, K.M., KIM, C.-K., MARTIN, P., Determination of

210

Po in

environmental materials: A review of analytical methodology, Appl. Radiat. Isot.

(2007)

267–279.

[3.36] INTERNATIONAL ATOMIC ENERGY AGENCY, A Procedure for the Determination

of Po-210 in Water Samples by Alpha Spectrometry, IAEA Analytical Quality in

Nuclear Applications Series No.

12, IAEA, Vienna (2009).

PROPERTIES

[3.37] KIM, C.-K., MARTIN, P., FAJGELJ, A., Quantification of measurement uncertainty in

the sequential determination of

210

Pb and

210

Po by liquid scintillation counting and

alpha-particle spectrometry, Accredit. Qual. Assur.

13 (2008) 691–702.

[3.38] KIM, C.-K., LEE, M.H., MARTIN, P., Method validation of a procedure for

determination of

210

Po in water using DDTC solvent extraction and Sr resin, J.

Radioanal. Nucl. Chem.

279 (2009) 639–646.

[3.39] MARTIN, P., HANCOCK, G.J., Routine Analysis of Naturally Occurring

Radionuclides in Environmental Samples by Alpha-Particle Spectrometry, Supervising

Scientist Report 180, Supervising Scientist, Darwin (2004).

[3.40] FLYNN, W.W., The determination of low levels of polonium-210 in the environmental

materials, Anal. Chim. Acta

43 (1968) 221–227.

[3.41] EHINGER, S.C., PACER, R.A., ROMINES, F.L., Separation of the radioelements

210

Pb–

210

Bi–

210

Po by spontaneous deposition onto noble metals and verification by

Cherenkov and liquid scintillation counting, J. Radioanal. Nucl. Chem.

(1986) 39–48.

[3.42] HENRICSSON, F., RANEBO, Y., HOLM, E., ROOS, P., Aspects on the analysis of

210

Po, J. Environ. Radioact. 102 (2011) 415–419.

[3.43] BAGNALL, K.W., The Chemistry of Selenium, Tellurium and Polonium, Elsevier,

Amsterdam

(1966).

Chapter 4

OCCURRENCE AND CYCLING OF

210

Pb AND

210

IN THE ENVIRONMENT

This chapter provides a general overview of the occurrence, behaviour and

cycling of

210

Pb and

210

Po in the environment and covers some major concepts

and pathways (see Fig.

4.1). The data presented give a feeling for the orders of

magnitude of activity concentrations which can be expected in environmental

media and the relative importance of various pathways. However, the

information is not comprehensive, nor does it provide details of the mechanisms

involved. Environmental media and pathways are discussed in more detail in

Chapters 5–10, including data on activity concentrations, transfer parameters and

fluxes.

4.1. OCCURRENCE OF

210

Pb AND

210

Po IN ROCK AND SOIL

The source of virtually all

210

Pb and

210

Po in the environment is

238

U in the

Earth’s crust. In unweathered rocks,

210

Pb and

210

Po are expected to be in secular

equilibrium with

238

U. A measure of typical natural concentrations is given by the

composition of the upper continental crust (see Table 4.1), which has an average

concentration of uranium of 2.7 ppm or 33 Bq/kg [4.2].

TABLE 4.1. TYPICAL URANIUM CONCENTRATIONS

IN VARIOUS ROCKS

Rock type U (ppm)

238

U (Bq/kg)

Basalts 0.1–1 1–12

Black shale 3–1 250 37–15 000

Gneiss 2 25

Limestone 2 25

Phosphates 50–300 620–3 700

Schist 2–5 25–62

Silicic rocks (granites, dacites) 2.2–6.1 27–75

Source: See Ref. [4.1].

CHAPTER 4

FIG. 4.1. Occurrence and cycling of

210

Pb and

210

Po in the environment.

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

Three major processes which transfers

210

Pb and

210

Po out of primary

sources in the crust include the following:

(a) Weathering of exposed rocks at the Earth’s surface;

(b) Leaching or dissolution of rocks by groundwater;

Surface weathering and subsequent geomorphological and geochemical

processes result in the formation of soils and sediments. These processes can

modify the concentration of uranium and separate the radionuclides within

the uranium series decay chain. The result can be disequilibrium between the

radionuclides (see Section 3.1.2). Surface soils are the interface between the

atmosphere and the subsurface hydrological system, and there are radionuclide

fluxes between these environmental compartments. Secular equilibrium for the

238

U decay chain of radionuclides in soils cannot be assumed, although the degree

of disequilibrium is not large in the majority of cases.

One mechanism which causes disequilibrium in the uranium decay chain

is exhalation of

222

Rn from surface soils to the atmosphere. This subsequently

decays to

210

Pb, which is then deposited to the Earth’s surface. Exhalation of

222

Rn results in a deficiency of

210

Pb relative to

226

Ra in the topsoil (primarily

the top 1 m), and deposition results in an excess. Approximately 90% of

210

deposition is by wet deposition and greatly depends on the rainfall [4.3, 4.4]. A

confounding factor is

210

Pb in leaf litter, which can lead to excess

210

Pb in the

surface layer [4.5, 4.6]. Since exhalation flux densities of

222

Rn from soils are

about a hundred times greater than those from oceans, which cover more than

70% of the planet,

210

Pb to oceans will result in deficiency of

210

Pb relative to

226

Ra in soils.

Polonium-210 is at, or very close to, secular equilibrium with

210

Pb in

almost all soils [4.7]. Its concentrations in surface soil vary considerably, and

Table 4.2 presents data from studies over the last forty years [4.8]. In areas of

normal radiation background, the concentrations most commonly lie within the

range of 10–200 Bq/kg in the absence of anthropogenic influences [4.7, 4.23].

The upper ranges are mainly due to areas of uranium mining.

The concentrations of

210

Po in surface soils vary across a wide range,

depending on the uranium concentration in the bedrock of the area, the depth

of the soil sampled, climatic conditions and soil properties

[4.10, 4.15, 4.22]. In

addition, phosphate fertilizers commonly contain significant concentrations of

210

Pb and

210

Po [4.24, 4.25].

CHAPTER 4

4.2. OCCURRENCE OF

210

Pb AND

210

Po IN THE ATMOSPHERE

The primary source of

210

Pb in the atmosphere is

222

Rn exhalation from soil

and its subsequent decay via short lived radon progeny. This has been estimated

to give rise to a

210

Pb flux to the atmosphere of 3.5 × 10

Bq/a, while fluxes

from other sources are more than one order of magnitude lower [4.4]. Much of

this

210

Pb is removed from the atmosphere by wet and dry deposition, but that

fraction which does not deposit decays via

210

Bi to

210

Po, giving a flux from this

source of 1.05 × 10

Bq/a.

Volcanic eruptions are a major source of

210

Po in the atmosphere due to

the volatility of polonium at elevated temperatures. An early study estimates

the average flux to be 2.4 × 10

Bq/a for

210

Po, but only 6.0 × 10

Bq/a for

210

Pb [4.26]. Compared to

222

Rn exhalation from soil, volcanic eruptions are a

highly variable source, with relatively large year to year variations.

Other sources include

222

Rn exhalation from the oceans (see Section 5.2.1),

resuspension of soils (see Section 5.4.1), forest and savanna fires (see

TABLE 4.2. CONCENTRATIONS OF

210

Po IN TOPSOIL

Origin

210

(Bq/kg, DW)

Ref.

Brazil 27–74 [4.9

]

Canada 20–22 000 [4.7, 4.10

]

Germany 11–210 [4.11–4.13

]

India 4–220 [4.14–4.16]

Malaysia 8 [4.17]

Spain 16–780 [4.18

]

United Kingdom 51 [4.19]

USA 10–60

70–15 000

[4.14, 4.20]

[4.21

]

Worldwide 8–220 [4.22]

Source:

See Ref. [4.8].

Near a uranium mine.

Near a coal plant.

Near a phosphate factory.

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

Section IV.2) and the combustion of fossil fuels. Globally, these contribute less

than 10% of the total fluxes of

210

Pb and

210

Po in the atmosphere [4.4], although

they are sometimes important on a local or regional scale.

Both

210

Pb and

210

Po are attached to aerosol particles in the atmosphere.

Radioactive decay is only a minor contributor to their removal from the

atmosphere due to their relatively short residence times. The primary removal

pathway is wet deposition to the Earth’s surface (see Section 5.3). A secondary

pathway is dry deposition to the surface, which accounts for about 10–15%

of total deposition for

210

Pb [4.3, 4.4]. The annual, total deposition flux of

210

Pb ranges from several Bq·m

−2

·a

−1

(as observed in the Antarctic) to several

hundred Bq·m

−2

·a

−1

The distribution of

210

Pb and

210

Po in the atmosphere is complex due to the

spatial and temporal variability in the sources, in wet and dry deposition fluxes,

and in atmospheric dispersion processes. The highest

210

Pb concentrations in

surface air are observed in the subtropical and temperate latitudes of the northern

hemisphere owing to the relatively large land masses there. Average annual

concentrations in this region mainly lie in the range of 0.2–1 mBq/m

[4.27].

Between the tropics, values are generally in the range of 0.1–0.5 mBq/m

while south of 30°S, they quickly trend down to less than 0.1 mBq/m

, which

is consistent with the radon exhalation flux density trends discussed above.

These data are predominantly for continental locations, where higher surface air

concentrations are on the account of greater

222

Rn exhalation fluxes from soils

than oceans.

There is also clear evidence of a seasonality effect in

210

Pb and

210

concentrations in surface air [4.28–4.31]. Concentrations are normally higher

during dry seasons due to higher exhalation of

222

Rn from the soil, greater

resuspension of soil dust and lower wet deposition rates, which can vary locally.

The vertical distribution of

210

Pb in the troposphere is highly variable.

Although mean vertical profiles of

222

Rn generally show a strong reduction in

concentration from the surface to the upper troposphere of a factor of 100 over

continental areas, for

210

Pb this reduction is generally less than a factor of 5,

and the concentration at the upper troposphere can be higher than in the lower

troposphere [4.32, 4.33]. The reason for this is scavenging of

210

Pb by wet and

dry deposition in the lower troposphere and by precipitation from the mid-layers.

The mean residence time for

210

Pb in the troposphere is about five days [4.28],

but it is significantly longer than this in the upper troposphere, resulting in higher

210

Po:

210

Pb ratios than in the lower and mid-level troposphere.

The geographical distributions of

210

Pb and

210

Po concentrations in

rainwater mostly follow those in surface air. The main factors governing

radionuclide concentrations in rainwater are the size of the aerosols containing

210

Pb and

210

Po, and the type and duration of rainfall [4.34]. Values for

210

CHAPTER 4

concentration in rainwater vary considerably at different locations, being most

commonly in the range of 10–1000

mBq/m

with

210

Po:

210

Pb activity ratios in the

range of 0.05–0.5.

In the stratosphere,

210

Pb and

210

Po are present due to the input of

210

and

222

Rn from stratosphere–troposphere exchanges, and the relatively long

(approximately a year) stratospheric residence times.

4.3. OCCURRENCE OF

210

Pb AND

210

Po IN TERRESTRIAL

ECOSYSTEMS

4.3.1.

210

Pb and

210

Po in plants

Terrestrial plants receive

210

Pb and

210

Po from three principal sources:

(a) Radioactive decay from

222

Rn already accumulated by the plant;

(b) Root uptake from the soil;

210

Pb and

210

Po deposited on the leaves.

It is generally believed that foliar uptake dominates for

210

Po, which is thus

accumulated primarily by absorption and further translocation within the plant

(see Ref.

[4.35]). However, the relative importance of the root and foliar uptake

pathways depends on the concentration of the radionuclides in the soil, the soil–

plant concentration ratio (CR), and the rate of deposition onto plant parts above

ground.

Another important consideration is that in crops such as roots, tubers,

cereals, nuts and legumes whose edible portion is protected by inedible plant parts,

activity concentrations should not be significantly affected by direct deposition.

In contrast, observed concentrations for leafy vegetables may be up to 3–5

times

higher due to deposition effects. Therefore, there can be a large variation in the

radionuclide concentrations in plants, resulting from differences in radionuclide

accumulation in different plant species, variable tissue distributions, climate

variations and soil properties.

Sheppard et al. [4.7] report activity concentrations in plants in Canada in

the range of 1–730

Bq/kg (DW) for

210

Pb and 0.3–460 Bq/kg (DW) for

210

Po.

Data of the same order of magnitude were reported in 1976 for

210

Pb and

210

concentrations in Russian plants by Ladinskaya et

al. [4.36].

Based on a wide review in 1990 of the data available for terrestrial

plants

[4.22, 4.36–4.39], Taskayev and Testov [4.40] report a mean value

of 11 Bq/kg

(DW) with a range of 0.04–111 Bq/kg (DW) for

210

Po activity

concentrations in grassy plants sampled in areas of normal background radiation.

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

Activity concentrations of

210

Pb and

210

Po in areas of high natural background are

found to be typically two to three orders of magnitude higher compared to normal

areas. Overall, these estimates are in agreement with assessments provided in

the 2004 review on

210

Po environmental behaviour performed by Coppin and

Roussel-Debet

[4.8] (see Table 4.3).

TABLE 4.3. CONCENTRATIONS OF

210

Po IN TERRESTRIAL PLANTS

Country Plant

Activity conc. of

210

(Bq/kg, DW)

Ref.

Areas undisturbed by humans

Germany Grass and hay 1.1–29.6 [4.12]

Grass

Heather

Juniper

Blueberry

22–160

20–88

20–38

[4.11]

India Tobacco 0.1–3.3 [4.14]

Portugal Vegetables

Fruits

Trees

0.084

0.06

0.37

[4.41]

United Kingdom Grass

Lichens

6.5–29

290–370

[4.37]

World data Grass 2.0–35 [4.42]

Tobacco 5.6–57 [4.22]

Areas disturbed by humans

Brazil, uranium mine at 1 km Vegetables 1–8 [4.9]

Canada, uranium mine at Key Lake Spruce needles

Labrador tea

1300

[4.10]

Spain, near phosphoric acid factory

Spartina spp.

5.1–40.6 [4.18]

United Kingdom, Cotswolds Grass 4–25 [4.43]

Source: See Ref. [4.8].

Note: The activity concentrations for Portugal are based on fresh weight.

CHAPTER 4

4.3.2.

210

Pb and

210

Po in animals

Overall, the distributions of

210

Pb and

210

Po in animals, and their activity

concentrations in different animal tissues reflect their intake with feed. The

activity concentrations of

210

Po in animals vary across four orders of magnitude

depending on the animal and organs (see Table

4.4). Lead-210 is largely retained

in bones, while

210

Po is distributed mainly in soft tissue. Target tissues for

210

are the spleen, liver and kidneys. The

210

Po:

210

Pb ratio typically exceeds unity

in soft tissue; for example, ratios greater than 100 have been reported for the

muscle tissue of wild boar

[4.51]. However, observed

210

Po:

210

Pb ratios are time

dependent and can be substantially modified by continuing radioactive decay

from

210

Pb, resulting in support for

210

Po in older animals.

TABLE 4.4. CONCENTRATIONS OF

210

Po IN ANIMAL TISSUES

Tissue

210

Po activity conc.

(Bq/kg, FW)

Ref.

Meat

Mutton

Beef and pork

0.11–0.43

0.037–2.0

[4.44, 4.45]

[4.46]

Liver

Beef, lamb and mutton

Chicken

Caribou

0.15–120

0.21–1.03

332

[4.11, 4.37, 4.42, 4.47]

[4.48]

[4.10]

Kidney

Beef, lamb and mutton

0.74–67

[4.37, 4.42, 4.47, 4.49]

Gizzards

Chicken

0.48

[4.48]

Offal 0.19–37

6.2 × 10

−2*

[4.22, 4.50]

Milk (0.33–6.7) × 10

−2

[4.12, 4.22]

Eggs 0.11–37 [4.46, 4.48, 4.50]

Source: See Ref. [4.8].

Prepared for consumption.

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

4.4. OCCURRENCE OF

210

Pb AND

210

Po IN MARINE ECOSYSTEMS

4.4.1.

210

Pb and

210

Po in marine water

Polonium enters the marine environment by deposition from the atmosphere

at the ocean surface, from the in

situ radioactive decay of soluble

226

Ra and from

the decay of

222

Rn gas exhaled from the sea-floor. It also enters the coastal seas

with river water and sediment discharges, which includes a component from

anthropogenic discharges. The resulting

210

Po concentrations in sea water are

largely dependent on the magnitude and distribution of these sources, on

210

binding to suspended particulate matter and on particle scavenging in the water

column.

Polonium-210 in coastal waters is largely associated with suspended

particulate matter (around 70–80%), probably due to higher suspended loads

and intensive mixing in coastal waters than in open ocean waters. Lead-210 in

coastal sea water is also mainly associated with the suspended matter, although

to a lesser extent than

210

Po. The concentrations of these radionuclides in coastal

waters do not show a significant seasonal variation, and thus the atmospheric

depositions do not significantly modify those concentrations.

In open waters, the ocean is stratified and vertical mixing is reduced.

Within the water column,

210

Po is produced from radon decay. Given that radon

concentrations in the water column increase from a minimum in the surface

layer to a maximum in deep-sea water, and the radon distribution is mainly

dependent on diffusion in the water column and marine currents, surface water

concentrations of

210

Po in the open ocean are generally low.

Deposition from the atmosphere results in an excess of

210

Pb over

226

Ra in

the surface layer of the ocean. Absorption of

210

Pb and

210

Po onto particulates

and uptake by phytoplankton and zooplankton generates a downward flux of

particulate associated activity. Although these particles are, in part, recycled in

intermediate ocean layers, in their settling pathway through the deeper ocean

layers, they adsorb and remove soluble

210

Pb and

210

Po from the water column,

thus creating an imbalance between

226

Ra and its progeny. Some of the biogenic

particulate materials generated in the upper layer of the ocean may reach the

deep-sea (abyssal) floor and add

210

Pb and

210

Po to the top sediment layer. The

210

Pb flux arriving at the abyssal sea-floor is around double the

210

Pb atmospheric

flux entering the oceans at the surface. The mean residence times for

210

Po in

ocean water layers have been calculated to be around 6–12 months in the upper

layer and somewhat longer, around 2 years, in the deep-sea layers [4.52, 4.53].

CHAPTER 4

4.4.2.

210

Pb and

210

Po in marine organisms

The isotopes

210

Pb and

210

Po have been well studied in marine wildlife

owing to their relatively high concentrations in comparison with those in

terrestrial organisms. Currently known

210

Pb and

210

Po activity concentrations in

marine biota cover several orders of magnitude. With a few exceptions in the

case of abyssal fauna,

210

Po is in excess over

210

Pb. Reported

210

Po:

210

Pb ratios

are mostly in the range of 1–100

[4.54]. Carvalho [4.54] finds that organisms

occupying upper trophic levels (carnivores and top predators) generally display

lower

210

Po concentrations than planktivorous organisms (primary herbivores).

Furthermore, related species occupying equivalent ecological niches in different

ecosystems (e.g. filter feeding bivalves in the intertidal zone and in the abyssal

sea-floor) display similar

210

Po activity concentrations and

210

Po:

210

Pb ratios.

For higher trophic level organisms, such as shrimp, molluscs and fish, the

primary source of

210

Po uptake is from ingested food, with very little uptake

from water

[4.55–4.58]. Inside these organisms, the distribution of

210

Po is not

homogeneous, and the tissues and organs with higher concentrations are those

associated with the digestive system.

Concentrations of

210

Pb do not increase from phytoplankton to copepods

as much as

210

Po, and the

210

Pb transfer to planktivorous fish, such as sardines,

is less efficient than

210

Po transfer. Consequently, the

210

Po:

210

Pb ratio in marine

food webs appears to increase with the trophic level, at about 10 in phytoplankton

and zooplankton, 3–10 in herbivorous fish, 50–100 in piscivores (carnivorous

fish) and 200 in the muscle of marine mammals (top predators)

[4.59].

4.5. OCCURRENCE OF

210

Pb AND

210

Po IN FRESHWATER

ECOSYSTEMS

4.5.1.

210

Pb and

210

Po in groundwater

Rainwater infiltrating the ground interacts with soils, sediments and

rocks, and so it can accumulate polonium from a number of sources. This

includes polonium produced from naturally occurring sources within soils in the

unsaturated zone, polonium that has been naturally deposited onto the surface

from atmospheric sources, and anthropogenic sources. These sources add to the

polonium carried in the rainwater as wet deposition. Polonium is also available

from progenitor nuclides in the soils and is derived from similar sources. The

infiltrating waters then recharge the underlying groundwater systems.

Weathering of uranium bearing mineral phases will not normally be a major

source term for

210

Pb and

210

Po in groundwater, as it typically occurs at rates

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

that are very long compared to the decay rate of the nuclides. Rather, they are

primarily available to the groundwater phase due to recoil following alpha decay

of progenitor atoms in the uranium decay series (see Sections

3.1.2 and 7.2). This

can take place from within the mineral grains, from absorption sites on grain

surfaces, and from within the groundwater phase.

Lead and polonium are both strongly particle reactive under most conditions

pertaining to groundwater. Consequently, the largest fraction of activity released

from mineral grains is adsorbed onto the aquifer rocks. The distribution of lead

and polonium in groundwater is therefore strictly controlled by the adsorption

characteristics of the surrounding minerals. Groundwater

210

Po concentrations

are typically in the range of 1 –30 mBq/L, although values of up to 19 Bq/L have

been recorded for brines.

The strong association of

210

Po with mineral surfaces and particles, and its

short half-life, means that significant transport of activity to a significant distance

from the site of production is unusual and measured concentrations generally

reflect the local supply rates and bulk adsorption coefficients. The study of

210

in groundwater is therefore primarily of interest from a human health perspective

when the water is to be used for drinking or irrigation, and as a means of studying

geochemical processes.

Groundwater discharge into surface waters, such as lakes, rivers and

wetlands, as well as into coastal waters. These discharges can carry polonium,

although the concentrations may be modified at the sediment–water interface.

The relationship between polonium in groundwater and surface waters has not

been studied in detail.

4.5.2.

210

Pb and

210

Po in surface water and sediments

The freshwater environment encompasses lotic (moving) water, such as

rivers and streams, and standing water bodies (lentic), such as lakes, ponds,

wetlands and bogs. The concentrations of

210

Po within moving water (e.g. a river)

would be expected to reflect the environmental conditions in the catchment area

as well as both natural and anthropogenic sources. The concentrations throughout

the year will vary depending on variable precipitation carrying particulates from

the surrounding areas to the water body.

The levels of

210

Po in surface water in freshwater systems are influenced

by in

situ decay, fluvial inputs and atmospheric deposition. Within a standing

water body (e.g. a lake), there is uptake of

210

Pb and

210

Po by particulates

(particularly biomass) in the water column. As these particulates settle to the

bottom sediments, polonium is scavenged from the water column. Polonium has

also been shown to become volatile in both fresh and marine waters by the action

of microorganisms

[4.60].

CHAPTER 4

The levels of

210

Pb and

210

Po in sediment are influenced by the cycling of

iron and manganese as the redox conditions change in the sediment. Overall,

a portion of the lead and polonium in sediment may be reintroduced into the

water column by diffusive processes; the remainder becomes buried by continual

deposition of fresh settling matter. In addition to cycling in sediment, depending

on the characteristics of the lake, the bottom layer within the water column may

become anoxic during portions of the year, which will influence the levels of

polonium in the water column. During the anoxic period, levels of

210

Pb and

210

Po can be significantly increased in the hypolimnion; due to differences in the

cycling, the

210

Po:

210

Pb ratio can be greater than one.

The activity concentrations of

210

Po in oxic waters are mostly in the range

of 1–5

mBq/L and are higher, up to 17 mBq/L, in seasonally anoxic ponds [4.61].

On a local scale, there can be a large range of typical values.

4.5.3.

210

Pb and

210

Po in freshwater organisms

The levels of

210

Po have been measured in freshwater biota, but there are not

as many data available as for the marine environment. When compared to marine

biota, a similar pattern in freshwater biota is observed: high concentrations in

plankton and lower concentrations in higher trophic biota. The distribution of

210

Po within an organism shows higher concentrations in the kidneys, liver and

organs of the digestive system. A higher accumulation of

210

Po is found in the

soft tissue of molluscs than in shells, whereas

210

Pb has been detected in higher

levels in the shells. In the soft tissue of an organism,

210

Po is typically found at

levels greater than

210

Pb [4.51].

Whole organism CRs for lead (FW) range from approximately

5 (amphibians) [4.62] to as high as 6000 (bivalve molluscs) [4.63]; CRs for

polonium are generally higher, from 2000 (fish and vascular plants) up to

1.3 × 10

(bivalves) [4.64, 4.65]. Lead and polonium CRs for wildlife groups in

freshwater ecosystems can be found in Table 7.3, Section 7.3.5.

4.6. OCCURRENCE OF

210

Pb AND

210

Po IN FOOD

(INCLUDING SEAFOOD) AND DRINKING WATER

The consumption of food and drinking water is the most important exposure

pathway for

210

Pb and

210

Po for humans. Data on activity concentrations of

210

Pb and

210

Po in drinking water for different countries are given in Table 4.5.

Mean activity concentrations of

210

Po in drinking water are in the range of

0.04–7600

mBq/L, and for most regions the upper mean value is less than

mBq/L. Furthermore,

210

Po concentrations can vary by more than four orders of

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

magnitude because of the diversity of water sources. High activity concentrations

210

Po are found in areas of high radiation background in Finland (up to

7600

mBq/L), while the lowest values are typical for public water supply systems

in countries with efficient water quality control systems, such as Italy, where the

concentrations of

210

Po in tap water are in the range of 0.1–5.9 mBq/L. Activity

concentrations of

210

Pb in drinking water are generally two to fivefold higher

than those measured for

210

Pb. The mean value for most countries is in the range

of 1.5–26.9

mBq/L.

TABLE 4.5. ACTIVITY CONCENTRATION OF

210

Pb AND

210

Po IN

DRINKING WATER

Country Origin

210

Pb (mBq/L)

210

Po (mBq/L)

Mean Range Mean Range

Brazil Mineral water 66 20–102 —

—

Czech Republic Public supply —

<40–350 9 <2–71

Finland Potable water —

0.2–21 000 —

0.2–7 600

Germany Potable water 1.5 0.2–170 0.5 0.1–40

India (Goa) Potable water —

2.2–11.5 —

1.7–7.0

Italy Tap water

Mineral water

—

0.13–5.9

<0.04–24

Poland Potable water 1.6 —

0.5 —

Portugal Public supply 18.5 2–390 —

—

Romania Water supply —

7.0–44 —

7.0–44

Russian Federation Water supply 26.9 2.3–68 2.72 1.2–7.1

Spain Public supply 18.7 0.23–59 3.63 1.8–19.3

United Kingdom Public supply —

40–200 —

—

USA Potable water —

0.1–1.5 —

—

Source: See Ref. [4.66].

—: data not available.

CHAPTER 4

Concentrations of

210

Pb and

210

Po in drinking water vary over a wide range

because of variations in the concentrations of uranium and its long lived progeny,

such as

226

Ra and

230

Th, in adjacent rocks, soil, atmosphere and groundwater.

High levels of

210

Pb and

210

Po have been found to be associated with regions of

uranium or thorium ore bodies. Mining and processing of uranium and other

minerals, and geothermal springs also lead to elevated concentrations of

210

and hence

210

Po in the environment.

Concentrations of

210

Pb and

210

Po in foods also vary over a wide range,

reflecting differences in uranium and its progeny in soil, the atmosphere

and water, differing farming practices and local features in food processing

(see Tables 4.6 and 4.7).

TABLE 4.6. CONCENTRATIONS OF

210

Pb AND

210

Po IN FOODS

Product Region Country

210

Pb (mBq/kg)

210

Po (mBq/kg)

Milk products North America USA 11 —

Asia China

India

—

Europe Germany

Poland

Romania

United Kingdom

5–280

10–15

35–88

2–80

13–140

20–220

Reference value 15 15

Meat products North America USA 18 —

Asia China

India

140

—

120

440

Europe Germany

Poland

Romania

United Kingdom

100–1 000

98–105

15–19

40–3 700

37–4 000

99–102

38–110

62–67 000

Reference value 80 60

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

TABLE 4.6. CONCENTRATIONS OF

210

Pb AND

210

Po IN FOODS (cont.)

Product Region Country

210

Pb (mBq/kg)

210

Po (mBq/kg)

Grain products North America USA 33–81 —

Asia China

India

—

15–120

Europe Germany

Poland

Romania

United Kingdom

40–4 000

110–160

49–59

56–120

37–1 900

90–140

20–360

27–260

Reference value 50 60

Leafy vegetables North America USA 41 —

Asia China

India

360

—

430

320

Europe Germany

Poland

United Kingdom

4–4 100

43–51

16–3 300

4–7 400

40–67

37–3 300

Reference value 80 100

Root vegetables

and fruits

North America USA 8–150 —

Asia China

India

—

16–140

Europe Germany

Poland

Romania

United Kingdom

20–4 900

24–93

19–44

18–76

22–5 200

28–210

12–140

—

Reference value 30 40

Fish products North America USA 14–1 800 150–55 000

Europe Germany

Poland

Portugal

United Kingdom

20–4 400

81–93

—

180–4 800

50–5 200

3 100–3 800

80–120 000

60–53 000

Reference value

200 2 000

Source: See Ref. [4.67].

—: data not available.

CHAPTER 4

The lowest ranges in

210

Pb and

210

Po concentrations are in milk products.

Two to sixfold higher concentrations were observed in meat and plant products,

where root vegetables had slightly lower mean

210

Pb and

210

Po concentrations

compared with other plant products. The highest concentrations were found

to be in fish products, and leafy vegetables demonstrated

210

Pb concentrations

higher than in grain products. High

210

Po concentrations are consistently found

TABLE 4.7. CONCENTRATIONS OF

210

Pb AND

210

Po IN MUSCLES OF

COMMON SEAFOOD

Species Body weight

(g)

No. of

samples

210

(Bq/kg, FW)

210

(Bq/kg, FW)

Sardine 51 4 66 ± 2 1.0 ± 0.02

Anchovy 5.5 6 9.4 ± 0.3 0.42 ± 0.01

Mackerel 120 4 19 ± 1 0.63 ± 0.4

Horse mackerel 200 2 5.2 ± 0.2 0.1 ± 0.02

Bigeye tuna 31 × 10

1 3.1 ± 0.09 0.46 ± 0.02

Hake 256 2 6.7 ± 0.3 0.15 ± 0.01

Red sea bream 275 1 2.4 ± 0.09 0.84 ± 0.02

Common sole 150 1 1.4 ± 0.04 0.14 ± 0.01

Skate 894 1 0.73 ± 0.03

0.12 ± 0.000 4

Squid 2 103 1 1.6 ± 0.04 0.41 ± 0.01

Common shrimp —

12 49 ± 1.5 1.1 ± 0.03

Clam —

6 152 ± 19 2.9 ± 0.1

Cockle —

5 9.4 ± 3 1.32 ± 0.06

Mussel —

12 132 ± 5 2.6 ± 0.1

Source: See Ref. [4.68].

Note: All samples were taken from the north-eastern Atlantic Ocean, off Portugal.

—: data not available.

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

in seafood products, and with

210

Po:

210

Pb ratios much higher than unity, usually

2–100

[4.67–4.69].

The great importance of

210

Po to dietary intake is found in many coastal

countries, such as Japan, the Marshall Islands, Portugal and South Africa [4.67]. A

global review of

210

Po in marine food suggests that representative concentrations

are 2.4 Bq/kg in fish, 6.0 Bq/kg in crustaceans and 15.0 Bq/kg in molluscs [4.70].

Consumption of seafood varies widely, both between countries and within a

single country. In addition,

210

Pb and

210

Po concentrations in seafood species vary

within three orders of magnitude (see Table 4.7). If representative consumption

rates are 13 kg/a of fish and 1 kg/a each of molluscs and crustaceans, the intake

210

Po with these foods would be 52 Bq/a [4.67].

If there are processing or distribution delays for fish products between catch

and consumption, the activity intake will be reduced owing to the radioactive

decay of

210

Po. Statistics quoted by Aarkrog et al. [4.70] indicate that 30% of

seafood is eaten fresh, 30% frozen, 20% smoked and 20% canned. Application

of a correction factor of 0.6 [4.67] suggests an intake of 31 Bq/a in seafood and a

weighted concentration of

210

Po in fish products of 2.1 Bq/kg.

4.7. ANTHROPOGENIC SOURCES OF

210

Po IN THE ENVIRONMENT

4.7.1. Industrial uses of human-made

210

There are a number of industrial uses of

210

Po, in particular as an alpha

emitter to eliminate static in manufacturing environments and to remove dust

under clean conditions. Owing to its short half-life, the

210

Po source needs to be

replaced approximately every twelve months. It has also been investigated as a

heat source for thermoelectric power devices for space applications, since the

energy released by decay is so great (140 W/g).

In nature,

210

Po occurs in extremely low concentrations; even uranium

ores contain less than 0.1 mg/t. However,

210

Po is generated in nuclear reactors

when stable

209

Bi is bombarded with neutrons. Around 100 g of

210

Po is produced

annually, largely at a single facility in the Russian Federation. There have not

been any widely reported releases of

210

Po from such facilities.

Earlier designs of nuclear weapons included a neutron source of beryllium

and polonium, and so weapons programmes required additional production of

210

Po. In the United Kingdom, this took place at Windscale Works, Sellafield. A

nuclear reactor fire there in 1957 led to the environmental release of radionuclides.

In addition to a study published in 1996 on the magnitude of the release of

210

(see Ref. [4.71]), a 2007 study of

137

Cs,

131

I and

210

Po estimated an approximate

release of 40 TBq of

210

Po, with a range of 14–110 TBq [4.72].

CHAPTER 4

4.7.2. Environmental release of naturally occurring radioactive material

Naturally occurring

210

Po is continuously released from geological materials

processed for industrial use. In a 1976 study, Moore et

al. [4.73] estimate that

anthropogenic sources could constitute up to 7% of the total

210

Po fluxes in the

atmosphere, while elevated concentrations in soils and waters are found near

related industrial activities.

4.7.2.1. Phosphate fertilizer and phosphogypsum waste

Phosphate rich rocks are extensively mined, largely to produce phosphate

fertilizer. These rocks, especially those of sedimentary origin, can have high

concentrations of uranium and daughter

210

Po. Both the products and waste of these

operations can therefore release radionuclides to the environment

[4.74, 4.75].

Phosphate rocks contain up to 3700

Bq/kg of uranium [4.75], with high

uranium content generally corresponding to high phosphorus content. Physical

processing of ore can increase radionuclide concentrations by up to 300%

[4.76].

During chemical processing, phosphate rock is reacted with sulphuric acid

and converted into phosphoric acid to make fertilizer, while phosphogypsum

is generated as a by-product. While uranium and

210

Pb primarily partition into

the phosphoric acid, much of the

226

Ra and up to 99% of the

210

Po enter the

phosphogypsum

[4.74, 4.77].

Most phosphogypsum is stored in piles (referred to as stacks), although

some is discharged into surface waters. A study in the Netherlands examined the

impact of phosphogypsum discharged by the fertilizer industry into the Rhine

(~1.6

TBq/a) and transported by suspension to the North Sea [4.78]. The

210

and

210

Po settle with particulates, with some of the

210

Po becoming involved

in biological activity in the upper marine water column (see Chapter

8). The

sediments thus remain a potential source of

210

Po, although much of the

210

Pb and

226

Ra — and hence probably the daughter

210

Po — is trapped within insoluble

fractions

[4.79].

Radionuclides can also escape from processing plants or phosphogypsum

storage stacks. For example, a study around a fertilizer plant in the Syrian Arab

Republic found elevated levels of

210

Po in surrounding soil, water and vegetation,

attributable to deposition of daughter isotopes of

222

Rn escaping from the

facility

[4.80]. A fertilizer production facility along the Tagus estuary, Portugal,

has been found to contribute radionuclides to near-shore sediments through

discharge of process waters and leaching by rainwater of phosphogypsum

piles

[4.81]. Overall, the radiological impact of naturally occurring radioactive

material industries on the European population is dominated by discharges from

fertilizer plants (see Ref. [4.82]).

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

The spread of phosphate fertilizer, containing

210

Pb, which decays to

210

Po,

can contribute to the naturally occurring radionuclides in soils. Phosphogypsum,

containing

210

Po, has also been used as a soil amendment for depleted soils in

some countries (see Ref.

[4.83]).

4.7.2.2. Fossil fuel production and combustion

Fossil fuels are often rich in uranium series nuclides, since the precursor

organic material was accumulated under reducing conditions favourable for

incorporation of uranium. Petroleum in the United States of America can contain

0.1–40

Bq/kg of

226

Ra [4.82], presumably with similar

210

Pb and

210

Po activities.

Production waste contains high levels of radium, with on average approximately

000 Bq/kg of radium in scale and 2800 Bq/kg of radium in sludge. These

wastes can be significant sources of

210

Pb and

210

Po when discharged, landfilled

or disposed of by land spreading

[4.84]. Waters discharged from production wells

can have high concentrations of

226

Ra, although low levels of

210

Po. For example,

waters from Norwegian wells have up to 16

Bq/L of

226

Ra but only 0.2–6 mBq/L

210

Po [4.85].

The behaviour of

210

Po in coal combustion power plants is considered

by Mora et

al. [4.86]. Coal typically has concentrations of 16–52 Bq/kg of

210

Po [4.75], which is volatilized during combustion. More than half of this is

condensed onto fly ash, although some also condenses within the installations,

and a small fraction remains in bottom ash. It is expected that

210

Pb behaves

similarly. Most of the process residues are deposited onto land, although some fly

ash is used in building materials

[4.75].

4.7.2.3. Metal production

One clear potential source of uranium series radionuclides is tailings

from uranium and radium mining sites, where leaching and runoff can disperse

contaminants. As discussed in Chapter 6, lead and polonium are relatively

immobile, except under highly acidic conditions. Radium is also relatively

immobile [4.87], and while uranium is more readily transported, in the absence

of other intermediate radionuclides, there will not be substantial

210

Po ingrowth

over short timescales (see Chapter 3). Therefore, the spread of

210

Po, and of

210

Pb followed by ingrowth of

210

Po, is often through discharge of acidic waters

and waste pile leachates, and the transport of airborne particles and suspended

sediments. The escape of

222

Rn to the atmosphere and subsequent deposition of

daughter nuclides is also an important route for spreading

210

Po (see Chapter 5).

Some of the geochemistry controlling radionuclide behaviour is summarized in

Ref. [4.88]. There have been various documented releases from mine tailings,

CHAPTER 4

in particular acute failures of containment structures: major events are listed in

Ref. [4.89] and an example of the release of

210

Po to the environment is presented

in chapter 7 of Ref. [4.90].

Ore processing for other metals can also release

210

Po and other

radionuclides. For example, Baxter et al. [4.91] report that raw materials for a

large tin smelter in the United Kingdom generally contain up to 2 × 10

Bq/kg

210

Po, and there was considerable loss of

210

Po in stack emissions, although

high concentrations were not found in surrounding soils. The technological

enhancement of radionuclides in primary products or waste materials can occur

during the processing of other metals as well [4.92]. Separation of zirconium

also releases radionuclides from zircon, which contains high concentrations of

uranium [4.93].

4.8. SUMMARY

The isotopes

210

Pb and

210

Po are widely distributed in the Earth’s crust

and soil. As polonium is only ever present at ultratrace levels and lead is usually

present at trace levels, their behaviour is heavily influenced by major ion

chemistry and processes such as adsorption, co-precipitation, and attachment to

colloids in aqueous systems and to aerosols in the atmosphere.

The activity concentrations of

210

Pb and

210

Po in the topsoil vary over a

wide range, with a range in most soils of 10–200 Bq/kg. The

210

Po:

210

Pb ratio in

soil is quite close to unity, implying that they are in secular equilibrium in soils.

Major sources to the atmosphere are decay of

222

Rn following its exhalation

from soils, and volcanic emissions (particularly for

210

Po), although there are

several other minor sources. Removal is primarily by deposition to the Earth’s

surface with precipitation.

For plants,

210

Po:

210

Pb activity ratios are in the range of 0.2–0.7 in above

ground biomass and 0.96–1.3 in roots, indicating different mobilities of

210

Pb and

210

Po in the soil–plant system as well as within plants. A greater variability in

concentrations is observed for animals, both terrestrial and aquatic. Lead-210 is

largely retained in bones, while

210

Po is distributed mainly in soft tissue. Target

tissues for

210

Po are the spleen, liver and kidneys. Relatively high concentrations

210

Po have been measured in marine animals, and ingestion of seafood is an

important pathway for radiological dose to humans.

Both lead and polonium are strongly particle reactive under most conditions

pertaining to groundwater, and most of the available activity is adsorbed onto

aquifer rocks. Their distribution is therefore strictly controlled by the adsorption

characteristics of the surrounding minerals. Groundwater

210

Po concentrations

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

are typically in the range of 1–30 mBq/L, although values up to 19 Bq/L have

been recorded for brines.

The realistic behaviour of

210

Po in the environment can be described if

based on both predictions of environmental behaviour of

210

Pb as a source and

transfer of

210

Po itself, and so the environmental behaviour of these radionuclides

needs to be considered together.

REFERENCES TO CHAPTER 4

[4.1] GASCOYNE, M., “Geochemistry of the actinides and their daughters”, Uranium-

Series Disequilibrium: Applications to Earth, Marine, and Environmental Sciences

(IVANOVICH, M., HARMON, R.S., Eds), Clarendon Press, Oxford (1992)

34–61.

[4.2] RUDNICK, R.L., GAO, S., “Composition of the continental crust”, Treatise on

Geochemistry, Vol.

3: The Crust (HOLLAND, H.D., TUREKIAN, K.K., Eds),

Elsevier, Oxford (2003)

1–64.

[4.3] BALKANSKI, Y.J., JACOB, D.J., GARDNER, G.M., GRAUSTEIN, W.C.,

TUREKIAN, K.K., Transport and residence times of tropospheric aerosols inferred

from a global three-dimensional simulation of

210

Pb, J. Geophys. Res. 98

(1993)

20 573–20 586.

[4.4] BASKARAN, M., Po-210 and Pb-210 as atmospheric tracers and global atmospheric

Pb-210 fallout: A review, J. Environ. Radioact.

102 (2011) 500–513.

[4.5] VAARAMAA, K., ARO, L., SOLATIE, D., LEHTO, J., Distribution of

210

Pb and

210

in boreal forest soil, Sci. Total Environ.

408 (2010) 6165–6171.

[4.6] BROWN, J.E., et al., Levels and transfer of

210

Po and

210

Pb in Nordic terrestrial

ecosystems, J. Environ. Radioact.

102 (2011) 430–437.

[4.7] SHEPPARD, S.C., SHEPPARD, M.I., ILIN, M., TAIT, J., SANIPELLI, B., Primordial

radionuclides in Canadian background sites: Secular equilibrium and isotopic

differences, J. Environ. Radioact.

99 (2008) 933–946.

[4.8] COPPIN, F., ROUSSEL-DEBET, S., Comportement du

210

Po en milieu terrestre: revue

bibliographique, Radioprot. 39 (2004) 39–58.

[4.9] SANTOS, P.L., GOUVEA, R.C., DUITA, I.R., GOUVEA, V.A., Accumulation of

210

Po in foodstuff cultivated in farms around the Brazilian mining and milling facilities

on Poços de Caldas plateau, J. Environ. Radioact.

11 (1990) 141–149.

[4.10] THOMAS, P.A., Radionuclides in the terrestrial ecosystem near a Canadian uranium

mill — Part

I: Distribution and doses, Health Phys. 78 (2000) 614–624.

[4.11] BUNZL, K., KRACKE, W., Distribution of

210

Pb and

210

Po, stable lead and fallout

137

Cs in soil, plants and moorland sheep of a heath, Sci. Total Environ. 39

(1984)

143–159.

[4.12] SCHÜTTELKOPF, H., KIEFER, H., “The Ra-226, Pb-210 and Po-210 contamination

of the Black Forest in natural radiation environment”, Natural Radiation Environment

(Proc. 2nd Special Symp. Bombay, 1981) (VOHRA, K.G., MISHRA, V., PILLAI, K.,

SADAÏWAN, S., Eds), John Wiley and Sons, New

York (1982).

CHAPTER 4

[4.13] ROSNER, G., BUNZL, K., HÖTZL, H., WINKLER, R., Low level measurements of

natural radionuclides in soil samples around a coal-fired power plant, Nucl. Instrum.

Methods Phys. Res.

223 (1984) 585–589.

[4.14] SINGH, D.R., NILEKANI, S.R., Measurement of polonium activity in Indian tobacco,

Health Phys.

31 (1976) 393–394.

[4.15] KARUNAKARA, D., et al., Distribution and enrichment of

210

Po in the environment

of Kaiga in South India, J. Environ. Radioact.

51 (2000) 349–362.

[4.16] SIDDAPPA, K., et al., Distribution of natural and artificial radioactivity components in

the environs of costal Karnataka, Kaiga and Goa (1991–94), Mangalore Univ. (1994).

[4.17] HOUMANI, Z.M.M., BRADLEY, D.A., MAAH, M.J., AHINED, Z., Concentration of

210

Po,

226

Ra and

228

Ac in non-siliceous environmental materials, Radiat. Phys.

Chem.

61 (2001) 665–668.

[4.18] MARTINEZ-AGUIRRE, A., GARCIA-LEON, M., Transfer of natural radionuclides

from soils to plants in a wet marshland, Appl. Radiat. Isot.

47 (1996) 1103–1108.

[4.19] SMITH-BRIGGS, J.L., BRADLEY, E.J., Measurement of natural radionuclides in UK

diet, Sci. Total Environ.

35 (1984) 431–440.

[4.20] UNITED STATES ATOMIC ENERGY COMMISSION, Polonium-210 in Soils and

Plants, USAEC Special Report, Vol.

11, USAEC (1980).

[4.21] LAPHAM, S.C., MILLARD, J.B., SAMET, J.M., Health implications of radionuclide

levels in cattle raised near U mining and milling facilities in Ambrosia

Lake,

New

Mexico, Health Phys. 56 (1989) 327–340.

[4.22] PARFENOV, Yu.D., Polonium-210 in the environment and in the human organism, At.

Energy Rev.

12 (1974) 75–143.

[4.23] PERSSON, B.R.R., HOLM, E., Polonium-210 and lead-210 in the terrestrial

environment: A historical review, J. Environ. Radioact.

102 (2011) 420–429.

[4.24] HARMSEN, K., DE HAAN, F.A.M., Occurrence and behaviour of uranium and

thorium in soil and water, Neth. J. Agric. Sci.

28 (1980) 40–62.

[4.25] SYED, H.S., Comparison studies adsorption of thorium and uranium on pure clay

minerals and local Malaysian soil sediments, J. Radioanal. Nucl. Chem.

241

(1999)

11–14.

[4.26] LAMBERT, G., BUISSON, A., SANAK, J., ARDOUIN, B., Modification of the

atmospheric polonium 210 to lead 210 ratio by volcanic emissions, J. Geophys. Res.

(1979)

6980–6986.

[4.27] PREISS, N., MÉLIÈRES, M.-A., POURCHET, M., A compilation of data on lead 210

concentration in surface air and fluxes at the air–surface and water–sediment interfaces,

J. Geophys. Res.

101 (1996) 28 847–28 862.

[4.28] ANAND, S.J.S., RANGARAJAN, C., Studies on the activity ratios of polonium-210

to lead-210 and their dry deposition velocities at Bombay in India, J. Environ.

Radioact.

11 (1990) 235–250.

[4.29] MARENCO, A., FONTAN, J., Sources of polonium-210 within the troposphere,

Tellus

24 (1972) 38–46.

[4.30] NHO, E.-Y., ARDOUIN, B., LE CLOAREC, M.F., RAMONET, M., Origins of

210

in the atmosphere at Lamto, Ivory

Coast: Biomass burning and saharan dusts, Atmos.

Environ.

30 (1996) 3705–3714.

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

[4.31] WINKLER, R., HÖTZL, H., CHATTERJEE, B., Analysis of

210

Pb and

210

concentrations in surface air by an alpha spectrometric method, Health Phys.

(1981)

495–503.

[4.32] MOORE, H.E., POET, S.E., MARTELL, E.A.,

222

Rn,

210

Pb,

210

Bi and

210

Po profiles

and aerosol residence times versus altitude, J. Geophys. Res.

78 (1973) 7065–7075.

[4.33] LIU, H., JACOB, D.J., BEY, I., YANTOSCA, R.M., Constraints from

210

Pb and

Be on

wet deposition and transport in a global three-dimensional chemical tracer model

driven by assimilated meteorological fields, J. Geophys. Res.

106

(2001)

12 109–12 128.

[4.34] MARTIN, P., Uranium and thorium series radionuclides in rainwater over several

tropical storms, J. Environ. Radioact.

65 (2003) 1–18.

[4.35] MAHON, D.C., MATHEWES, R.W., Uptake of naturally occurring radioisotopes by

vegetation in a region of high radioactivity, Can. J. Soil Sci. 63 (1983) 281–290.

[4.36] LADINSKAYA, L.A., PARFENOV, Yu.D., ARUTYNOV, Ya.C., POPOVA, D.K.,

Natural concentrations of Pb-210 and Po-210 in plants of steppe ecosystems,

Ecologia

1 (1976) 32–35 (in Russian).

[4.37] HILL, C.R., Polonium-210 in man, Nature 208 (1965) 423–428.

[4.38] BEASLEY, T.M., PALMER, H.E., Lead-210 and polonium-210 in biological samples

from Alaska, Health Phys.

152 (1966) 1062–1064.

[4.39] HOLTZMAN, R.B., Natural levels of lead-210, polonium-210 and radium-226 in

humans and biota of the Arctic, Nature

210 (1966) 1094–1097.

[4.40] TASKAYEV, A.I., TESTOV, B.V., “Major components of doses to biological

organisms from incorporated primordial radionuclides”, Heavy Natural Radionuclides

in Biosphere (ALEXAKHIN, R.M., Ed.), Nauka, Moscow (1990) 201–217

(in Russian).

[4.41] CARVALHO, F.P., Contribution à l’étude du cycle du

210

Po et du

210

Pb dans

l’environnement, PhD Thesis, UFR Sciences, Univ. Nice Sophia Antipolis

(1990).

[4.42] JAWOROWSKI, Z., Radioactive lead in the environment and in the human body, At.

Energy Rev.

1 (1969) 3–45.

[4.43] SMITH-BRIGGS, J.L., Natural radionuclides near a coal-fired power station, Nucl.

Instrum. Meth. Phys. Res.

223 (1984) 590–592.

[4.44] KHANDEKAR, R.N., Polonium-210 in Bombay diet, Health Phys. 33 (1977) 148–150.

[4.45] LADINSKAYA, L.A., PARFENOV, Yu.D., POPOV, D.K., FEDOROVA, A.V.,

210

and

210

Po content in air, water, foodstuffs, and the human body, Arch. Environ.

Health

27 (1973) 254–258.

[4.46] GLOBEL, B., MUTH, H., “Natürliche Radioaktivität in Trinkwasser, Nahrungsmitteln

und im Menschen in Deutschland”, The Radiological Burden of Man from Natural

Radioactivity in the Countries of the European Communities (Proc. Sem. Paris, 1979),

Commission of the European Communities, Luxembourg (1979)

385–418.

[4.47] BUNZL, K., KRACKE, W., KREUZER, W.,

210

Pb and

210

Po in liver and kidneys of

cattle, I.

Animals from an area with little traffic or industry, Health Phys. 37

(1979)

323–330.

[4.48] IZAK-BIRAN, T., et al., Concentrations of U and Po in animal feed supplements, in

poultry meat and in eggs, Health Phys.

56 (1989) 315–319.

CHAPTER 4

[4.49] HILL, C.R., Identification of α emitters in normal biological materials, Health Phys. 8

(1962)

17–25.

[4.50] SMITH-BRIGGS, J.L., BRADLEY, E.J., POTTER, M.D., The ratio of lead-210 to

polonium-210 in UK diet, Sci. Total Environ.

54 (1986) 127–133.

[4.51] MARTIN, P., HANCOCK, G.J., JOHNSTON, A., MURRAY, A.S., Natural-series

radionuclides in traditional North Australian Aboriginal foods, J. Environ. Radioact.

(1998)

37–58.

[4.52] BACON, M.P., “Reactive radionuclides as tracers of oceanic particle flux”, Marine

Radioactivity (LIVINGSTON, H.D., Ed.), Elsevier, Amsterdam (2004)

139–164.

[4.53] CARVALHO, F.P., Distribution, cycling and mean residence time of

226

Ra,

210

Pb and

210

Po in the Tagus estuary, Sci. Total Environ. 196 (1997) 151–161.

[4.54] CARVALHO, F.P., Polonium (

210

Po) and lead (

210

Pb) in marine organisms and their

transfer in marine food chains, J. Environ. Radioact. 102 (2011) 462–472.

[4.55] CARVALHO, F.P., FOWLER, S.W., An experimental study on the bioaccumulation

and turnover of polonium-210 and lead-210 in marine shrimp, Mar. Ecol. Prog.

Ser.

102 (1993) 125–133.

[4.56] CARVALHO, F.P., FOWLER, S.W., A double-tracer technique to determine the

relative importance of water and food as sources of polonium-210 to marine prawns

and fish, Mar. Ecol. Prog. Ser.

103 (1994) 251–264.

[4.57] STEWART, G.M., et al., Contrasting transfer of polonium-210 and lead-210 across

three trophic levels in marine plankton, Mar. Ecol. Prog. Ser.

290 (2005) 27–33.

[4.58] WILDGUST, M.A., McDONALD, P., WHITE, K.N., Assimilation of

210

Po by the

mussel Mytilus edulis from the alga Isochrysis galbana, Mar. Biol.

136 (2000) 49–53.

[4.59] GODOY, J.M., SICILIANO, S., DE CARVALHO, Z.L., DE MOURA, J.F., GODOY,

M.L.D.P.,

210

Polonium content of small cetaceans from Southeastern Brazil, J. Environ.

Radioact. 106 (2012) 35–39.

[4.60] KIM, G., HUSSAIN, N., CHURCH, T.M., Excess

210

Po in the coastal atmosphere,

Tellus

52B (2000) 74–80.

[4.61] KIM, G., KIM, S.J., HARADA, K., SCHULTZ, M.K., BURNETT, W.C., Enrichment

of excess

210

Po in anoxic ponds, Environ. Sci. Technol. 39 (2005) 4894–4899.

[4.62] YANKOVICH, T.L., et al., Whole-body to tissue concentration ratios for use in biota

dose assessments for animals, Radiat. Environ. Biophys.

49 (2010) 549–565.

[4.63] RYAN, B., BOLLHÖFER, A., MARTIN, P., Radionuclides and metals in freshwater

mussels of the upper South Alligator River, Australia, J. Environ. Radioact.

(2008)

509–526.

[4.64] HOWARD, B.J., et al., The IAEA handbook on radionuclide transfer to wildlife,

J. Environ. Radioact. 121 (2013) 55–74.

[4.65] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer to Wildlife, Technical Reports Series

No.

479, IAEA, Vienna (2014).

[4.66] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Sources and Effects of Ionizing Radiation, UNSCEAR 2008 Report to

the General Assembly, with Scientific Annexes, Vol. II: Effects, United Nations,

New York (2008) Annex E.

OCCURRENCE AND CYCLING IN THE ENVIRONMENT

[4.67] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Sources and Effects of Ionizing Radiation, UNSCEAR 2000 Report to

the General Assembly, with Scientific Annexes, Vol.

I: Sources, United Nations,

New

York (2000) Annex B.

[4.68] CARVALHO, F.P.,

210

Po and

210

Pb intake by the Portuguese population: The

contribution of seafood in the dietary intake of

210

Po and

210

Pb, Health Phys. 69

(1995)

469–480.

[4.69] LEE, C.W., et al., Assessment of

210

Po in foodstuffs consumed in Korea, J. Radioanal.

Nucl. Chem.

279 (2009) 519–522.

[4.70] AARKROG, A., et al., A comparison of doses from

137

Cs and

210

Po in marine food: A

major international study, J. Environ. Radioact.

34 (1997) 69–90.

[4.71] CHAMBERLAIN, A.C., Emissions from Sellafield and activities in soil, Sci. Total

Environ.

177 (1996) 259–280.

[4.72] GARLAND, J.A., WAKEFORD, R., Atmospheric emissions from the Windscale

accident of October

1957, Atmos. Environ. 41 (2007) 3904–3920.

[4.73] MOORE, H.E., MARTELL, E.A., POET, S.E., Sources of polonium-210 in

atmosphere, Environ. Sci. Technol.

10 (1976) 586–591.

[4.74] RUTHERFORD, P.M., DUDAS, M.J., SAMEK, R.A., Environmental impacts of

phosphogypsum, Sci. Total Environ.

149 (1994) 1–38.

[4.75] INTERNATIONAL ATOMIC ENERGY AGENCY, Extent of Environmental

Contamination by Naturally Occurring Radioactive Material (NORM) and

Technological Options for Mitigation, Technical Reports Series No.

419, IAEA,

Vienna

(2003).

[4.76] ROESSLER, C.D., SMITH, Z.A., BOLCH, W.E., PRINCE, R.J., Uranium and

radium-226 in Florida phosphate materials, Health Phys.

37 (1979) 269–277.

[4.77] HURST, F.J., ARNOLD, W.D., A discussion of the uranium control in phosphogypsum,

Hydrometall.

9 (1982) 69–82.

[4.78] VAN DER HEIJDE, H.B., KLIJN, P.-J., PASSCHIER, W.F., Radiological impacts of

the disposal of phosphogypsum, Radiat. Prot. Dosim.

24 (1988) 419–423.

[4.79] SANTOS, A.J.G., MAZZILLI, B.P., FÁVARO, D.I.T., SILVA, P.S.C., Partitioning of

radionuclides and trace elements in phosphogypsum and its source materials based on

sequential extraction methods, J. Environ. Radioact.

87 (2006) 52–61.

[4.80] AOUN, M., EL SAMRANI, A.G., LARTIGES, B.S., KAZPARD, V., SAAD, Z.,

Releases of phosphate fertilizer industry in the surrounding environment: Investigation

on heavy metals and polonium-210 in soil, J. Environ. Sci.

22 (2010) 1387–1397.

[4.81] CARVALHO, F.P.,

210

Pb and

210

Po in sediments and suspended matter in the Tagus

estuary, Portugal: Local enhancement of natural levels by wastes from phosphate ore

processing industry, Sci. Total Environ.

159 (1995) 201–214.

[4.82] BETTI, M., et al., Results of the European Commission Marina II Study — Part II:

Effects of discharges of naturally occurring radioactive material, J. Environ.

Radioact.

74 (2004) 255–277.

[4.83] BERETKA, J., “The current state of utilisation of phosphogypsum in Australia”,

Phosphogypsum (Proc. Third Int. Symp. Orlando, 1990), Publication No. 01-060-083,

Vol. I, Florida Institute of Phosphate Research, Bartow, FL (1990) 394–401.

CHAPTER 4

[4.84] BOU-RABEE, F., AL-ZAMEL, A.Z., AL-FARES, R.A., BEM, H., Technologically

enhanced naturally occurring radioactive materials in the oil industry (TENORM): A

review, Nukleonika

54 (2009) 3–9.

[4.85] NORWEGIAN RADIATION PROTECTION AGENCY, Natural Radioactivity in

Produced Water from the Norwegian Oil and Gas Industry in 2003, StrålevernRapport

2005:2, Østerås, NRPA, Oslo

(2004).

[4.86] MORA, J.C., ROBLES, B., CORBACHO, J.A., GASCO, C., GASQUEZ, M.J.,

Modelling the behaviour of

210

Po in high temperature processes, J. Environ.

Radioact.

102 (2011) 520–526.

[4.87] PORCELLI, D., “Investigating groundwater processes using U- and Th- series

nuclides”, U-Th Series Radionuclides in Aquatic Systems, Radioactivity in the

Environment, Vol.

13 (KRISHNASWAMI, S., COCHRAN, J.K., Eds), Elsevier,

Amsterdam (2008)

105–153.

[4.88] ABDELOUAS, A., Uranium mill tailings: Geochemistry, mineralogy, and

environmental impact, Elements

2 (2006) 335–341.

[4.89] INTERNATIONAL ATOMIC ENERGY AGENCY, The Long Term Stabilization of

Uranium Mill Tailings, IAEA-TECDOC-1403, IAEA, Vienna

(2004).

[4.90] INTERNATIONAL ATOMIC ENERGY AGENCY, The Environmental Behaviour of

Radium: Revised Edition, Technical Reports Series No. 476, IAEA, Vienna (2014).

[4.91] BAXTER, M.S., MacKENZIE, A.B., EAST, B.W., SCOTT, E.M., Natural decay series

radionuclides in and around a large metal refinery, J. Environ. Radioact.

(1996)

115–133.

[4.92] BAXTER, M.S., Technologically enhanced radioactivity: An overview, J. Environ.

Radioact.

32 (1996) 3–17.

[4.93] SELBY, J.H., “The industrial uses of zircon and zirconia and the radiological

consequences of these uses”, Naturally Occurring Radioactive Material (NORM

V),

(Proc. Int. Symp. Seville, 2007), IAEA, Vienna (2007)

95–116.

Chapter 5

210

Pb AND

210

Po IN ATMOSPHERIC SYSTEMS

5.1. INTRODUCTION TO THE ATMOSPHERIC ENVIRONMENT

The depth of the gaseous envelope surrounding Earth is not firmly fixed but

is estimated to extend, at a maximum, approximately 1000 km from the surface

of the crust. Despite comprising less than 15% of the total planetary radius, the

nature of the atmospheric system as a whole has an overwhelming influence on

almost all the biological and geochemical conditions on Earth. Not only does it

affect human exposure to external influences and how its processes occur, it also

influences many of the Earth’s physical characteristics.

The temperature and pressure of the atmosphere vary considerably

with height above the surface. Pressure decreases exponentially with height,

and altitude is often given in units of pressure (hPa) rather than distance. The

temperature variation is more complex and is a crucial determinant of atmospheric

mixing processes. The rate of decrease in temperature with height is termed the

temperature lapse rate (°C/km), whose sign changes several times with height,

giving rise to identifiable layers in the atmosphere (see Fig. 5.1). In the lowest

of these, the troposphere, the lapse rate is normally positive (i.e. temperature

decreases with height), resulting in rapid vertical mixing. In the stratosphere, the

lapse rate is zero or negative due to absorption of solar ultraviolet radiation by

ozone, and vertical mixing is slow. Above the stratosphere lie the mesosphere,

thermosphere and exosphere. Together, the troposphere and stratosphere

comprise the lower atmosphere, which is the region of interest here.

The actual temperature variations with height are crucial determinants

of atmospheric mixing processes and are more complex than those shown in

Fig. 5.1. They depend not only on the time of the year but also on variations

in the structure of the atmosphere: for example, the tropopause (the boundary

between the troposphere and stratosphere) is around 8 km over the poles to but

16 km over the tropics.

The troposphere can be further divided into the planetary boundary layer

(PBL) and the free atmosphere. The behaviour of the PBL is strongly influenced

by the Earth’s surface and the wind within it is affected by surface drag.

Furthermore, turbulence is common and vertical mixing is generally quite strong.

However, a negative temperature lapse rate can exist within the PBL, leading to

a layer of stable air above the surface called an inversion layer. This occurs when

the Earth’s surface is cooler than the air above, such as on nights with clear skies.

CHAPTER 5

The height of the PBL is about 1 km, although this depends on surface features

and atmospheric conditions.

In the atmosphere,

210

Bi,

210

Pb and

210

Po are attached to aerosol particles,

with a diameter distribution typically covering several orders of magnitude, from

around 10

−3

μm to 10

μm. The behaviour of these radionuclides reflect those

of the particles to which they are attached. Airborne particles may be broadly

classified into coarse particles (>1 μm) and fine particles (<1 μm). Fine particles

may be further categorized into Aitken nuclei (primary aerosols: 10

−3

–10

−1

μm)

and particles in the accumulation range (10

−1

–10

μm). These size range

classifications are approximate and vary depending on the situation [5.2, 5.3].

Coarse particles are primarily generated by mechanical means (e.g. dust

resuspension by wind), whereas fine particles originate from processes such as

condensation of gases. Aitken nuclei form particles in the accumulation range

through condensation and coagulation. Sedimentation (gravitational settling) is

a significant removal mechanism for coarse particles, but this is not the case for

fine particles.

The size distribution of particles to which radionuclides are attached

depends on several factors, including the source of the radionuclide and the

length of time it has been suspended. For example, airborne

210

Pb originating

from resuspension of dust will be predominantly on coarse particles, whereas that

Source: See Ref. [5.1].

FIG. 5.1. Vertical structure of the lower atmosphere.

ATMOSPHERIC SYSTEMS

produced by the decay of free

222

Rn present in the atmosphere is predominantly

on fine particles. In the latter case, on initial production, a large proportion of

the

210

Pb is attached to Aitken particles, but with time the proportion in the

accumulation mode increases.

5.2. SOURCES OF

210

Pb AND

210

Po IN THE ATMOSPHERE

The primary source of

210

Pb in the atmosphere is

222

Rn exhalation from the

Earth’s surface (mainly from soil) and its subsequent decay via short lived radon

progeny to

210

Pb. There are several other sources, including volcanic activity,

resuspension of soils, forest and savanna fires, and combustion of fossil fuel.

However, these other sources have been estimated to contribute less than 5% to

the total

210

Pb budget.

Exhalation of

222

Rn from the Earth’s surface with subsequent decay through

210

Pb to its progeny is one of the primary sources of

210

Po in the atmosphere.

However, due to the relative volatility of polonium, some other sources are

important. In particular, volcanic activity has been estimated to be as important

a contributor to the total budget as the exhalation of

222

Rn. Each of these sources

has different characteristics in terms of spatial and temporal variability, and

each has a different influence on the

210

Po:

210

Pb ratio and on the aerosol size

distribution to which

210

Pb and

210

Po are attached [5.4]. Overall global fluxes for

210

Pb and

210

Po have been estimated to be approximately 3.64 × 10

Bq/a and

3.82

× 10

Bq/a, respectively [5.5].

5.2.1. Radon exhalation

Radon exhalation flux densities from soil are dependent on a number of

factors, in particular the

226

Ra activity concentration in the soil, soil moisture,

bulk density and porosity

[5.6]. Although flux densities vary over several

orders of magnitude, sites where they are highly elevated are small on a

regional or global scale, and for the majority of soils, they are in the range of

10–50

mBq·m

−2

·s

−1

. Radon exhalation flux densities from the ocean surfaces

have been estimated to be about two orders of magnitude lower than those

from ice-free land surfaces

[5.5, 5.7]. As a result of these factors, flux densities

are highly variable over the Earth’s surface, and the major influences are the

proportion of ice-free land followed by soil properties, including the radium

activity concentration

[5.8].

Figure 5.2 shows the latitudinal dependence of radon flux densities, based

on the flux model of Schery and Wasiolek

[5.7]. In this plot, the latitude axis

is scaled such that the area enclosed by the curve between any two latitudes is

CHAPTER 5

proportional to the radon flux (Bq/s) for that area [5.10]. The region between the

Tropics of Cancer and Capricorn contributes around 38% to the total flux. Mean

latitudinal exhalation

222

Rn flux densities drop dramatically south of 30°S and

north of 70°N owing to the presence of large oceans and ice-covered land areas.

The southern hemisphere contributes only about 26% of the Earth’s

total radon flux. The radioactive half-life of radon (3.82

days) and the mean

lifetime of

210

Pb in the PBL, about 5–7 days [5.11], are considerably smaller

than the timescale required for mixing of air between the northern and southern

hemispheres. As a consequence, the lower radon flux results in lower average

210

Pb and

210

Po concentrations in the air, and lower deposition fluxes to the

Earth’s surface in the southern hemisphere

[5.12].

The radon exhalation flux density for ice-free soil is affected by a number

of factors. Of the temporally variable factors, soil moisture has the greatest

influence

[5.13]. The relationship is complex, but rainfall generally results in

a reduction in exhalation. Most studies of seasonal variability show reductions

in the wetter months: winter in temperate zones

[5.14] and wet monsoon in the

Mean latitudinal

222

Rn exhalation flux density

(mBq m

-2

-1

)

0 5 10 15 20 25

Latitude (degrees)

-10

-20

-30

-40

-50

-90

Source: Figure 1 of Ref. [5.9].

FIG. 5.2. Estimated mean annual latitudinal radon exhalation flux density.

ATMOSPHERIC SYSTEMS

tropics [5.15]. However, the variability is relatively small over an interannual

time frame.

Following exhalation, radon decays in the atmosphere to a series of short

lived progeny (

218

Po,

214

Pb,

214

Bi and

214

Po), which attach to fine (submicron)

aerosol particles. The activity concentrations of these short lived progeny have

been observed to be high, resulting in increased gamma dose rates close to the

ground during, and immediately following, rainstorms

[5.16, 5.17].

5.2.2. Volcanic activity

Volcanic activity releases thorium and uranium series radionuclides to the

atmosphere, including

210

Pb,

210

Po and

222

Rn, with

210

Po in excess over

210

in volcanic plumes due to its volatility

[5.18, 5.19]. The excess is sufficiently

high for

210

Po to be used as a tracer for long range transport of volcanic

plumes

[5.20]. Measurements close to Mount Sakurajima, Japan, show that

210

is predominantly attached to fine particles smaller than 2

μm [5.21].

The fluxes of

210

Pb and

210

Po to the atmosphere from this source have

been estimated to be 6.0

× 10

Bq/a and 2.4 × 10

Bq/a, respectively [5.22].

In the case of

210

Po, this is of a similar order of magnitude to that from ingrowth

following

222

Rn exhalation from soil. However, this flux is difficult to quantify

accurately. Moreover, it greatly varies depending on the timing and location

of volcanic activity. One approach is to use of sulphate measurements as an

indicator for the influence of volcanic activity, in conjunction with

210

Pb and

210

Po measurements [5.23]. Volcanic eruptions can also lead to an injection

210

Pb and

210

Po into the stratosphere; in a 1979 study, Lambert et al. [5.24]

estimate this flux to be about 2% of that into the troposphere.

5.2.3. Resuspended soil dust

The suspension of dust can in some cases be a significant contributor to

concentrations of

210

Pb and its progeny in the atmosphere. This is especially

true in the PBL over continents. The sources often contribute only on a local or

regional scale, owing to the significant role sedimentation plays in the removal of

resuspended particles to the surface. Nevertheless dispersion can be significant,

even over relatively large distances

[5.20].

Globally, the flux has been estimated at 3.3

× 10

Bq/a for

210

Pb and

210

Po [5.5, 5.25], which means it contributes around 1% and 10%, respectively,

of the total budgets from all sources. However, because it primarily affects the

PBL above significant land masses, it can be a very significant contributor to the

observed concentrations at ground level.

CHAPTER 5

The isotopes

210

Bi,

210

Pb and

210

Po are normally in secular equilibrium in

topsoils, and

210

Pb is also normally in excess over its progeny in the troposphere.

Consequently, the resuspension of dust has the effect of increasing the

210

Bi:

210

and

210

Bi:

210

Po ratios. This can contribute to difficulties with using

210

Pb–

210

and

210

Pb–

210

Po couples to calculate atmospheric

210

Pb residence times

(see Section 5.4). One approach to counter this is to use a determination of a

member of the uranium decay chain above

210

Pb to apply a correction [5.26, 5.27].

5.2.4. Biomass burning

Polonium-210 has been shown to be enriched in fire plumes, due to the

volatility of elemental polonium and polonium compounds at temperatures in the

low hundreds of degrees Celsius

[5.28]. Estimates of total worldwide biomass

burning are around (2–5) × 10

kg CO

released per year [5.29], and the carbon

content of dried plant materials is around 45–50%

[5.30]. As noted in Chapter 4,

concentrations of

210

Po in vegetable matter vary considerably, but an indicative

average value is 10

Bq/kg. This would give a total annual flux of

210

Po from this

source of around 2

× 10

Bq/a if all of the available

210

Po were released. This

release would be at ground level, and the

210

Po would be available for removal by

wet and dry deposition processes.

On a global and time averaged scale, biomass burning should not be a

significant source of

210

Po (nor of

210

Pb). Although in the case of large burning

events, it can have a local or regional effect

[5.20, 5.31]. The convective draught

from fires can also transport

222

Rn to the free troposphere, with subsequent decay

210

Pb [5.32]. Biomass burning is discussed in more detail in Appendix I V.

5.2.5. Wind blown sea spray and emission of volatile polonium compounds

from water surfaces

The isotopes

210

Pb and

210

Po are enriched in the ocean surface microlayer,

with the

210

Po:

210

Pb activity ratio at around 2 [5.33]. Wind blown sea spray has

been shown to be a minor contributor to the atmospheric budget, estimated at

3.7

× 10

Bq/a and 7.4 × 10

Bq/a for

210

Pb and

210

Po, respectively [5.5, 5.25].

In addition, there is evidence that volatile compounds, such as dimethyl polonide,

can be formed and released from ocean and fresh water surfaces

[5.34–5.36].

5.2.6. Anthropogenic sources

There are many anthropogenic sources of

210

Pb and

210

Po. The most

significant are fossil fuel (particularly coal) burning and dispersion of

phosphate fertilizers and gypsum by-products. Global fluxes are estimated to be

ATMOSPHERIC SYSTEMS

8.2 × 10

Bq/a from fossil fuels for both

210

Pb and

210

Po, and 2.0 × 10

Bq/a and

× 10

Bq/a from phosphate for

210

Pb and

210

Po, respectively [5.5]. In addition,

much of the world’s biomass burning is anthropogenic (see Section 5.2.4).

Fossil fuel burning and most other anthropogenic sources impact most

heavily on concentrations in urban areas. A study in Seoul, Republic of

Korea

[5.37], concludes that the dominant fraction of

210

Po in precipitation

samples is linked to anthropogenic sources, primarily from the burning of coal,

with minor contributions from biomass burning.

5.3. PROCESSES FOR THE REMOVAL OF

210

Pb AND

210

Po FROM THE

ATMOSPHERE

Radioactive decay is only a minor contributor to the removal of

210

Pb and

210

Po on account of their short residence times in the troposphere. The primary

removal pathway is wet deposition to the Earth’s surface, with a secondary

pathway being dry deposition. Using the estimates of global source terms for

210

and

210

Po given in Section 5.2, the average global deposition flux density over

the Earth’s surface should be approximately 71

Bq·m

−2

·a

−1

and 7.5 Bq·m

−2

·a

−1

for

210

Pb and

210

Po, respectively. However, annual total deposition flux densities

for

210

Pb range from several Bq·m

−2

·a

−1

(as observed in Antarctica) to several

hundred Bq·m

−2

·a

−1

5.3.1. Wet deposition

Wet deposition refers to removal of atmospheric constituents to the ground

surface by precipitation (e.g. rain, snow and hail). Globally, wet deposition is

the primary process for removal of

210

Pb and

210

Po from the troposphere, and

occurs mainly by precipitation from either stratiform or convective clouds. The

removal of

210

Pb and

210

Po essentially reflects the removal of the aerosol particles

to which they are attached. The term rainout refers to the removal of aerosol

which has been scavenged within the cloud, while washout refers to scavenging

which occurs below the cloud during the precipitation event.

The concentrations of

210

Pb and

210

Po in precipitation are a complex result

of several factors and include the following:

— Their concentrations in the air column;

— The size distribution of the aerosol to which they are attached;

— The effective height of the precipitating column;

— The relative contributions from rainout and washout;

CHAPTER 5

— The growth in raindrop size within the cloud due to ascent in convective

updrafts;

— Evaporation of the raindrop during its descent;

— The type and duration of rainfall.

Reported values for

210

Pb concentration in rainwater vary

considerably at different locations, being most commonly in the range of

10–1000

mBq/L

with

210

Po:

210

Pb activity ratios of 0.005–1 or even greater.

Some measured concentrations in rainwater are provided in Table 5.1, showing

the variability in concentrations between locations and over extended periods at

one location.

In many cases,

210

Pb and

210

Po activity concentrations are greater at the start

of the precipitation event than later in the event, probably primarily due to the

reduction in washout as the air below the cloud becomes depleted

[5.27, 5.43].

A related effect is that activity concentrations in the precipitation are

generally inversely proportional to the amount of precipitation (e.g. see

Refs

[5.39, 5.40, 5.44]).

TABLE 5.1.

210

Pb AND

210

Po CONCENTRATIONS AND

210

Po:

210

ACTIVITY RATIOS IN RAINWATER

Location

210

Pb (mBq/L)

210

Po (mBq/L)

210

Po:

210

Ref.

Range Avg. Range Avg. Range Avg.

Izmir, Turkey 9–198 51 2–35 8 0.03–1.09 —

[5.38]

Geneva,

Switzerland

60–3300 —

—

[5.39]

Jabiru, Australia 46–1700 147 0.9–790 20 0.02–0.46 0.14 [5.27]

Galveston, USA 34–3600 132 —

—

[5.40]

Detroit, USA 105–1233 472 2.2–126 23 0.005–

0.283

0.049 [5.41]

Monaco —

—

0.25 [5.42]

—: data not available.

ATMOSPHERIC SYSTEMS

5.3.2. Dry deposition

Globally, dry deposition accounts for about 10–15% of total deposition for

210

Pb [5.5, 5.45]. However, this proportion varies spatially and temporally. The

contribution from dry deposition is greater for arid regions and can exceed the

wet fallout in these areas

[5.46].

Dry deposition of

210

Pb and

210

Po consists of two components: activity

present on resuspended dust; and activity from other sources and primarily

attached to fine particles. Deposition of activity on resuspended dust will be more

strongly influenced by sedimentation, although under some conditions such dust

can travel thousands of kilometres. Cases of such long distance transport have

been well documented for dusts originating from subtropical arid regions such as

North

Africa and the Middle East, Central Asia and Australia [5.20].

On its resuspension, the

210

Po:

210

Pb activity ratio of dust can be expected

to be approximately one. Since coarse particles may dominate the dry

deposition, this can result in a significantly higher

210

Po:

210

Pb ratio than in wet

deposition

[5.41].

5.3.3. Deposition fluxes

Figure 5.3 shows the results of an aerosol transport model simulation for

deposition flux density of

210

Pb on a latitudinal basis [5.47]. There is a peak in

deposition from convective rain in the tropics. The greater deposition flux in the

northern hemisphere compared with the southern hemisphere is due to the higher

222

Rn source term (see Fig. 5.1 and Section 5.4.1). The very low deposition flux

in Antarctica is due to a combination of the low southern hemisphere source

term, the low precipitation rate, the lack of significant ice-free land masses close

to Antarctica, and the band of strong westerly winds around 40–50°S, which act

as a barrier to low level movement of air from the land masses to the north.

As discussed, wet deposition dominates the total deposition at most

locations, and Fig.

5.3 shows that this is the case for all latitudinal bands. The

annual wet deposition flux (Bq·m

−2

·a

−1

) of a radionuclide can be calculated from

the product of the total precipitation over the year and the (precipitation weighted)

average concentration in the rainwater. Hence, both the amount of precipitation

and the concentration are important in evaluating these fluxes.

An example of the importance of multiple factors in determining deposition

fluxes is the waters on the east of Japan, where

210

Pb deposition fluxes of up to

800

Bq·m

−2

·a

−1

have been recorded. These high fluxes are due to a combination

of air masses with high

210

Pb concentrations from the surface layer of the Asian

continent being moved by strong winter monsoons, mixing with ascending

convection clouds over the body of water between Japan and the Asian mainland,

CHAPTER 5

and subsequent precipitation (rainfall and snowfall) over the coast and mountain

ridge of Japan with high

210

Pb concentrations in the precipitation [5.44, 5.48].

High

210

Po deposition fluxes and elevated

210

Po:

210

Pb ratios have also been

recorded in this region

[5.49].

The annual amount of precipitation is usually a very strong determinant

of depositional fluxes. This is true both between locations, and for interseasonal

and interannual variability at the same location

[5.39, 5.50]. However, Beks et

al.

[5.51] report that the

210

Pb depositional fluxes at two sites in the Netherlands

were mainly correlated with the number of times that there was heavy rain

or thunderstorms, rather than annual rainfall. The deposition fluxes were

approximately 75

Bq·m

−2

·a

−1

, which is considerably lower than the studies in

Japan on account of the predominantly westerly winds.

5.4. DISTRIBUTION OF

210

Pb AND

210

Po IN THE ATMOSPHERE

5.4.1. Concentrations at the land surface

In general, the most important influences on concentrations observed at the

surface are the presence of significant land masses within the prevailing wind

direction (as a source of

222

Rn), the amount of resuspension of dust, the degree of

wet deposition occurring in the region and the degree of vertical mixing of the air

Source: Figure 3 of Ref. [5.47].

FIG. 5.3. Zonally averaged annual deposition flux of

210

Pb (Bq·m

−2

·a

−1

) due to convective

rain, synoptic rain and dry deposition (reproduced with permission courtesy of American

Geophysical Union).

ATMOSPHERIC SYSTEMS

mass (see Section 4.2). The presence of other sources, such as volcanic activity

or forest fires, can also be important, especially for

210

Po.

Figure 5.4 shows average annual surface air concentrations of

210

for a number of locations against latitude

[5.12]. As a first approximation, the

latitudinal variations in concentrations reflect the variations in the

222

Rn source

term, which in turn, reflect the distribution of ice-free land masses. However,

for any specific location, factors such as predominant wind direction, degree of

vertical mixing and removal by precipitation also play important roles.

The highest

210

Pb concentrations in surface air are observed in the

subtropical and temperate latitudes of the northern hemisphere owing to the

relatively large land masses there. Values for average annual concentrations in

this region mainly lie in the range of 0.2–1

mBq/m

[5.12]. Between the tropics,

values are generally in the range of 0.1–0.5

mBq/m

, while south of around 30°S,

they quickly trend down to less than 0.1

mBq/m

, owing to the lack of large

ice-free land masses and consistent with the radon exhalation flux density trends

already discussed.

210

Pb concentration in surface air (mBq SCM

-1

)

0.0 0.2 0.4 0.6 0.8 1.0 1.2

Latitude (degrees)

-10

-20

-30

-40

-50

-90

Polar regions

Americas

Atlantic Ocean

Europe, Africa

Asia

Indian Ocean

Australia

Pacific Ocean

Source: Figure 2.13 of Ref. [5.10].

Note: Each point represents a separate measurement location.

FIG. 5.4. Mean annual

210

Pb concentration in surface air as a function of latitude (reproduced

with permission courtesy of Elsevier).

CHAPTER 5

The concentrations in the high latitude regions of the northern hemisphere

are about an order of magnitude higher than over Antarctica, primarily due to the

proximity of significant land masses to the south, and low level transport from

mid-latitudes to the Arctic in winter and spring, coincident with a pollution event

known as Arctic haze

[5.12, 5.52].

The concentrations are generally lower for the island or coastal sites (open

symbols in Fig. 5.4) than for continental sites. Since Fig. 5.4 shows data for land

based monitoring stations, the latitudinal average concentrations can be expected

to be lower than indicated, with the island sites being more representative of the

concentration over ocean areas.

There is clear evidence of a seasonality effect in

210

Pb and

210

concentrations in surface air

[5.20, 5.53–5.55]. Concentrations are normally

higher during dry seasons due to higher exhalation of

222

Rn from the soil,

greater resuspension of soil dust and lower wet deposition rates, which can vary

locally. However, the effect of air mass origin can also play a significant role in

seasonality effects, depending on the local conditions

[5.56].

5.4.2. Vertical profiles

Mean vertical profiles of

222

Rn generally show a strong reduction in

concentration from the surface to the upper troposphere of a factor of 100 over

continental areas

[5.57–5.59]. This reflects the ground level source, and the fact

that radon is only removed by radioactive decay with a half-life of 3.8 days.

Nevertheless, vertical profiles are highly variable, reflecting geographical

variability in the source term, as well as vertical and horizontal mixing processes

in the PBL and the free atmosphere. Average profiles have been modelled using a

one dimensional turbulent diffusion equation

[5.60]. However, mixing processes

are much more dynamic than is implied by such models, including rapid

transport from the lower to the upper troposphere in convective clouds

[5.61].

Discontinuities have also been observed at the PBL–free atmosphere interface,

and between cloud layers and the above cloud layer

[5.59].

The vertical distributions of

210

Pb and

210

Po in the troposphere are

more complex than that of

222

Rn due to progressive ingrowth and removal by

entrainment in precipitation. For

210

Pb, the reduction between ground level

and upper tropospheric concentrations are generally less than a factor of

five, and the concentration in the upper troposphere can be higher than in the

cloud layer

[5.57, 5.62]. The reason for this is scavenging of

210

Pb by wet and

dry deposition in the lower troposphere, and by convective and stratiform

precipitation from the mid-layers

[5.53]. There is also a stratospheric contribution

resulting from stratosphere–troposphere exchange, which predominantly affects

upper tropospheric concentrations.

ATMOSPHERIC SYSTEMS

In oceanic and polar regions, the vertical distributions of

210

Pb and

210

Po can differ from those discussed above owing to the much reduced radon

exhalation flux densities at the surface. Their concentrations can increase with

altitude owing to a combination of the transport of radon to the free troposphere

due to convective events, with subsequent formation of

210

Pb [5.32, 5.63] and

contributions from stratospheric air

[5.52]. Elevated levels of

210

Pb in the free

troposphere over the South

Pacific are attributed to radon convection due to

fires, most likely in Africa

[5.32]. In a study of the western hemisphere Arctic,

210

Pb concentrations were elevated at 3–6 km compared to those below 1 km at

latitudes greater than 65°N

[5.64].

Average

210

Pb and

210

Po concentrations in the stratosphere are about

0.3

mBq/SCM (standard cubic metre) at 0°C and 1000 hPa, with relatively little

variation with altitude and latitude

[5.22, 5.65]. As discussed in Section 5.2,

some transfer of

210

Pb and

210

Po to the stratosphere can occur owing to

volcanic eruptions

[5.24]. Apart from this, direct transfers of

210

Pb and

210

to the stratosphere are negligible due to their removal by ice crystals formed in

tropospheric air entering the lower stratosphere [5.66]. Rather,

222

Rn present in

these air masses decays to

210

Pb, with substantial subsequent ingrowth of

210

due to the relatively long stratospheric residence times (~1 year). This transfer

occurs predominantly under conditions of strong convective activity, bringing

air from the lower troposphere with higher concentrations of radon and resulting

in sufficient radon being transferred to support the observed stratospheric

210

activity concentrations

[5.64]. Strong convective activity, as well as tropopause

folds, linked to the subtropical and polar jet streams, also result in the transfer of

stratospheric air to the troposphere

[5.67].

REFERENCES TO CHAPTER 5

[5.1] NATIONAL OCEANIC AND ATMOSPHERIC ADMINISTRATION, NATIONAL

AERONAUTICS AND SPACE ADMINISTRATION, UNITED STATES AIR

FORCE, US Standard Atmosphere, 1976, US Government Printing Office, Washington,

(1976).

[5.2] SEINFELD, J.H., PANDIS, S.N., “Properties of the atmospheric aerosol”, Atmospheric

Chemistry and Physics: From Air Pollution to Climate Change, 2nd edn, Wiley and

Sons, Hoboken, NJ (2006)

350–395.

[5.3] PAPASTEFANOU, C., “Radioactive aerosol analysis”, Handbook of Radioactivity

Analysis, 3rd edn (L’ANNUNCIATA, M., Ed.), Elsevier, Amsterdam (2012)

727–767.

[5.4] GAFFNEY, J.S., MARLEY, N.A., CUNNINGHAM, M.M., Natural radionuclides in

fine aerosols in the Pittsburgh area, Atmos. Environ.

38 (2004) 3191–3200.

[5.5] BASKARAN, M., Po-210 and Pb-210 as atmospheric tracers and global atmospheric

Pb-210 fallout: A review, J. Environ. Radioact.

102 (2011) 500–513.

CHAPTER 5

[5.6] INTERNATIONAL ATOMIC ENERGY AGENCY, Measurement and Calculation of

Radon Releases from NORM Residues, Technical Reports Series No.

474, IAEA,

Vienna

(2013).

[5.7] SCHERY, S.D., WASIOLEK, M.A., “Modelling

222

Rn flux from the earth’s surface”,

Radon and Thoron in the Human Environment (Proc. 7th Tohwa Univ. Int. Symp.

Fukuoka, 1997) (KATASE, A., SHIMO, M., Eds), World Scientific, Tokyo

(1998)

207–217.

[5.8] Sources and Measurements of Radon and Radon Progeny Applied to Climate and Air

Quality Studies (Proc. Technical Mtg Vienna, 2009), IAEA, Vienna

(2012).

[5.9] MARTIN, P., “A note on the contribution of tropical regions to the Earth’s radon flux”,

ibid., pp.

75–76.

[5.10] MARTIN, P., McBRIDE, J.L., “Radionuclide behaviour and transport in the tropical

atmospheric environment”, Tropical Radioecology, Radioactivity in the Environment,

Vol. 18 (TWINING, J., Ed.), Elsevier, Amsterdam (2012) 59–91.

[5.11] RANGARAJAN, C., EAPEN, C.D., The use of natural radioactive tracers in a study of

atmospheric residence times, Tellus

B 42 (1990) 142–147.

[5.12] PREISS, N., MÉLIÈRES, M.-A., POURCHET, M., A compilation of data on lead 210

concentration in surface air and fluxes at the air–surface and water–sediment interfaces,

J. Geophys. Res.

101 (1996) 28 847–28 862.

[5.13] JHA, S., KHAN, A.H., MISHRA, U.C., A study of the

222

Rn flux from soil in the U

mineralised belt at Jaduguda, J. Environ. Radioact.

49 (2000) 157–169.

[5.14] WHITTLESTONE, S., ZAHOROWSKI, W., SCHERY, S.D., Radon flux variability

with season and location in Tasmania, Australia, J. Radioanal. Nucl. Chem.

236

(1998)

213–217.

[5.15] LAWRENCE, C.E., AKBER, R.A., BOLLHÖFER, A., MARTIN, P., Radon-222

exhalation from open ground on and around a uranium mine in the wet-dry tropics, J.

Environ. Radioact.

100 (2009) 1–8.

[5.16] FUJINAMI, N., Observational study of the scavenging of radon daughters by

precipitation from the atmosphere, Environ. Int.

22 (1996) 181–185.

[5.17] STÖHLKER, U., BLEHER, M., CONEN, F., BÄNNINGER, D., “Harmonization of

ambient dose rate monitoring provides for large scale estimates of radon flux density

and soil moisture changes”, Sources and Measurements of Radon and Radon Progeny

Applied to Climate and Air Quality Studies (Proc. Technical Mtg Vienna, 2009),

IAEA, Vienna (2012)

53–63.

[5.18] SATO, J., Natural radionuclides in volcanic activity, Appl. Radiat. Isot. 58

(2003)

393–399.

[5.19] GAUTHIER, P.-J., LE CLOAREC, M.-F., CONDOMINES, M., Degassing processes

at Stromboli volcano inferred from short-lived disequilibria (

210

Pb–

210

Bi–

210

Po) in

volcanic gases, J. Volcanol. Geotherm. Res.

102 (2000) 1–19.

[5.20] NHO, E.-Y., ARDOUIN, B., LE CLOAREC, M.-F., RAMONET, M., Origins of

210

in the atmosphere at Lamto, Ivory

Coast: Biomass burning and Saharan dusts, Atmos.

Environ.

30 (1996) 3705–3714.

ATMOSPHERIC SYSTEMS

[5.21] ASHIKAWA, N., et al., The size distribution of

210

Po in the atmosphere around

Sakurajima in Kagoshima prefecture, Japan, J. Radioanal. Nucl. Chem. 230

(1998)

97–104.

[5.22] LAMBERT, G., ARDOUIN, B., POLIAN, G., Volcanic output of long-lived radon

daughters, J. Geophys. Res.

87 (1982) 11 103–11 108.

[5.23] SHENG, Z., KURODA, P.K., Atmospheric injections of polonium-210 from the recent

volcanic eruptions, Geochem. J.

19 (1985) 1–10.

[5.24] LAMBERT, G., BUISSON, A., SANAK, J., ARDOUIN, B., Modification of the

atmospheric polonium 210 to lead 210 ratio by volcanic emissions, J. Geophys. Res.

(1979)

6980–6986.

[5.25] ROBBINS, J.A., “Geochemical and geophysical applications of radioactive lead”, The

Biogeochemistry of Lead in the Environment (TRIAGE, J.O., Ed.), Elsevier,

Amsterdam (1978)

285–393.

[5.26] CARVALHO, F.P., Origins and concentrations of

222

Rn,

210

Pb,

210

Bi and

210

Po in the

surface air at Lisbon, Portugal, at the Atlantic edge of the European continental

landmass, Atmos. Environ.

29 (1995) 1809–1819.

[5.27] MARTIN, P., Uranium and thorium series radionuclides in rainwater over several

tropical storms, J. Environ. Radioact.

65 (2003) 1–18.

[5.28] LAMBERT, G., LE CLOAREC, M.-F., ARDOUIN, B., BONSANG, B., “Long-lived

radon daughters signature of Savanna Fires”, Global Biomass Burning (LEVIN, J.S.,

Ed.), MIT

Press, London (1991) 181–184.

[5.29] CRUTZEN, P.J., ANDREAE, M.O., Biomass burning in the tropics: Impact on

atmospheric chemistry and biogeochemical cycles, Science

250 (1990) 1669–1678.

[5.30] FRIEDL, A., PADOUVAS, E., ROTTER, H., VARMUZA, K., Prediction of heating

values of biomass fuel from elemental composition, Anal. Chim. Acta

544

(2005)

191–198.

[5.31] PAATERO, J., et al., Resuspension of radionuclides into the atmosphere due to forest

fires, J. Radioanal. Nucl. Chem.

282 (2009) 473–476.

[5.32] DIBB, J.E., et al., Constraints on the age and dilution of Pacific Exploratory Mission:

Tropics biomass burning plumes from the natural radionuclide tracer

210

Pb, J. Geophys.

Res.

104 (1999) 16 233–16 241.

[5.33] BACON, M.P., ELZERMAN, A.W., Enrichment of

210

Pb and

210

Po in the sea-surface

microlayer, Nature

284 (1980) 332–334.

[5.34] KIM, G., HUSSAIN, N., CHURCH, T.M., Excess

210

Po in the coastal atmosphere,

Tellus

52B (2000) 74–80.

[5.35] MOMOSHIMA, N., SONG, L.-X., OSAKI, S., MAEDA, Y., Biologically induced Po

emission from fresh water, J. Environ. Radioact.

63 (2002) 187–197.

[5.36] CHURCH, T.M., SARIN, M.M., “U- and Th-series nuclides in the atmosphere: Supply,

exchange, scavenging and applications to aquatic processes”, Radioactivity in the

Environment, Vol.

13 (KRISHNASWAMI, S., COCHRAN, J.K., Eds), Elsevier,

Amsterdam (2008)

11–45.

[5.37] YAN, G., CHO, H.-M., LEE, I., KIM, G., Significant emissions of

210

Po by coal

burning into the urban atmosphere of Seoul, Korea, Atmos. Environ.

54 (2012) 80–85.

CHAPTER 5

[5.38] UĞUR, A., ÖZDEN, B., FILIZOK, I., Determination of

210

Po and

210

Pb concentrations

in atmospheric deposition in İzmir (Aegean Sea-Turkey), Atmos. Environ.

(2011)

4809–4813.

[5.39] CAILLET, S., ARPAGAUS, P., MONNA, F., DOMINIK, J., Factors controlling

and

210

Pb atmospheric deposition as revealed by sampling individual rain events in the

region of Geneva, Switzerland, J. Environ. Radioact.

53 (2001) 241–256.

[5.40] BASKARAN, M., COLEMAN, C.H., SANTSCHI, P.H., Atmospheric depositional

fluxes of

Be and

210

Pb at Galveston and College Station, Texas, J. Geophys. Res. 98

(1993)

20 555–20 571.

[5.41] McNEARY, D., BASKARAN, M., Residence times and temporal variations of

210

Po in

aerosols and precipitation from southeastern Michigan, United States, J. Geophys.

Res.

112 (2007).

[5.42] TATEDA, Y., CARVALHO, F.P., FOWLER, S.W., MIQUEL, J.-C., Fractionation of

210

Po and

210

Pb in coastal waters of the NW Mediterranean continental margin, Cont.

Shelf Res.

23 (2003) 295–316.

[5.43] GAVINI, M.B., BECK, J.N., KURODA, P.K., Mean residence times of the long-lived

radon daughters in the atmosphere, J. Geophys. Res.

79 (1974) 4447–4452.

[5.44] HIROSE, K., KIKAWADA, Y., DOI, T., SU, C.-C., YAMAMOTO, M,

210

Pb deposition

in the far East Asia: Controlling factors of its spatial and temporal variations, J.

Environ. Radioact.

102 (2011) 514–519.

[5.45] BALKANSKI, Y.J., JACOB, D.J., GARDNER, G.M., GRAUSTEIN, W.C.,

TUREKIAN, K.K., Transport and residence times of tropospheric aerosols inferred

from a global three-dimensional simulation of

210

Pb, J. Geophys. Res. 98 (1993)

573–20 586.

[5.46] RASTOGI, N., SARIN, M.M., Atmospheric

210

Pb and

Be in ambient aerosols over

low- and high-altitude sites in semiarid region: Temporal variability and transport

processes, J. Geophys. Res.

113 (2008).

[5.47] GUELLE, W., BALKANSKI, Y.J., SCHULZ, M., DULAC, F., MONFRAY, P., Wet

deposition in a global size-dependent aerosol transport model: 1.

Comparison of a

year

210

Pb simulation with ground measurements, J. Geophys. Res. 103 (1998)

429–11 445.

[5.48] YAMAMOTO, M., et al., Seasonal and spatial variation of atmospheric

210

Pb and

deposition: Features of the Japan Sea side of Japan, J. Environ. Radioact.

(2006)

110–131.

[5.49] TATEDA, Y., IWAO, K., High

210

Po atmospheric deposition flux in the subtropical

coastal area of Japan, J. Environ. Radioact.

99 (2008) 98–108.

[5.50] BASKARAN, M., A search for the seasonal variability on the depositional fluxes of

Be and

210

Pb, J. Geophys. Res. 100 (1995) 2833–2840.

[5.51] BEKS, J.P., EISMA, D., VAN DER PLICHT, J., A record of atmospheric

210

deposition in the Netherlands, Sci. Total Environ.

222 (1998) 35–44.

[5.52] DIBB, J.E., et al., Estimation of stratospheric input to the Arctic troposphere:

Be and

Be in aerosols at Alert, Canada, J. Geophys. Res. 99 (1994) 12 855–12 864.

ATMOSPHERIC SYSTEMS

[5.53] ANAND, S.J.S., RANGARAJAN, C., Studies on the activity ratios of polonium-210

to lead-210 and their dry-deposition velocities at Bombay in India, J. Environ.

Radioact.

11 (1990) 235–250.

[5.54] MARENCO, A., FONTAN, J., Sources of polonium-210 within the troposphere,

Tellus

24 (1972) 38–46.

[5.55] WINKLER, R., HÖTZL, H., CHATTERJEE, B., Analysis of

210

Pb and

210

concentrations in surface air by an alpha spectrometric method, Health Phys.

(1981)

495–503.

[5.56] ALI, N., et al., The effect of air mass origin on the ambient concentrations of

Be and

210

Pb in Islamabad, Pakistan, J. Environ. Radioact. 102 (2011) 35–42.

[5.57] MOORE, H.E., POET, S.E., MARTELL, E.A.,

222

Rn,

210

Pb,

210

Bi and

210

Po profiles

and aerosol residence times versus altitude, J. Geophys. Res.

78 (1973) 7065–7075.

[5.58] LIU, S.C., McAFEE, J.R., CICERONE, R.J., Radon 222 and tropospheric vertical

transport, J. Geophys. Res.

89 (1984) 7291–7297.

[5.59] ZAHOROWSKI, W., et al., “

222

Rn observations for climate and air quality studies”,

Sources and Measurements of Radon and Radon Progeny Applied to Climate and Air

Quality Studies (Proc. Tech. Mtg Vienna, 2009), IAEA, Vienna (2012)

77–96.

[5.60] JACOBI, W., ANDRÉ, K., The vertical distribution of radon 222, radon 220 and their

decay products in the atmosphere, J. Geophys. Res.

68 (1963) 3799–3814.

[5.61] KRITZ, M.A., LE ROULLEY, J.-C., DANIELSEN, E.F., The China Clipper: Fast

advective transport of radon-rich air from the Asian boundary layer to the upper

troposphere near California, Tellus

42B (1990) 46–61.

[5.62] LIU, H., JACOB, D.J., BEY, I., YANTOSCA, R.M., Constraints from

210

Pb and

Be on

wet deposition and transport in a global three-dimensional chemical tracer model

driven by assimilated meteorological fields, J. Geophys. Res.

106 (2001)

109–12 128.

[5.63] GIANNAKOPOULOS, C., et al., A three-dimensional modelling study of the

correlations of

210

Pb with HNO

and peroxyacetylnitrate (PAN) at remote oceanic

sites, J. Geophys. Res.

105 (2000) 1947–1956.

[5.64] DIBB, J.E., TALBOT, R.W., GREGORY, G.L., Beryllium 7 and lead 210 in the

western hemisphere Arctic atmosphere: Observations from three recent aircraft-based

sampling programs, J. Geophys. Res.

97 (1992) 16 709–16 715.

[5.65] FEELY, H.W., SEITZ, H., Use of lead 210 as a tracer of transport processes in the

stratosphere, J. Geophys. Res.

75 (1970) 2885–2894.

[5.66] KRITZ, M.A., et al., Radon measurements in the lower tropical stratosphere: Evidence

for rapid vertical transport and dehydration of tropospheric air, J. Geophys. Res.

(1993)

8725–8736.

[5.67] BARAY, J.-L., ANCELLET, G., RANDRIAMBELO, T., BALDY, S., Tropical cyclone

Marlene and stratosphere–troposphere exchange, J. Geophys. Res.

104 (1999)

953–13 970.

Chapter 6

210

Pb AND

210

Po IN TERRESTRIAL SYSTEMS

6.1. DESCRIPTION OF THE TERRESTRIAL ENVIRONMENT

For the purposes of this chapter, the terrestrial environment is arbitrarily

defined as encompassing the following range of related components: the

atmosphere (particularly the troposphere); the lithosphere (particularly soils);

and the surface and groundwater with which they are associated. Within these

compartments, the terrestrial environment is also populated by a diverse range of

plants and animals. The exchanges within and between these compartments are

the environmental processes on which terrestrial ecosystems depend. However,

the interactions among the various abiotic and biotic environments vary from

location to location. The International Union of Radioecology (IUR) suggests

that interaction matrices could present information on relevant processes in the

terrestrial environments in a synthesized and structured way [6.1].

Each environment has its own unique chemical, physical and biological

characteristics. Generic interaction matrices developed by the IUR [6.1] provide

a convenient checklist for the initial evaluation of relevant environmental

processes. The IUR indicates that the major components of the terrestrial

environment closely relate to those discussed in the BIOMASS project [6.2]. In

the IUR matrices, however, plants and animals are treated as distinct components

rather than aggregated biota. Similarly, soils are subdivided into several distinct

components. The main components in any terrestrial system are thus the

following [6.1]:

(a) The atmosphere, particularly the troposphere, which comprises the region

above the soil surface, including the regions both above and below the plant

canopy.

(b) Surface aquatic bodies (i.e. rivers, streams and lakes) and groundwater,

including aquifers.

(d) Animals (macrofauna) in the terrestrial environment, inhabiting both the

soil surface and areas within the soil (i.e. burrowing animals).

(e) Soil microbiota, comprising microflora and protozoa but excluding larger

organisms such as earthworms and burrowing mammals.

CHAPTER 6

(f) The soil, comprising minerals and organic material, aqueous solutions and

the content of air in a soil sample. The characteristics of these various parts

are influenced by the biotic and abiotic conditions in operation.

(g) The interface with the geosphere. This is an artificial frontier. In terrestrial

environments, it will normally be represented by weathered, saturated or

unsaturated materials.

6.2. MAJOR TRANSFER PROCESSES AND PATHWAYS

The decay of

238

U is the ultimate source of

210

Pb and

210

Po (see also

Chapter 4). As a primordial radioisotope,

238

U is present in every component of

the environment to varying degrees. The decay of

238

U leads to the formation

of a series of different elements, including

210

Pb and

210

Po, each with its own

characteristic environmental behaviour. As environmental processes are always

dynamic, there may be disequilibrium between members of the decay series,

depending on how open the system is to external influences (e.g. see Ref.

[6.3]).

The major environmental processes which cause disequilibria and the transfer

210

Pb and

210

Po, or their relatively long lived progenitors (including

226

and

230

Th), from primary sources in the crust to other compartments include the

following:

— Weathering of exposed rocks;

— Leaching or dissolution of rocks by groundwater;

— Volcanic eruptions;

— Fires and burning fossil fuels;

— Exhalation of

222

Rn.

Given that

210

Pb and

210

Po are intrinsically linked by their radiogenic

association, a combined consideration of both of these radionuclides is necessary

to explain the observed patterns of

210

Po transfer in the environment (see Fig. 6.1).

The two main categories of biotic transfer are flora and fauna. Terrestrial

plants can be exposed to

210

Pb and

210

Po via their surface (bark and cuticle of

foliage), interaction with the atmosphere or via their roots, which can take up

the radionuclides from the soil solution. Atmosphere–plant interactions are

explored in Section

6.3. The mobility of radioisotopes in soil and their uptake by

plant roots is discussed in Sections

6.4 and 6.5, and

210

Po transfer in animals in

Section 6.6.

TERRESTRIAL SYSTEMS

6.3. ATMOSPHERE–PLANT INTERACTIONS

Polonium and lead activity concentrations on the surfaces of plants can be

affected by many processes, including:

(a) Interception of gases, aerosols or dusts directly from the atmosphere, or by

processes such as rain splash from the soil;

(b) Weathering, which results in the removal of radionuclides from plant

surfaces by various physical processes and comprises the loss of

radionuclides from the plant surface due to various processes, including

wash-off due to rain or irrigation and surface abrasion;

(d) Translocation of radionuclides from the plant surface to other, internal

tissues of the plant.

References [6.4, 6.5] provide detailed discussions about the concepts and

processes involved in air to plant interactions and how these processes can be

quantified. This section provides a broad overview of these ideas and identifies

information specific to polonium.

Plants can come into contact with

210

Pb and

210

Po by deposition of

aerosols from

222

Rn decay or from resuspended dust. The contribution of

atmospheric deposition to the total amount of radionuclides in and on plants

FIG. 6.1. Pathways controlling the movement and disposition of

210

Po in terrestrial

environments.

CHAPTER 6

is significant [6.6, 6.7] and can account for up to 95% [6.8]. The two basic

mechanisms governing the rate of radionuclide deposition from the atmosphere

to plants are wet and dry deposition processes. Wet deposition is the transport of

radionuclides from the atmosphere with precipitation (rain, snow or hail); and

dry deposition is a direct flux of radionuclides from the atmosphere in the form

of gases or particulates. Pietrzak-Flis and Skowrońska-Smolak [6.8] find that the

incorporation of the radionuclides occurs mainly as a result of wet deposition.

In addition to deposition, losses of radioactivity through weathering

influence the overall amount of radioactivity retained on the plant surface.

Pietrzak-Flis and Skowrońska-Smolak [6.8] measure these processes and

determine experimentally that the deposition of

210

Po onto a variety of grasses

and crops is in the range of 2–23 Bq/m

(6–62 Bq/m

in the case of

210

Pb), with

retention proportional to their surface area and the presence of surface features

(in this example possibly due to pockets on barley chaff, but may include factors

such as glabrous or hirsute leaves).

With regard to translocation of the deposited radioactivity, there is

unfortunately limited information available specifically on polonium, thus

restricting the capacity to quantitatively model the relevant foliar uptake

processes. If polonium specific data are absent, appropriate values might be

identified from suitable analogues

[6.9]. Although similar chemical properties of

different elements do not necessarily mean similar behaviour inside the plant, this

approach is still often used in practice, in the absence of other data. Analogues

used for polonium include sulphur, selenium and tellurium (see Section

3.2), all

of which are in Group

16 in the periodic table (chalcogens).

6.3.1. Interception

The extent to which plants can retain radionuclides depends on many

physical, chemical and biological factors. The simplest parameter used for

quantification is the interception fraction f, which is defined as the ratio of

the activity initially retained by the standing vegetation (immediately after

deposition) to the total activity deposited

[6.10]. Since the interception fraction

depends on the stage of development of the plant, the interception fraction is

normalized to the standing biomass B (dry mass) to yield the mass interception

fraction f

/kg) [6.10]. As the leaf is the main interface between the atmosphere

and vegetation, it is also possible to normalize the interception fraction to the leaf

area index f

LAI

[6.10].

There are no specific data on interception fractions for

210

Po. Using

as a chemically similar analogue, an f

of 0.35 was derived across five species

based on wet deposition

[6.11]. This value was slightly higher than that derived

TERRESTRIAL SYSTEMS

for

131

I (0.27) and was around three times lower compared with the other

radionuclides used in the experiment (

109

Cd,

144

Ce,

Cr and

Sr).

Particle size is another key parameter in determining the extent of dry

deposition along with the vegetation type (often characterized by surface

roughness). Interception is more effective for small particles and reactive

gases

[6.10]. No specific information on interception of polonium by dry

deposition has been identified in the literature. However, data on interception for

different particle sizes, across a range of radionuclides and vegetation types, can

be found in Ref. [6.10].

6.3.2. Physical and biological weathering

Weathering is the loss of contamination from plants. The magnitude of the

weathering loss of a radionuclide depends on many factors, including its solubility,

strength of adsorption to the plant surface, degree of penetration into the inner

flesh and leachability from the interior [6.12]. Biological factors (e.g. regular or

seasonal exfoliation) also play a part in the weathering process. Thus, a complex

interaction of those factors leads to the observed difference in weathering loss

among radionuclides, plant species and their growth stages [6.12]. Despite this

complexity, in models of environmental radionuclide behaviour, weathering is

normally described by a single exponential function characterized by a first order

rate constant λ

or a weathering half-life T

[6.12].

ln2

l =

(6.1)

Data from numerous studies have shown limited differences in weathering

between cationic species (such as caesium, cobalt, manganese, ruthenium and

strontium) for most plant species, and that T

is dependent on plant characteristics

as well as on plant growth stage at the time of deposition

[6.12]. In general,

therefore, for different elements and plant groups, an assumption can be made

that the weathering half-life for polonium will be in the range of 7.9–49

days (see

table 1 of Ref. [6.12]).

6.3.3. Foliar translocation

Translocation is the process by which elements, water and nutrients

are moved within vascular plants in the phloem and xylem, and can refer to

contaminants either taken up by plants roots or absorbed from the surface of

plants and transferred to other parts that have not been contaminated directly. The

rate of this process reflects the mobility of an element within the plant and can

CHAPTER 6

vary greatly. Translocation from roots is discussed in Section 6.5. To quantify this

process for material deposited directly onto the foliage, the translocation factor f

normally used represents the proportion of foliage activity per 1

of crop at the

time of deposition (Bq/m

) associated with the activity in the edible part per 1 m

of crop at harvest time (Bq/m

) expressed as a percentage [6.13].

There is a lack of information on polonium translocation values; hence, an

analogue could be considered here. However, there is no appropriate analogue

for translocation based on foliar deposition (f

). The only analogue data available

are the proportional translocation of tellurium in root vegetables, reported as

0.8

[6.14, 6.15]. However, Ham et al. [6.7] and Pietrzak-Flis and Skowrońska-

Smolak

[6.8] report that the movement of polonium from roots is negligible.

This highlights the uncertainty associated with the analogue approach, as the

experimental data for polonium have not indicated a translocation factor as high

as that reported for tellurium.

6.4. MOBILITY OF POLONIUM IN SOILS

Surface soils are the interface between the atmosphere, the parent rocks

and the subsurface hydrological system. There are radionuclide fluxes between

these environmental compartments which govern

210

Pb and

210

Po activity

concentrations and how they change in the surface soil. Generally,

210

Po is at, or

very close to, secular equilibrium with

210

Pb in almost all soils [6.16]. However,

the distribution of

210

Po in soil is rather inhomogeneous (see Chapter 4) [6.17].

The concentrations of radionuclides in surface soils vary depending on a number

of processes and are discussed in Sections 6.4.1–6.4.4.

6.4.1. Resuspension

In general, three types of process are used to describe the lateral spread of

contaminants which have been deposited onto surface soil

[6.18, 6.19]: surface

creep, saltation and resuspension. Resuspension occurs when wind exerts a

force exceeding the degree of adherence of particles to the surface material. The

forces in action are the weight of the particle and the adherence, as well as the

aerodynamic loads relating to the flow of wind. There are two basic approaches

used to model resuspension. The resuspension factor is the ratio of volumetric

air concentration to soil contamination. The resuspension rate approach is based

on the ratio of particle flow density to soil contamination, which depends on

the resuspended material (particle size, shape and adherence), the surface type

(roughness and humidity), the time lapsed after deposition, and mechanical

actions (soil processing), if any.

TERRESTRIAL SYSTEMS

No specific information is available for polonium. However, because the

variability of measured resuspension factors and rates are very high and the

accuracies of these models’ predictions are rather low, there is no substantial

evidence that the elemental characteristics of radionuclides influence the

processes governing resuspension. This means that the data and models validated

for other radionuclides can also be used for polonium. Overall, the resuspension

factors measured directly after the acute depositions are around 10

−5

/m in

residential areas, on a site undergoing cleanup operations and on an arid site, and

from 10

−8

/m to 10

−6

/m in humid areas. The resuspension factor decreases by up

to 10

−9

/m over 3–4 years after the deposition. Resuspended dust can be a very

localized phenomenon or can reach global scales, such as with the resuspension

of dusts from nuclear test sites in China to Japan and even further

[6.20].

6.4.2. Radon exhalation

One of the major mechanisms which determines changes in

210

concentrations (and hence

210

Po) in topsoil is exhalation of a fraction of the noble

gas

222

Rn from surface soils to the atmosphere, with subsequent decay to

210

in the atmosphere and removal to the Earth’s surface by wet and dry deposition

(see Section 5.3). Depending on the local soil and meteorological conditions,

these processes can result in either an excess or deficiency of

210

Pb relative to

226

Ra in the topsoil (primarily in the top 1 m of soil).

Since exhalation flux densities of

222

Rn from soils are about a hundred

times greater than that from oceans, deposition of soil-origin

210

Pb to the oceans

can be expected to cause a deficiency of

210

Pb and

210

Po relative to

226

Ra in

soils. Globally, approximately 90% of

210

Pb deposition is by wet deposition, and

local deposition greatly depends on rainfall

[6.21, 6.22]. A confounding factor

210

Pb in leaf litter, which can lead to excess

210

Pb and

210

Po in the surface

layer

[6.23, 6.24]. Owing to the permanent flux from the atmosphere and other

processes, activity concentrations in the topsoil (0–10

cm) tend to be higher

compared with the deeper soil, where

210

Po is in approximate equilibrium with

the

210

Pb and

222

Rn values, within a few tens of Bq/kg [6.17].

6.4.3. Partitioning of the soil Po:Pb ratio into solution

A portion of the radioactivity present in the soil will be adsorbed to soil

particles, while some will dissolve into soil water and hence become available for

root uptake by plants. Significant progress has been made to describe radionuclide

sorption in heterogeneous solids. Nevertheless, models describing sorption are

still mostly based on empirical, solid–liquid distribution coefficient (K

) values,

CHAPTER 6

which are estimated from sorption studies using reasonably well characterized

homogeneous surfaces (see Section 2.1).

A summary of available K

values for lead and polonium for soils, grouped

on the basis of texture and organic matter content criteria, as well as data on

the effect of pH (excluding organic soils), are given in Table 6.1. Soils are

grouped according to the sand and clay mineral percentages, and the organic

matter content of the soil. Some variation is due to pH (8%) and clay content

(24%) but not cation exchange capacity [6.25]. These numbers are estimates,

since adsorption is controlled by a larger range of conditions, including the

concentration of competing ions, the mineralogical and organic composition of

the soils, and the grain size of the constituents (and so the available surface area

for adsorption per gram).

TABLE 6.1. K

VALUES FOR SOILS GROUPED ACCORDING TO THE

TEXTURE/ORGANIC MATTER CRITERION (L/kg)

Element Soil group

GM GSD AM SD Min. Max.

Pb All soils 23 2.0 × 10

1.0 × 10

1.5 × 10

3.3 × 10

2.5 × 10

1.3 × 10

Sand 9 2.2 × 10

4 4.0 × 10

4.3 × 10

2.5 × 10

1.3 × 10

Loam 5 1.0 × 10

3 1.5 × 10

1.6 × 10

3.6 × 10

4.3 × 10

Clay 2 n.a. n.a. 6.6 × 10

n.a. 5.4 × 10

1.3 × 10

Organic 5 2.5 × 10

3 3.7 × 10

3.8 × 10

8.8 × 10

1.0 × 10

Unspecified 2 n.a. n.a. 5.9 × 10

n.a. 1.6 × 10

1.0 × 10

Po All soils 44 2.1 × 10

5 5.6 × 10

1.1 × 10

1.2 × 10

7.0 × 10

Sand 14 1.0 × 10

6 7.4 × 10

1.9 × 10

1.7 × 10

7.0 × 10

Loam 27 2.3 × 10

4 4.6 × 10

4.6 × 10

1.2 × 10

1.8 × 10

Clay 2 n.a. n.a. 1.7 × 10

n.a. 7.2 × 10

2.7 × 10

Organic 1 n.a. n.a. 6.6 × 10

n.a. n.a. n.a.

Source: See Ref. [6.5].

Note: The sand group is ≥65% sand and <18% clay; the clay group is ≥35% clay; the

organic group is ≥20% organic matter; and the rest is assigned to the loam group and

unspecified. AM — arithmetic mean; GM — geometric mean; GSD — geometric

standard deviation; n.a. — not applicable; SD — standard deviation.

TERRESTRIAL SYSTEMS

Table 6.1 shows that K

values for polonium tend to be lower than those for

lead. This difference can lead to some disequilibrium between these radionuclides

in the fraction assumed to be available for root transfer.

6.5.

210

Pb AND

210

Po TRANSFER FROM SOIL TO PLANTS

6.5.1. Processes governing

210

Pb and

210

Po transfer to plants

Polonium can enter plants via several routes. Most large scale contamination

is generally through air deposition or from contaminated soil. Contamination

can also occur via contaminated surface water (inundation and irrigation) or

contaminated groundwater (upwelling and irrigation)

[6.8, 6.26]. The relative

importance of these pathways depends on the concentration of the radionuclides

in the soil, the soil–plant concentration ratio (CR) and the rate of deposition onto

plant parts above ground. An important implication is that radionuclide activity

concentrations in crops such as in roots, tubers, cereals and legumes, where the

edible portion is protected by inedible plant parts, should not be significantly

affected by direct deposition; whereas the converse is true for leafy vegetables.

Plant type and the ecosystem influence the values thus derived. Transfer factors

(TFs) range across orders of magnitude, depending on the plant and soil type

variations assessed

[6.5].

6.5.2.

210

Pb and

210

Po transfer to mosses, plants and lichens

The factors that determine variability in radionuclide transfer to plants

generally including the following [6.27]:

(a) The form in which the activity enters or is present in the soil (e.g. as

particles, as aerosol or in solution);

(b) The physicochemical properties of the radionuclide;

(d) The type of soil and its physicochemical characteristics;

(e) The type of crop;

(f) Crop management practices (the application of fertilizers, irrigation,

ploughing and liming);

(g) Climatic conditions;

(h) The experimental conditions under which the TFs were obtained.

CHAPTER 6

The significance of these factors varies depending on the radionuclide of

interest and the contamination scenario. Table 6.2 presents data from the 1970s

on the influence of different steppe ecosystems on

210

Po transfer to plants.

From Table 6.2, the activity concentrations in the above ground biomass

are in the range of 3.3–16.3

Bq/kg (DW) for

210

Po and 9.3–49.6 Bq/kg (DW) for

210

Pb. For both

210

Pb and

210

Po, there are higher concentrations in the underground

biomass (roots). Maximum activity concentrations of

210

Pb and

210

Po occur in

perennial sawn grasses, probably because of the application of mineral fertilizers

containing a mixture of naturally occurring radionuclides

[6.29]. The

210

Po:

210

ratios in plants are lower than those for topsoils, indicating higher transfer rates

for

210

Pb compared with

210

Po [6.29].

Taskayev and Testov

[6.29] present data on

210

Po transfer to 24 species of

plants and mosses representative of northern taiga and collected from areas with

naturally elevated radiation backgrounds (see Table 6.3).

One type of vegetation of particular interest in the assessment of

210

Po is

lichen. Lichens are slow growing symbiotic associations (fungi plus algae or

cyanobacteria) that grow on trees, rocks and soils. They have no true roots and,

despite accessing some mineral nutrition from rocks or organic nutrients from

bark, are highly dependent on the atmosphere for nutrition. They have a long

lifespan; thus, activity concentrations of

210

Po are supported by

210

Pb ingrowth

in addition to direct deposition. Lichens are important for human populations in

northern climates because reindeer or caribou (which are key components of the

human diet) ingest large quantities in winter. The importance of lichens has not

been considered to the same extent for food chains involving herbivorous species

in other climatic zones.

In a 1978 study, Ermolayeva-Makovskaya and Litver

[6.30] report a high

mean value of 260

Bq/kg (DW) for both

210

Pb and

210

Po in lichens of the polar

and subpolar zones of the Russian Federation, from the Cola Peninsula to Chukot.

Slightly lower values (70–212

Bq/kg, DW) of

210

Po activity concentrations in

lichens in Nordic terrestrial ecosystems are presented in Refs [6.24, 6.31].

Data on lichens in northern Canada show mean activity concentrations of

275

Bq/kg (DW) for Cladina mitis and 622 Bq/kg (DW) for Cetraria

nivalis

[6.32]. The data indicate that the highest activity concentrations of

210

and

210

Po in above ground biomass occur in mosses and lichens, followed by

club-mosses and ferns, and then grasses.

6.5.3.

210

Pb and

210

Po transfer to berries, mushrooms and understory

species

Several comprehensive studies in Finland and France have greatly extended

the information on

210

Pb and

210

Po transfers to forest products, mainly mushrooms

TERRESTRIAL SYSTEMS

TABLE 6.2. ACTIVITY CONCENTRATIONS OF

210

Po AND

210

Pb IN MEADOW PLANTS OF DIFFERENT STEPPE

ECOSYSTEMS

Ecosystem

210

Po (Bq/kg, DW)

210

Pb (Bq/kg, DW)

210

Po:

210

Pb ratio

Above ground

biomass

Roots Ratio Above ground

biomass

Roots Ratio Above ground

biomass

Roots

Natural steppe 7.0 ± 4.4 42 ± 14 0.17 ± 0.1 17.5 ± 8.1 44 ± 15 0.41 ± 0.1 0.39 ± 0.1 0.96 ± 0.1

Saline meadow 3.7 ± 0.7 25 ± 24 0.16 ± 0.1 9.3 ± 2.6 31 ± 18 0.42 ± 0.3 0.4 ± 0.05 1.25 ± 0.2

Bog meadow 3.3 ± 2.0 23 ± 7 0.14 ± 0.1 11.8 ± 2.0 18 ± 4 0.62 ± 0.3 0.28 ± 0.05 1.25 ± 0.2

Sawn grasses 16.3 ± 6.7 —

—

49.6 ± 23.7 —

—

0.34 ± 0.08 —

Trees/bushes 13.7 ± 4.4 —

—

32.6 ± 13.3 —

—

Mean

8.1

± 5.9 30 ± 19 0.19 ± 0.1 21.5 ± 17 32 ± 17 0.47 ± 0.3 0.38 ± 0.12 0.97 ± 0.3

Source: See Ref. [6.28].

—: data not available.

CHAPTER 6

and berries, at different forest sites (Scots pine, Norway spruce, mixed forest and

Downy birch).

Vaaramaa et al.

[6.23] present data on

210

Pb and

210

Po in wild berries

(blueberry and lingonberry) and mushrooms in boreal forest ecosystems in

Finland. The lingonberry is an evergreen dwarf shrub and the blueberry is a

deciduous plant (see Table 6.4).

TABLE 6.3. CONCENTRATIONS OF

210

Po IN PLANTS OF

NORTHERN TAIGA

Group Species

210

(Bq/kg, DW)

Mosses

Conocephalum conicum

1480

Polytrichum commune

353

Mnium cinclidioides

293

Pleurozium schreberi

292

Sphagnum

274

Club-mosses

Lycopodium annotinum L.

100

Gymnocarpium dryopteris

Ferns

Dryopteris filix-mas L.

Grasses

Solidago virgaurea L.

Trientalis L.

Caltha palustris

Polygonum L.

Geum rivale L.

Filipendula ulmaria

Veratrum lobelianum

Anemone

Chamerion angustifolium

Equisetum arvense

Juncus L.

Sanguisorba officinalis

Source: See Ref. [6.29].

Note: Range is 13–1480 Bq/kg (DW); overall mean is 145 Bq/kg (DW);

standard deviation is 300 Bq/kg (DW); geometric mean is

61 Bq/kg (DW); median is 42 Bq/kg (DW); geometric standard

deviation is 1.1.

TERRESTRIAL SYSTEMS

TABLE 6.4. ACTIVITY CONCENTRATIONS OF

210

Po AND

210

Pb IN WILD BERRIES IN FINLAND

Wild berry Site

210

(Bq/kg, DW)

210

(Bq/kg, DW)

Blueberry N. Finland

S. Finland

0.7 ± 0.1

1.7 ± 0.3

3.2 ± 0.6

2.5 ± 0.4

Lingonberry N. Finland

S. Finland

1.4 ± 0.2

3.2 ± 0.3

2.2 ± 0.3

7.5 ± 1.0

Source: See Ref. [6.23].

The activity concentrations of

210

Pb and

210

Po in wild berries do not show

clear dependence on the type of forest and are quite similar to those reported by

Solatie et al.

[6.33] for Finnish Lapland and McDonald et al. [6.34] for samples

taken in England and Wales.

The highest

210

Pb (

210

Po) concentrations (dry weight) in blueberry shrubs

were found to be the following

[6.23]:

— Lignified brown stem of the blueberry shrub: 35 (80) Bq/kg;

— Roots: 33 (48) Bq/kg;

— Green stem 30 (29) Bq/kg;

— Leaves 16 (27) Bq/kg.

For lingonberry shrubs, the highest

210

Po concentrations (dry weight) were

found in the roots and rhizome (100 Bq/kg), followed by the current year stem

(90 Bq/kg), the lignified brown stem (30 Bq/kg), and the current year and old

leaves (22–26 Bq/kg) [6.23]. In general, the activity concentrations of

210

Pb and

210

Po in leaves and small tree branches were higher than those in above ground

biomass of grassy plants because of higher interception of the radionuclides from

the atmosphere, due to a higher leaf area index and a longer period of vegetation

during the year.

Although the

210

Po:

210

Pb activity ratios in the organic horizon of the

studied forests was close to 0.9, in most cases the concentrations of

210

Po in plant

tissues exceeded those of

210

Pb (

210

Po:

210

Pb varied from 1.5 to 2.5), except in

the lignified stem of lingonberry where the concentration of

210

Pb was 1.5 times

higher than

210

Po [6.23]. This might indicate that

210

Po is characterized by both

a higher transfer rate than

210

Pb from forest soil to plant compartments and by a

higher mobility within the plant. However, more precise research is needed to

confirm or reject these statements.

CHAPTER 6

According to the literature, fungi have been shown to accumulate stable

lead to much higher concentrations than berries

[6.35]. However, because of

their short lifespan, direct atmospheric deposition on the surfaces of the fruiting

bodies of mushrooms cannot be the major uptake route for elements

[6.36]. Thus,

environmental properties of the site where fungi grow can be more important for

radionuclide transfer

[6.23, 6.35, 6.36] provide information on

210

Pb and

210

transfer to many species of mushrooms.

Vaaramaa et

al. [6.23] find that the activity concentrations for

210

Pb and

210

Po in mushrooms in Finland are generally higher than wild berries, especially in

the case of

210

Po. The dry weight range for

210

Pb is 0.97–16.2 Bq/kg, and is much

wider for

210

Po, 13.9–1200 Bq/kg. They find the highest mean concentration of

210

Po in the fruiting bodies of the following families [6.23]:

— Boletaceae: 380 Bq/kg (DW);

— Russulaceae: 32.2 Bq/kg (DW);

— Cortinariaceae: 10 Bq/kg (DW).

They find the minimum mean values for

210

Po for mushrooms in the

following families [6.23]:

— Hygrophorus: 2.5/3.1 Bq/kg (DW) in the cap/stipe;

— Albatrellus: 3.8 Bq/kg (DW) in the whole fruit body;

— Ratio of the activity concentration between the caps and stipes: 1.6–7.1.

As for

210

Pb, considering the same species in different forest types,

Vaaramaa et al. [6.23] find the lowest activity concentrations in mushrooms

of the Boletaceae family (3.0 Bq/kg, DW) and the highest in Russulaceae

(11.1 Bq/kg, DW) and Cortinariaceae (11.1 Bq/kg, DW) [6.23]. Overall, these

data are similar to the

210

Pb concentrations (1.76–36.5 Bq/kg, DW) measured in

various mushrooms collected from the forests of France

[6.37].

In the northern Finland site, Vaaramaa et al. [6.23] report:

“No differences between forest types [Norway spruce, Scots pine, Downy

birch and mixed forest] were found in the concentrations of

210

Pb in the

whole fruiting bodies (cap and stipe combined), except in the activity

concentration of

210

Pb in Cortinarius armillatus L. collected from mixed

forest….

.......

TERRESTRIAL SYSTEMS

“The activity concentrations of

210

Pb were higher in the caps of the

mushrooms than in the stipes for Lactarius rufus

L. and Russula paludosa L.

The average cap-to-stipe ratio of

210

Pb in the Boletaceae family was 1.0,

which is in agreement with the mean ratio for

210

Pb (1.1) in Xerocomus

badius

L. (Boletaceae family) analyzed by Malinowska et al. [Ref. [6.38]].”

These values were statistically significantly higher than those for

210

Pb,

reflecting higher mobility of

210

Po among tissues of mushrooms. Vaaramaa et

al. [6.23] find that:

“The

210

Po/

210

Pb ratio in the mushrooms varied widely from 0.87 to 322 in

whole fruiting bodies, mainly being higher than one. The lowest value was

recorded in R.

paludosa L. and the highest in L. vulpinum L. In S. luteus L.

the ratio of

210

Po/

210

Pb has been reported to be 22 and 47 in Finnish Lapland

(Solatie et

al. [Ref. [6.33]]). In the present study, the ratio was 17 for the

same species in Scots pine forest of the northern Finland site.... The results

may indicate that

210

Po is more effectively taken up from soil to the fruiting

bodies than

210

Pb.”

6.5.4.

210

Pb and

210

Po transfer to agricultural plants

Several comprehensive reviews of radionuclide transfer from soil to plants,

including recommended TF values for both natural and artificial radionuclides

can be found in Refs

[6.4, 6.5] and in Table 6.5 for lead and polonium. The

TF values are for 14 plant groups according to the classifications presented in

Chapter

4 (see Ref. [6.5] for assignments of individual plants to these groups,

and the plant compartments considered). The TF values depend on the plants and

soil types assessed. Among the lowest values for both elements is maize grain

(5.2 × 10

−4

for lead and 1.8 × 10

−5

for polonium); while the highest are 1.0 for

polonium in pasture vegetation and 25 for lead in leafy vegetables.

However, when analysing the data in Table 6.5, it should be noted that

the radioactivity in the plant (or plant compartments) used for assessing the TF

(or CR) values is not only acquired through root transfer. The above ground

biomass might also be contaminated because of resuspension or direct deposition

of radionuclides from the atmosphere [6.8] (see also Section 6.3). These

atmospheric fluxes are of secondary importance for many radionuclides (such as

137

Cs and

Sr), but that might not be the case for

210

Pb and

210

Po, for which the

foliar pathway dominates in some contamination scenarios. In this context, plants

with edible compartments located above ground do not experience the same

contamination conditions as root crops and tubers, which can be contaminated

CHAPTER 6

TABLE 6.5. SOIL TO PLANT TRANSFER FACTORS FOR LEAD AND POLONIUM

Element Plant group Plant compartment Soil group

Mean/value GSD/SD

Min. Max.

Temperate environments

Pb Cereals Grain All 9 1.1 × 10

−2

3.6 1.9 × 10

−3

4.8 × 10

−2

Stems and shoots All 4 2.3 × 10

−2

3.5 5.1 × 10

−3

9.6 × 10

−2

Maize Grain All 9 1.2 × 10

−3

2.3 5.2 × 10

−4

3.8 × 10

−3

Stems and shoots All 3 2.8 × 10

−3

6.6 6.0 × 10

−4

2.3 × 10

−2

Leafy vegetables Leaves All 31 8.0 × 10

−2

1.3 × 10

3.2 × 10

−3

2.5 × 10

Sand 4 7.3 × 10

−2

1.5 4.9 × 10

−2

1.1 × 10

−1

Loam 3 8.2 × 10

−1

1.0 7.9 × 10

−1

8.6 × 10

−1

Clay 7 2.8 × 10

−2

4.1 4.1 × 10

−3

1.2 × 10

−1

Non-leafy vegetables Fruits, heads, berries,

buds

All 5 1.5 × 10

−2

2.6 × 10

1.5 × 10

−3

3.9

Stems and shoots All 2 8.8 × 10

−3

—

5.8 × 10

−3

1.2 × 10

−2

TERRESTRIAL SYSTEMS

TABLE 6.5. SOIL TO PLANT TRANSFER FACTORS FOR LEAD AND POLONIUM (cont.)

Element Plant group Plant compartment Soil group

Mean/value GSD/SD

Min. Max.

Pb Leguminous vegetables Seeds and pods All 17 5.3 × 10

−3

1.2 × 10

4.6 × 10

−4

4.9

Sand 3 2.7 × 10

−3

3.2 6.5 × 10

−4

8.9 × 10

−3

Loam 5 1.4 × 10

−3

4.4 6.5 × 10

−4

8.9 × 10

−3

Clay 4 8.0 × 10

−4

1.0 4.6 × 10

−4

1.0 × 10

−2

Stems and shoots All 1 8.0 × 10

−4

—

Root crops Roots All 27 1.5 × 10

−2

1.6 × 10

2.4 × 10

−4

3.3

Sand 5 6.4 × 10

−2

1.6 4.2 × 10

−2

1.2 × 10

−1

Loam 5 2.3 × 10

−3

4.7 2.4 × 10

−4

1.7 × 10

−2

Stems and shoots All 12 6.3 × 10

−2

150 3.0 × 10

−3

1.6 × 10

Tubers Tubers All 30 1.5 × 10

−3

7.4 1.5 × 10

−4

2.6

Sand 5 6.4 × 10

−3

3.5 1.6 × 10

−3

3.9 × 10

−2

Loam 17 5.2 × 10

−4

2.4 1.5 × 10

−4

2.3 × 10

−3

Grasses Stems and shoots All 17 3.1 × 10

−1

1.8 1.1 × 10

−1

1.0

CHAPTER 6

TABLE 6.5. SOIL TO PLANT TRANSFER FACTORS FOR LEAD AND POLONIUM (cont.)

Element Plant group Plant compartment Soil group

Mean/value GSD/SD

Min. Max.

Pb Pasture Stems and shoots All 34 9.2 × 10

−2

4.8 2.2 × 10

−3

1.0

Leguminous fodder Stems and shoots All 1 1.6 × 10

−2

n.a.

Po Cereals Grain All 2 2.4 × 10

−4

—

2.2 × 10

−4

2.6 × 10

−4

Maize Grain All 2 2.4 × 10

−4

—

1.8 × 10

−5

4.7 × 10

−4

Leafy vegetables Leaves All 12 7.4 × 10

−3

6.9 2.5 × 10

−4

5.0 × 10

−2

Non-leafy vegetables Stems and shoots All 2 1.9 × 10

−4

—

1.6 × 10

−5

3.7 × 10

−4

Leguminous vegetables Seeds and pods All 4 2.7 × 10

−4

3.9 6.0 × 10

−5

1.0 × 10

−3

Root crops Roots All 10 5.8 × 10

−3

4.3 2.4 × 10

−4

4.9 × 10

−2

Stems and shoots All 2 7.7 × 10

−2

—

5.8 × 10

−2

9.7 × 10

−2

Tubers Tubers All 9 2.7 × 10

−3

5.8 1.4 × 10

−4

3.4 × 10

−2

Pasture Stems and shoots All 10 1.2 × 10

−1

4.2 2.2 × 10

−2

1.0

Leguminous fodder Stems and shoots All 2 1.1 × 10

−2

—

2.6 × 10

−5

2.2 × 10

−4

TERRESTRIAL SYSTEMS

TABLE 6.5. SOIL TO PLANT TRANSFER FACTORS FOR LEAD AND POLONIUM (cont.)

Element Plant group Plant compartment Soil group

Mean/value GSD/SD

Min. Max.

Tropical environments

Cereals Grain Sand 1 2.5 × 10

−3

n.a.

Grasses Leaves Sand 9 2.1 × 10

−1

2.2 5.9 × 10

−2

1.00

Herbs Leaves Sand 2 3.7 × 10

−1

—

2.0 × 10

−2

7.1 × 10

−1

Leguminous

vegetables

Grain All 9 3.3 × 10

−3

2.4 6.5 × 10

−4

8.9 × 10

−3

Sand 3 3.4 × 10

−3

1.3 2.8 × 10

−3

4.4 × 10

−3

Loam 6 3.2 × 10

−3

3.0 6.5 × 10

−4

8.9 × 10

−3

Other crops Sand 18 2.3 × 10

−1

2.7 1.4 × 10

−2

1.0

Non-leafy vegetables All 2 7.0 × 10

−3

—

7.0 × 10

−3

7.0 × 10

−3

Pasture Stems and shoots Unspecified 1 3.0 × 10

−1

n.a.

Root crops Roots Loam 3 2.4 × 10

−3

1.5 1.8 × 10

−3

4.0 × 10

−3

CHAPTER 6

TABLE 6.5. SOIL TO PLANT TRANSFER FACTORS FOR LEAD AND POLONIUM (cont.)

Element Plant group Plant compartment Soil group

Mean/value GSD/SD

Min. Max.

Tubers Tubers All 16 5.7 × 10

−4

2.4 1.5 × 10

−4

2.3 × 10

−3

Sand 1 1.6 × 10

−3

n.a.

Loam 15 5.3 × 10

−4

2.4 1.5 × 10

−4

2.3 × 10

−3

Maize Grain All 6 8.5 × 10

−4

2.1 5.2 × 10

−4

3.8 × 10

−3

Sand 1 5.2 × 10

−4

n.a.

Loam 5 9.3 × 10

−4

2.2 5.9 × 10

−4

3.8 × 10

−3

Source: See Ref. [6.5].

Geometric standard deviation/standard deviation.

—: data not available.

n.a.: not applicable.

Data for polonium not available.

TERRESTRIAL SYSTEMS

through root uptake. Therefore, observed concentrations for leafy vegetables can

be up to 3–5 times higher due to deposition effects. It is also known that slowly

maturing plants have much lower transfer coefficients for

210

Po than rapidly

maturing plants under greenhouse conditions [6.39].

On the other hand, the concept of soil to plant TFs is adopted as a

reasonable empirical measure of plant contamination under steady state

conditions, when radionuclide fluxes corresponding to different pathways are

already well balanced. A general point is that the data in Table

6.5 also reflect

existing differences in mobility between radionuclides in soil and tend to be in

compliance with the data on K

values given in Table 6.1. In particular, TF values

for clay soils (where available) are normally about three times lower than those

for light sandy soils, and this ratio is close to the ratio for the K

values for the

same radionuclides and soil types.

6.6.

210

Pb AND

210

Po TRANSFER TO ANIMALS

In this section, transfer to animals of both lead and polonium are considered

because

210

Pb transferred to animal tissues decays to

210

Po. Therefore, its transfer

characteristics and rates to various animals are also relevant for polonium.

6.6.1. Processes governing distribution of

210

Pb and

210

Po in animals

Lead-210 is largely retained in bone, while

210

Po is distributed mainly

in soft tissue (see Table

6.6). Target tissues for

210

Po are the spleen, liver and

kidneys. However, the observed

210

Po:

210

Pb ratios in tissues are time dependent

and can be substantially modified by continuing radioactive decay of

210

resulting in support for

210

Po, as well as by the difference in biological half-lives

of these radionuclides in the different tissues. The rates of these processes greatly

depend on the animals and tissues of interest, and can vary across a wide range.

Following

210

Po administration to rats under laboratory conditions, the

highest concentrations are observed in the spleen, kidney, liver and other soft

tissue [6.40]. In wild animals (e.g. roe deer), the highest concentrations are

observed in kidney, liver and bone [6.30].

Some of the most detailed evidence of animal dependent distribution of

210

Po to

210

Pb is given by Maslov and Testov [6.41] in the 1970s, who report data

210

Pb and

210

Po activity in small mouse-like rodents (bank vole, six months

old) and reindeer (six years old) (see Table 6.7).

Differences in the accumulation of

210

Po, and the ratio with

210

Pb, in small

rodents and reindeer can be explained by the origin of the contamination. In the

bank vole, the

210

Po accumulated in some target organs (e.g. spleen, kidney and

100

CHAPTER 6

other soft tissue) originates from the diet and not, to any great extent, from

210

decay in the body. Therefore, the

210

Po:

210

Pb ratios in different tissues of these

small animals (with relatively short lifespans) typically exceed unity. In contrast,

the ratios in reindeer are normally less than one (0.02–0.3), since bone contains a

substantial amount of older

210

Pb accumulated in the animals as a consequence of

their longer lifespan. Data for other species are given in Table

6.8.

The highest accumulation of

210

Po was observed in small rodents and

earthworms, followed by reptiles and amphibians. For

210

Pb, the highest activity

concentrations (around 50 Bq/kg, DW) were found in earthworms. The mobility

210

Po varies in different species:

210

Po:

210

Pb ratios were highest for rodents,

TABLE 6.6. CONCENTRATIONS OF

210

Pb AND

210

Po:

210

Pb RATIOS IN

BONE AND MUSCLE OF ANIMALS IN FINLAND

Animal

Bone Muscle

210

(Bq/kg, DW)

210

Po:

210

(Bq/kg, DW)

210

Po:

210

Deer 1338 0.56 0.17 24

Moose 18.5 0.54 0.06 12.5

Wolf 9.6 0.69 0.02 100

Wolverine 22.9 0.67 0.08 90

TABLE 6.7. ACTIVITY OF

210

Po IN ANIMALS WITH DIFFERENT

LIFESPANS

Tissue

Bank vole Reindeer

210

(mBq)

210

Po ratio

to bone

210

(mBq)

210

Po ratio

to bone

Bone 1–7.8 1 765.9 1

Muscle 27 1.5 23.7 0.02

Liver 23 1.3 261.2 0.3

Spleen 177.6 10 45.9 0.05

Source: See Ref. [6.41].

101

TERRESTRIAL SYSTEMS

while for earthworms they were close to unity. Ratios greater than 100 have also

been reported in the muscle of wild boar [6.42].

6.6.2. Quantification of

210

Pb and

210

Po transfers to agricultural animals

6.6.2.1. Absorption

The International Commission on Radiological Protection (ICRP)

recommends a relatively high fractional gastrointestinal absorption value of 0.5

for polonium [6.43]. Considering ruminants in particular, data specifically for

polonium gastrointestinal absorption have not been identified; however, a lower

mean fractional gastrointestinal absorption value of 0.04 for lead is reported by

TABLE 6.8. WHOLE BODY (EXCLUDING GASTROINTESTINAL TRACT)

ACTIVITY CONCENTRATIONS AND ESTIMATED

210

Po:

210

Pb RATIOS IN

NORDIC WILD SPECIES

Species N

210

Po (Bq/kg, DW)

210

Pb (Bq/kg, DW)

210

Po:

210

Mean SD Mean SD

Common shrew 9 37.2 19 0.3 0.4 258

Bank vole 8 65.1 17.3 1.5 1.1 77

Willow grouse 5 3.3 1.2 2.1 1.6 2.2

Red earthworm 2 48.6 28.7 50 8 1.0

Grey worm 5 57.2 42.3 47 23 1.5

Viper 1 20.9

21.84

—

13.05

12.09

—

Frog 1 5.65

5.88

—

3.39

2.98

—

Source: Adapted from Ref. [6.24].

Note: SD — standard deviation. The intestinal tracts of the worms were not removed.

The

210

Po:

210

Pb ratios are averages based on individual sample values rather than species

means.

—: data not available.

102

CHAPTER 6

Howard et al. [6.44]. The value is within the range of 0.01–0.1 recommended for

polonium in humans by the ICRP [6.43].

6.6.2.2. Transfer to animal food products

The major pathway for polonium transfer to animals is through ingestion

of feed and water. Parameters used to estimate that transfer are the transfer

coefficient — F

and F

for milk and flesh, respectively — and the CR (for a

definition, see Chapter

2). Given its easy absorption in the gut, the transfer to

meat and milk is relatively high for some animal products. As lead accumulates

in bone, its transfer to milk and meat (although higher in some offal) might be

expected to be lower than polonium. However, data to make the comparison are

presently inadequate to test this assumption for most products (see Table 6.9).

The transfer coefficient for polonium to cow milk is based on four data

sources and is lower than that for goat milk [6.45–6.48]. The difference does not

mean that higher activity concentrations are expected in goat milk, as transfer

coefficients include the amount of feed ingested [6.4, 6.5, 6.49]. Since CRs for

milk for all ruminants are similar, it is likely that the polonium CR for cow milk

would be a reasonable value to use for polonium in goat and sheep milk [6.49].

The polonium CR for beef is based on only one data source

[6.50] and there is

some uncertainty concerning its derivation (lack of information on weight basis),

so it has not been included in Table

6.9.

Using data from the United Nations Scientific Committee on the Effects

of Atomic Radiation (UNSCEAR), the reference concentration (FW) for

210

in milk and meat is 15

mBq/kg and 60 mBq/kg, respectively [6.51]. When this

is compared with a mean value (33.5 Bq/kg, DW) of natural steppe and sawn

grass (see Table

6.2) and all grass species (see Table 6.3), the respective CRs are

4.5

× 10

−4

and 1.8 × 19

−3

. These CRs are based on specific types of ecosystem for

vegetation and unassociated milk and are therefore uncertain. However, the value

for cow milk is similar to Ref.

[6.5]. A direct comparison of transfer of lead and

polonium is only possible for cow milk. In this case, both F

and CRs are similar

for the two elements. The CR for beef would be expected to be lower than for

reindeer meat, so the value given based on UNSCEAR data

[6.51] might be more

appropriate than the value in Ref. [6.5].

The transfer coefficient for polonium is relatively high for poultry meat and

egg contents. However, this might be the result of the values being derived from

unrelated feed and food samples

[6.52].

The equivalent transfer values for lead are given in Table

6.10. In

addition, similar values for cow milk of F

and CR of 3.6 × 10

−4

and 7.3 × 10

−3

respectively, are reported in Ref.

[6.53], which also reports higher values of

and CR for beef (n = 3) for lead of 0.01 and 0.11, respectively. The values

103

TERRESTRIAL SYSTEMS

incorporate review information from animal nutrition studies as well as other

literature on metal transfer.

6.6.3. Transfer to terrestrial animals

The transfer of polonium to products such as game species is not quantified

in Refs

[6.4, 6.5], and therefore, data specifically on the human food chain are

not available. However, the transfer of radionuclides to wildlife has recently

been published in Refs

[6.54, 6.55], in which the transfer to terrestrial animals is

reported as CRs between the fresh weight whole body (excluding shell, gut and

contents, feathers and pelt) and dry weight soil for a few species (see Table

6.11).

The CR

wo-soil

for polonium transfer to terrestrial animals is high compared

to many other radionuclides. These values are not directly comparable to those

given above for beef, as the CRs are for the whole organism (including bone),

and the denominator is soil, not ingested feed. The especially high transfer of

polonium to reindeer is well known and has already been discussed

[6.56]. The

wo-soil

for different mammal categories varies widely, with a much lower value

for herbivores than those carnivores or omnivores. The high CR

wo-soil

for reptiles

is influenced by relatively high values from Australia

[6.57] and may be due to

acidic spray from mine tailings

[6.58].

TABLE 6.9. TRANSFER COEFFICIENTS AND CONCENTRATION RATIOS

FOR POLONIUM TRANSFER TO ANIMALS AND FOODSTUFFS

Animal product

Transfer coefficient (d/L or d/kg) Concentration ratio

Mean GSD Min. Max.

Cow milk 4 2.1 × 10

−4

1.8 8.9 × 10

−5

3. × 10

−4

1 2.4 × 10

−3

Goat milk 2 2.2 × 1

−3

n.a.

1.8 × 10

−3

2.7 × 10

−3

—

Poultry 1 2.4 n.a.

n.a.

—

Reindeer meat

—

5 × 10

−2

—

5 × 10

−2

Egg 1 3.1 n.a.

n.a.

—

Sources: See Refs [6.5, 6.31].

Note: CR — concentration ratio; GSD — geometric standard deviation.

n.a.: not applicable.

—: data not available.

Assuming lichen intake [6.31].

104

CHAPTER 6

TABLE 6.11. TRANSFER OF POLONIUM TO THE WHOLE BODY OF

TERRESTIAL ANIMALS

Wildlife group

wo-soil

AM ASD GM GSD Min. Max.

Annelid 7 1.0 × 10

−1

3.9 × 10

−2

9.6 × 10

−2

1.4 —

—

Bird

Herbivorous

1.0 × 10

−2

2.9 × 10

−3

9.6 × 10

−3

1.3

—

Mammal

Carnivorous

Herbivorous

Omnivorous

Rangifer spp.

199

8.6 × 10

−2

1.2 × 10

−1

2.9 × 10

−3

2.1 × 10

−1

2.5

2.1 × 10

−1

8.7 × 10

−2

1.9 × 10

−3

1.2 × 10

−1

3.7

3.3 × 10

−2

9.7 × 10

−2

2.4 × 10

−3

1.8 × 10

−1

1.4

4.0

1.9

1.8

1.7

3.0

2.4 × 10

−4

1.9 × 10

−2

2.4 × 10

−4

7.5 × 10

−4

5.9 × 10

−1

1.1

1.4 × 10

−1

9.5 × 10

−3

2.6 × 10

−1

Reptile 15 9.5 23 3.6 4.0 1.9 × 10

–2

Source: See Refs [6.54, 6.55].

Note: CR

wo-soil

is the activity concentration in the whole organism (in Bq/kg, FW) divided by

the activity concentration in the soil (in Bq/kg, DW). Rangifer spp. include reindeer

and caribou. AM — arithmetic mean; ASD — arithmetic standard deviation; CR —

concentration ratio; GM — geometric mean; GSD — geometric standard deviation.

—: data not available.

TABLE 6.10. TRANSFER COEFFICIENTS FOR LEAD TRANSFER TO

ANIMALS AND FOODSTUFFS

Animal product

Transfer coefficient (d/L and d/kg) Concentration ratio

Mean GSD Min. Max.

Cow milk 4 1.9 × 10

−4

1.0 7.3 × 10

−6

1.2 × 10

−3

1 2.4 × 10

−3

Goat milk 1 6.0 × 10

−3

n.a.

1 9.0 × 10

−3

Sheep milk 1 3.5 × 10

−2

n.a.

1 3.0 × 10

−2

Beef 5 7.0 × 10

−4

2.5 2.0 × 10

−4

1.6 × 10

−3

11 7.7 × 10

−2

Mutton 2 7.1 × 10

−3

—

4.0 × 10

−3

1.0 × 10

−2

3 1.2 × 10

−2

Pork —

—

2 6.6 × 10

−1

Source: See Ref. [6.5].

Note: CR — concentration ratio; GSD — geometric standard deviation.

n.a.: not applicable.

—: data not available.

105

TERRESTRIAL SYSTEMS

The CR

wo-soil

for lead transfer to terrestrial animals (see Table 6.12) are

generally lower than those for polonium. As for polonium, there is relatively

high transfer to Rangifer species. The mammal values are similar for all of

the subcategories. The mammal values in Tables

6.11 and 6.12 are not directly

comparable to those given for beef in Tables

6.9 and 6.10 as the transfer

coefficient includes dietary intake and the CR compares muscle (not whole body)

with ingested feed and not soil.

TABLE 6.12. TRANSFER OF LEAD TO THE WHOLE BODY OF DIFFERENT

WILDLIFE GROUPS

Wildlife group

wo-soil

AM ASD GM GSD Min. Max.

Amphibian 24 1.2 × 10

−1

5.2 × 10

−1

2.7 × 10

−2

5.6 8.8 × 10

−4

2.8 × 10

−1

Annelid 647 5.2 × 10

−1

7.5 × 10

−1

2.9 × 10

−1

2.9 2.3 × 10

−3

2.8

Arachnid 2 5.3 × 10

−2

—

4.3 × 10

−2

6.2 × 10

−2

Arthropod 561 4.0 × 10

−1

4.7 × 10

−1

2.6 × 10

−1

2.5 4.6 × 10

−3

1.0

Bird

Carnivorous

424

6.2 × 10

−2

1.7 × 10

−1

2.1 × 10

−2

4.4

—

Mammal

Carnivorous

Herbivorous

Omnivorous

Rangifer spp.

515

368

270

3.8 × 10

−2

4.7 × 10

−2

2.0 × 10

−2

1.2 × 10

−2

3.6

3.6 × 10

−2

2.8 × 10

−2

2.7 × 10

−2

6.3 × 10

−2

3.3

2.8 × 10

−2

4.0 × 10

−2

1.2 × 10

−2

2.2 × 10

−3

2.7

2.2

1.7

2.8

6.3

2.2

2.7 × 10

−4

8.8 × 10

−3

1.9 × 10

−3

2.7 × 10

−4

4.0 × 10

−1

2.0 × 10

−1

7.7 × 10

−2

2.0 × 10

−1

3.9 × 10

−2

Gastropod 47 7.3 × 10

−3

1.3 × 10

−2

3.6 × 10

−3

3.3 6.1 × 10

−4

3.8 × 10

−2

Reptile

Carnivorous

3.7 × 10

−1

3.8 × 10

−2

1.0

1.6 × 10

−1

1.3 × 10

−1

8.7 × 10

−3

4.3

5.6

1.4 × 10

−3

1.4 × 10

−3

1.2

7.0 × 10

−2

Source: See Refs [6.54, 6.55].

Note: CR

wo-soil

is the activity concentration in the whole organism (in Bq/kg, FW) divided by

the activity concentration in the soil (in Bq/kg, DW). Rangifer spp. include reindeer

and caribou. AM — arithmetic mean; ASD — arithmetic standard deviation; CR —

concentration ratio; GM — geometric mean; GSD — geometric standard deviation.

—: data not available.

106

CHAPTER 6

Generally, the behaviour of

210

Po in terrestrial ecosystems depends on

many processes which have been studied little. The available data are largely

based on field observations that attempt to evaluate the role of individual

factors governing the behaviour of

210

Po. However, it is not always possible to

identify these processes in uncontrolled conditions. Furthermore, the quantity

210

Po in environmental compartments depends to some extent on the ambient

concentrations of

210

Pb and the behaviour of

210

Pb and

210

Po. Therefore, both

aspects should be considered when trying to explain any observed patterns

210

Po accumulation in specific environmental receptors. The generic data

presented in this chapter can be used for assessing the impacts of polonium on

the environment. However, these values are often based on only a few measured

values or single observations, which are sometimes contradictory. Therefore, the

data presented here should be considered with some caution, bearing in mind the

relatively high uncertainty associated with many of the values.

REFERENCES TO CHAPTER 6

[6.1] INTERNATIONAL UNION OF RADIOECOLOGY, Recommendations for Improving

Predictions of the Long-term Behaviour of

Cl,

Tc,

237

Np and

238

U, IUR

Report

6, IUR, Warrington (2006).

[6.2] INTERNATIONAL ATOMIC ENERGY AGENCY, ‘Reference Biospheres’ for Solid

Radioactive Waste Disposal, Report of BIOMASS Theme 1 of the BIOsphere

Modelling and ASSessment (BIOMASS) Programme, Part of the IAEA Co-ordinated

Research Project on Biosphere Modelling and Assessment (BIOMASS),

IAEA-BIOMASS-6, IAEA, Vienna (2003).

[6.3] IVANOVICH, M., HARMON, R.S. (Eds), Uranium Series Disequilibrium:

Applications to Environmental Problems, Clarendon Press, Oxford

(1982).

[6.4] INTERNATIONAL ATOMIC ENERGY AGENCY, Quantification of Radionuclide

Transfer in Terrestrial and Freshwater Environments for Radiological Assessments,

IAEA-TECDOC-1616, IAEA, Vienna

(2009).

[6.5] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer in Terrestrial and Freshwater Environments,

Technical Reports Series No.

472, IAEA, Vienna (2010).

[6.6] MAHON, D.C., MATHEWES, R.W., Uptake of naturally occurring radioisotopes by

vegetation in a region of high radioactivity, Can. J. Soil Sci.

63 (1983) 281–290.

[6.7] HAM, G.J., EWERS, L.W., WILKINS, B.T., Variations in Concentrations of Naturally

Occurring Radionuclides in Foodstuffs, NRPB-M892, National Radiological

Protection Board, Chilton

(1998).

[6.8] PIETRZAK-FLIS, Z., SKOWROŃSKA-SMOLAK, M., Transfer of

210

Pb and

210

Po to

plants via root system and above-ground interception, Sci. Total Environ.

162

(1995)

139–147.

107

TERRESTRIAL SYSTEMS

[6.9] LECLERC, E., TAGAMI, K., UCHIDA, S., VARGA, B., “Use of analogues”,

Quantification of Radionuclide Transfer in Terrestrial and Freshwater Environments

for Radiological Assessments, IAEA-TECDOC-1616, IAEA, Vienna (2009)

605–614.

[6.10] PRÖHL, G., “Interception”, ibid., pp 29–43.

[6.11] HOFFMAN, F.O., THIESSEN, K.M., RAEL, R.M., Comparison of interception and

initial retention of wet-deposited contaminants on leaves of different vegetation types,

Atmos. Environ.

29 (1995) 1771–1775.

[6.12] LECLERC, E., CHOI, Y.H., “Weathering”, Quantification of Radionuclide Transfer in

Terrestrial and Freshwater Environments for Radiological Assessments,

IAEA-TECDOC-1616, IAEA, Vienna (2009)

45–48.

[6.13] LECLERC, E., COLLE, C., MADOZ-ESCANDE, C., CHOI, Y.H., “Translocation”,

ibid., pp

49–62.

[6.14] BAEZA, A., PANIAGUA, J.M., RUFO, M., STERLING, A., BARANDICA, J.,

Radiocaesium and radiostrontium uptake by turnips and broad beans via leaf and root

absorption, Appl. Radiat. Isot.

50 (1999) 467–474.

[6.15] BRAMBILLA, M., FORTUNATI, P., CARINI, F., Foliar and root uptake of

134

Cs,

and

Zn in processing tomato plants (Lycopersicon esculentum Mill.), J. Environ.

Radioact.

60 (2002) 351–363.

[6.16] SHEPPARD, S.C., SHEPPARD, M.I., ILIN, M., TAIT, J., SANIPELLI, B., Primordial

radionuclides in Canadian background sites: Secular equilibrium and isotopic

differences, J. Environ. Radioact.

99 (2008) 933–946.

[6.17] COPPIN, F., ROUSSEL-DEBET, S., Comportement du

210

Po en milieu terrestre: revue

bibliographique, Radioprot.

39 (2004) 39–58.

[6.18] JOURDAIN, F., “Resuspension”, Quantification of Radionuclide Transfer in Terrestrial

and Freshwater Environments for Radiological Assessments, IAEA-TECDOC-1616,

IAEA, Vienna (2009) 63–68.

[6.19] ALFARO, S.C., GAUDICHET, A., GOMES, L., MAILLÉ, M., Modelling the size

distribution of a soil aerosol produced by sandblasting, J. Geophys. Res.

102 (1997)

239–11 249.

[6.20] KIDO, H., FUJIWARA, H., JAMSRAN, U., ENDO, A., The simulation of long-range

transport of

137

Cs from East Asia to Japan in 2002 and 2006, J. Environ. Radioact. 103

(2012)

7–14.

[6.21] BALKANSKI, Y.J., JACOB, D.J., GARDNER, G.M., GRAUSTEIN, W.C.,

TUREKIAN, K.K., Transport and residence times of tropospheric aerosols inferred

from a global three-dimensional simulation of

210

Pb, J. Geophys. Res. 98 (1993)

573–20 586.

[6.22] BASKARAN, M., Po-210 and Pb-210 as atmospheric tracers and global atmospheric

Pb-210 fallout: A review, J. Environ. Radioact.

102 (2011) 500–513.

[6.23] VAARAMAA, K., ARO, L., SOLATIE, D., LEHTO, J., Distribution of

210

Pb and

210

concentrations in wild berries and mushrooms in boreal forest ecosystems, Sci. Total

Environ.

408 (2009) 84–91.

[6.24] BROWN, J.E., et al., Levels and transfer of

210

Po and

210

Pb in Nordic terrestrial

ecosystems, J. Environ. Radioact.

102 (2011) 430–437.

108

CHAPTER 6

[6.25] VANDENHOVE, H., GIL-GARCIA, C., RIGOL, A., VIDAL, M., New best estimates

for radionuclide solid–liquid distribution coefficients in soils — Part

2: Naturally

occurring radionuclides, J. Environ. Radioact.

100 (2009) 697–703.

[6.26] PETTERSSON, H.B.L., HALLSTADIUS, L., REDVALL, R., HOLM, E.,

Radioecology in the vicinity of prospected uranium mining sites in the subarctic

environment, J. Environ. Radioact.

6 (1988) 25–40.

[6.27] SANZHAROVA, N., FESENKO, S., REED, E., “Processses governing radionuclide

transfer to plants”, Quantification of Radionuclide Transfer in Terrestrial and

Freshwater Environments for Radiological Assessments, IAEA-TECDOC-1616,

IAEA, Vienna (2009)

123–138.

[6.28] LADINSKAYA, L.A., PARFENOV, Yu.D., ARUTYNOV, P.S., POPOVA, D.K.,

Natural concentrations of Pb-210 and Po-210 in plants of steppe ecosystems,

Ehkologiya 1 (1976) 32–35 (in Russian).

[6.29] TASKAYEV, A.I., TESTOV, B.V., “Major components of doses to biological

organisms from incorporated primordial radionuclides”, Heavy Natural Radionuclides

in Biosphere (ALEXAKHIN, R.M., Ed.), Nauka, Moscow (1990) 201–217

(in Russian).

[6.30] ERMOLAYEVA-MAKOVSKAYA, A.P., LITVER, B.Y., Lead-210 and polonium-210

in biosphere, Atomizdat, Moscow (1978) (in

Russian).

[6.31] SKUTERUD, L., GWYNN, J.P., GAARE, E., STEINNES, E., HOVE, K.,

Sr,

210

and

210

Pb in lichen and reindeer in Norway, J. Environ. Radioact. 84 (2005) 441–456.

[6.32] THOMAS, P.A., SHEARD, J.W., SWANSON, S., Transfer of

210

Po and

210

Pb through

the lichen–caribou–wolf food chain of northern Canada, Health Phys.

(1994)

666–677.

[6.33] SOLATIE, D., JUNTTILA, M., VESTERBACKA, P.,

210

Po and

210

Pb in the food chain

lichen–reindeer–man, Radiochemistry (Radiokhimiya)

48 (2006) 632–633.

[6.34] McDONALD, P., JACKSON, D., LEONARD, D.R.P., McKAY, K., An assessment of

210

Pb and

210

Po in terrestrial foodstuffs from regions of England and Wales, J. Environ.

Radioact.

43 (1999) 15–29.

[6.35] PELKONEN, R., ALFTHAN, G., JÄRVINEN, O., Cadmium, Lead, Arsenic and

Nickel in Wild Edible Mushrooms, The Finnish Environment

17, Finnish Environment

Institute, Helsinki

(2006).

[6.36] SVOBODA, L., ZIMMERMANNOVÁ, K., KALAČ, P., Concentrations of mercury,

cadmium, lead and copper in fruiting bodies of edible mushrooms in an emission area

of a copper smelter and a mercury smelter, Sci. Total Environ.

246 (2000) 61–67.

[6.37] KIRCHNER, G., DAILLANT, O., Accumulation of

210

Pb,

226

Ra and radioactive

cesium by fungi, Sci. Total Environ.

222 (1998) 63–70.

[6.38] MALINOWSKA, E., SZEFER, P., BOJANOWSKI, R., Radionuclides content in

Xerocomus badius and other commercial mushrooms from several regions of Poland,

Food Chem.

97 (2006) 19–24.

[6.39] WATTERS, R.L., HANSEN, W.R., The hazards implication of the transfer of

unsupported

210

Po from alkaline soil to plants, Health Phys. 18 (1970) 409–413.

[6.40] PARFENOV, Yu.D., Polonium-210 in the environment and in the human organism, At.

Energy Rev.

12 (1974) 75–143.

109

TERRESTRIAL SYSTEMS

[6.41] MASLOV, V.B., TESTOV, B.V., Activity Concentrations and Distribution of

210

Po and

210

Pb in Animals, Inhibiting Areas with Elevated Radiation Background, Results of

Radioecological Research in Natural Environment, Syktuvkar, Komi Branch, Academy

of Science of the Former USSR (1971)

98–106.

[6.42] MARTIN, P., HANCOCK, G.J., JOHNSTON, A., MURRAY, A.S., Natural-series

radionuclides in traditional North Australian Aboriginal foods, J. Environ. Radioact.

(1998)

37–58.

[6.43] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION, Human

Alimentary Tract Model for Radiological Protection, Publication 100, Elsevier,

Amsterdam (2007).

[6.44] HOWARD, B.J., BERESFORD, N.A., BARNETT, C.L., FESENKO, S.,

Gastrointestinal fractional absorption of radionuclides in adult domestic ruminants, J.

Environ. Radioact.

100 (2009) 1069–1078.

[6.45] WATTERS, R.L., McINROY, J.F., The transfer of

210

Po from cattle feed to milk,

Health Phys.

16 (1969) 221–225.

[6.46] JOHNSON, J.E., WATTERS, R.L.,

210

PoO

Metabolism in Ruminants, C00-2044-5,

Colorado State Univ., Colorado

(1972).

[6.47] MOROZ, B.B., PARFENOV, Yu.D., Metabolism and biological effects of

polonium-210, At. Energy Rev.

10 (1972) 175–232.

[6.48] SCHÜTTELKOPF, H., KIEFER, H., “The Ra-226, Pb-210 and Po-210 contamination

of the Black Forest in natural radiation environment”, Natural Radiation Environment

(Proc. 2nd Special Symp. Bombay, 1981) (VOHRA, K.G., MISHRA, V., PILLAI, K.,

SADAÏWAN, S., Eds), John Wiley and Sons, New

York (1982).

[6.49] HOWARD, B.J., BERESFORD, N.A., BARNETT, C.L., FESENKO, S., Quantifying

the transfer of radionuclides to food products from domestic farm animals, J. Environ.

Radioact.

100 (2009) 767–773.

[6.50] MILOSEVIC, Z., HORSIC, E., KLJAJIC, R., BAUMAN, A., “Distribution of

uranium,

226

Ra,

210

Pb and

210

Po in the ecological cycle in mountain regions of central

Yugoslavia”, IRPA

5, Jerusalem, International Radiation Protection Association,

Fontenay-aux-Roses (1980)

1123–1126.

[6.51] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Sources and Effects of Ionizing Radiation UNSCEAR 2000 Report to

the General Assembly, with Scientific Annexes, Vol. I: Sources, United Nations,

New York (2000) Annex B.

[6.52] IZAK-BIRAN, T., et al., Concentrations of U and Po in animal feed supplements, in

poultry meat and in eggs, Health Phys.

56 (1989) 315–319.

[6.53] SHEPPARD, S.C., LONG, J.M., SANIPELLI, B., Verification of radionuclide transfer

factors to domestic-animal food products, using indigenous elements and with

emphasis on iodine, J. Environ. Radioact.

101 (2010) 895–901.

[6.54] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer to Wildlife, Technical Reports Series

No.

479, IAEA, Vienna (2014).

[6.55] HOWARD, B.J., et al., The IAEA handbook on radionuclide transfer to wildlife, J.

Environ. Radioact.

121 (2013) 55–74.

110

CHAPTER 6

[6.56] PERSSON, B.R.R., HOLM, E., Polonium-210 and lead-210 in the terrestrial

environment: A historical review, J. Environ. Radioact.

102 (2011) 420–429.

[6.57] READ, J., PICKERING, R., Ecological and toxicological effects of exposure to an

acidic, radioactive tailings storage, Environ. Monit. Assess.

54 (1999) 69–85.

[6.58] WOOD, M.D., BERESFORD, N.A., SEMENOV, D.V., YANKOVICH, T.L.,

COPPLESTONE, D., Radionuclide transfer to reptiles, Radiat. Environ. Biophys.

(2010)

509–530.

111

Chapter 7

210

Pb AND

210

Po IN FRESHWATER AND

GROUNDWATER SYSTEMS

7.1. DESCRIPTION OF THE HYDROLOGICAL CYCLE

The continuous movement of water on, above and below the surface of the

Earth is called the hydrological cycle. Throughout this cycle, water changes from

one phase to another: from solid to liquid to vapour (see Fig.

7.1).

Precipitation runs off the land surface into streams and rivers to fill

standing bodies of water, such as natural lakes and artificial impoundments, on

its way to the sea and comes into contact with river and lake-bed sediments as it

moves. Some evaporates, but a large proportion of precipitation also infiltrates

the ground, where it interacts with soils and rocks. The infiltrating waters then

recharge the underlying groundwater systems (see Section

7.2). Groundwater

discharges into surface waters, such as lakes, rivers and wetlands, as well as into

coastal waters. All these waters and discharges can carry polonium.

The surface water environment includes streams, rivers, lakes, bogs,

fens and wetlands, along with the sediment in each of these water bodies (see

Section 7.3). Polonium in surface water can be directly transferred to other

surface water bodies, or can deposit and migrate in bottom sediments. It can also

FIG. 7.1. Illustration of the hydrological cycle.

112

CHAPTER 7

be taken up by freshwater organisms, thus entering the food chain. The transfer in

surface water occurs through a combination of processes, including

[7.1]:

(a) Diffusion and dispersion caused by gradients of concentration and the

turbulent motion of water;

(b) Transport caused by water currents;

(suspended solids);

(d) Deposition and remobilization following interaction with suspended matter

and bottom sediments.

Hydraulic processes responsible for radionuclide transfer by water are

mainly independent of radionuclide properties. Exchanges of radionuclides

between the dissolved and solid phases, as well as interaction with abiotic

suspended matter and bottom sediments, are strongly radionuclide specific.

Quantification of radionuclide redistribution in freshwater ecosystems requires

the evaluation of the physicochemical properties of radionuclides in water

and sediments. A description of how these processes can be represented in a

quantitative manner is provided in Ref. [7.2].

Polonium in water can then transfer to biota, such as plants, phytoplankton,

zooplankton, other invertebrates (including molluscs and crustaceans), fish and

water based amphibians, as well as to reptiles, mammals and birds that obtain

dietary components from the aquatic environment.

7.2. GROU NDWATER

7.2.1. Groundwater environment

Groundwater is an important component of the hydrological cycle and

source of drinking water for humans. A simplified typical aquifer system is shown

in Fig. 7.2. While water recharging the groundwater contains some

210

Po from

rainwater, any surface deposition or releases and interactions in the vadose zone,

the

210

Po concentrations further in the aquifer are controlled by local processes,

as steady state conditions are established over very short distances and tend to

break down over longer distances. In recharge zones, as precipitation enters the

subsurface, there are a number of sources of polonium in the water:

— Polonium contained in the rainwater or deposited on soils from the

atmosphere (i.e. wet or dry deposition; see Chapter

5);

— Polonium already in soils (see Chapter 6);

113

FRESHWATER AND GROUNDWATER SYSTEMS

— Anthropogenic releases (see Chapter 4);

— Polonium produced by progenitor nuclides in the

238

U decay series

(see Fig. 3.1, in Chapter 3) that have similar origins.

However, owing to its short half-life,

210

Po from sources confined to the near

surface cannot migrate far before decaying away. As the water begins to interact

with the surrounding minerals, both within the vadose zone and the underlying

aquifers, there are a number of factors that control the distribution of polonium

in the groundwater, including: the decay of the parent

210

Bi (which is the very

short lived daughter of

210

Pb) that is dissolved or adsorbed; the decay of

210

Po,

supply by recoil from mineral grains; and adsorption. The short half-life means

that the concentration is determined by the conditions in the immediate area.

Groundwater discharges into surface waters (see Section 7.3) and coastal waters

and are thus important inputs to these environments. These inputs are likely to

be moderated by interactions within sediments at the sediment–water interface,

although there is insufficient data to document the general characteristics of the

relationships between discharging groundwater and surface waters.

The behaviour of polonium in groundwater has been studied much less than

on other radionuclides — presumably because it is relatively immobile and short

lived. However, it has sometimes been found to be a significant contributor to the

overall gross alpha activity of groundwater, which has in turn led to a number of

regional studies.

7.2.2. Sources of

210

Pb and

210

Po in groundwater

An important input of

210

Po is the decay of parent isotopes dissolved within

the groundwater or adsorbed onto aquifer surfaces. Therefore, the distribution of

FIG. 7.2. A simplified view of groundwater flow and the dominant processes controlling the

distribution of

210

Po within aquifers.

114

CHAPTER 7

progenitor atoms is important, with the distribution of uranium controlling the

distribution and release of

230

Th, which, in turn, controls that of

226

Ra, which

affects the release of

222

Rn by recoil, and so the concentrations of the short lived

nuclides further along the chain (see Section

7.2.5).

A dominant input to groundwater for many constituents is weathering of

aquifer rocks and thus chemical releases from mineral structures

[7.3]. This is

important for uranium, thorium and radium (all progenitors of

210

Po), and is

responsible for much of the redistribution of the long lived nuclides. For short

lived nuclides, however, this is generally not important compared to other

sources, and so can be neglected

[7.4].

The most important mechanism for the release of short lived nuclides from

within mineral structures is recoil (see Figs

7.2 and 7.3). Whenever an atom

is subject to energetic alpha decay, the daughter atom is recoiled in a random

direction (though directly opposite to that of the alpha particle) over a distance

of approximately 20 nm in most mineral lattices [7.3]. When this recoil path

crosses a grain boundary (mineral surface), the daughter atom can be released to

a solution in the groundwater

[7.5, 7.6]. Atoms implanted into adjacent phases

may escape as well, along the alpha track defect in the mineral lattice generated

during the implantation event

[7.3].

The stopping power of water is substantially lower than mineral lattices,

and so atoms can cross pores and become implanted in adjacent grains, from

which the ion may migrate back to the solution [7.3]. This distance is somewhat

dependent on the lattice characteristics as well as on the decay energy (see the

review in Ref. [7.4]), although the recoil characteristics of the alpha decay

Source: Figure 2 of Ref. [7.3].

FIG. 7.3. A major source of short lived nuclides in groundwater is recoil (reproduced with

permission courtesy of Elsevier).

115

FRESHWATER AND GROUNDWATER SYSTEMS

products along the decay series are similar. Quantifying this input is discussed

further in Section 7.2.5.

Owing to the short half-life of

210

Po, its levels in groundwater are generally

in a steady state, such that the removal rates are equal to the input rates. While

much of the polonium is adsorbed, this may be readily exchangeable with

dissolved species (see Section 7.2.5). The main removal process of adsorbed

and dissolved

210

Pb is simply decay. Irreversible incorporation of

210

Po into

precipitating phases is generally too slow to be a significant removal flux.

7.2.3. Groundwater concentrations

Porcelli [7.3] reports that reconnaissance studies of groundwater across the

United States of America [7.7] and across California [7.8] find groundwater

210

concentrations to be typically less than 40 mBq/L. Values as high as 16 Bq/L

have been recorded in a well in Finland [7.9], and brines in Australia can have

values of up to 19 Bq/L [7.10]. A number of factors control the variations in

concentrations of

210

Po in groundwater, including variations in input and removal

rates. These factors are considered below, followed by an evaluation of polonium

adsorption coefficients.

7.2.4. Particles and colloids in groundwater

Polonium has been found to be associated with colloids and particles in

groundwater, as well as being dissolved. This is expected because polonium is

very particle reactive, and particles are often composed of clays and compounds

of manganese and iron that have strong adsorption coefficients for polonium

(see Chapter 3). Not separating these from truly dissolved constituents can lead

to misleading conclusions regarding chemical behaviour. These species are

usually operationally defined by filtration methods, with particles being material

excluded from 0.2 μm or 0.45 μm filters, and colloids constituting material

passing this filter but not molecular filters or ‘ultrafilters’ of 5–20 kilodaltons.

Some large molecules, such as humic acids, are in the colloidal pool, while iron

and manganese oxyhydroxides and clays may be present at sizes that can be

either colloidal or particulate.

There have been several polonium speciation studies of groundwater

from Finland. In low ionic strength Ca–HCO

waters, 35–68% of

210

Po was

bound to particles, while more than 90% was bound to particles in higher

salinity Na–Cl waters that have high concentrations of iron, manganese and

humic acids [7.11]. The colloid fraction (between 5 kilodaltons and <0.45 μm)

contained approximately less than 10% [7.12]. Vesterbacka et al. [7.13] report

that, on average, 86% of

210

Pb in groundwater is found in the large particle

116

CHAPTER 7

fraction (>0.45 μm). However, for waters that are rich in organics, iron and

manganese, and which have high Fe:Mn ratios, the majority of

210

Po is found in

either the intermediate particle (100 kilodaltons and 0.45 μm) or small particle

(<100 kilodaltons) fractions.

Polonium, with such low natural elemental concentrations (due to its

relatively short half-life), clearly does not form its own particles but rather is

adsorbed onto, or incorporated into, other phases. While early laboratory

experiments suggested that polonium in simple solutions can form colloids [7.14],

it is present in extremely low concentrations compared to other colloid forming

elements and is highly reactive with surfaces; so that in natural environments, it

is likely to be associated with colloids of mixed composition [7.12].

Studies of groundwater from aquifers in Florida [7.15] and Nevada [7.16],

United States of America, provide comparisons between filtered and unfiltered

samples (see Fig. 7.4). Data cluster around the line for equal concentrations,

especially for the Nevada samples, although significant fractions of

210

Po appear

to be present on particles (data above the line). However, there are significant

deviations in the Florida samples. Filtering of particles from Nevada groundwater

removed 5–47% of

210

Po [7.16]. One of the highest concentrations recorded

(95 Bq/L) is for an unfiltered Florida groundwater; the filtered sample contained

17%

210

Po [7.17]. Such a high concentration suggests that particles with high

226

Ra were present. The most perplexing are some samples that appear to have

higher concentrations when filtered. This can only be a sampling or analytical

artefact.

The strong association of polonium with particles has a number of

implications for interpreting polonium behaviour. Particles can be mobile,

Source: See Refs [7.15, 7.16].

FIG. 7.4. Data for groundwater (filtered and unfiltered) from aquifers in Florida and Nevada,

United States of America.

117

FRESHWATER AND GROUNDWATER SYSTEMS

although not necessarily at the same rate as dissolved species. There might be

sampling biases in quantifying the particle load, since the particle load in samples

collected from wells might not be representative of aquifer groundwater; removal

of particles can occur in the filter pack around the well; or increased mobilization

of particles around the well can occur from pumping, resulting in enhanced flow

rates. Furthermore, if particle bound polonium is not separated, the amount of

dissolved species will be overestimated. In some cases, a significant fraction

of polonium can also be colloid bound. Overall, such effects should be further

investigated, especially where unusual occurrences of polonium occur. Routine

practice should include reporting data for both unfiltered and filtered samples,

and filtration should be done during collection, or as soon as possible after, since

association with particles can change during storage.

7.2.5. In situ adsorption coefficients

7.2.5.1. Quantifying polonium adsorption

As with most trace elements, the extent of

210

Po adsorption can be

determined from the analysis of aquifer rocks or laboratory experiments.

However, it is difficult to obtain a bulk average aquifer value for a process that

is expected to vary substantially between samples due to the heterogeneous

distribution of secondary phases. Furthermore, as discussed in Chapter 4, this has

been done for polonium on only a small number of soils, sediments and aquifer

materials. Porcelli [7.3] finds that:

“Studies involving decay-series systematics of nuclides in groundwater have

quantified parameters of water–rock interaction from direct measurements

of waters alone, with source terms and adsorption coefficients inferred from

comparison between the different radionuclides. Sampling of large volumes

or from a number of wells can obtain parameter values for larger volumes

of aquifer. In particular, the rate of recoil supply to groundwater, which

supplies all of the daughter nuclides at similar rates, and the evolution and

steady state concentrations of the short-lived nuclides can be determined.

Then bulk adsorption partition coefficients can be obtained.”

7.2.5.2. Calculating average aquifer adsorption coefficients

Porcelli [7.3] reports that mathematical treatments of simple aquifer models

have been extensively developed [7.18–7.23] and summarized in Refs [7.4, 7.24].

These provide operational rates of transport and of water–rock interaction

118

CHAPTER 7

processes, regardless of the actual mechanisms involved, and so complement

physical chemistry studies that can identify the controlling chemical processes.

These studies have so far largely considered steady state groundwater profiles;

that is, they assume that the conditions at each location within the aquifer do

not change, and so groundwater radionuclide concentrations are constant at each

location but may change along flow lines. Furthermore [7.3]:

“In general, the decay series systematics for groundwater behavior can

become quite complex, as there are a number of parameters that affect each

isotope, and the resulting distribution then affects the production of the

next isotope in the chain. However, the element with the simplest and most

predictable behavior is

222

Rn, an unreactive noble gas, and the distribution

of isotopes further along the

238

U decay series, including

210

Po, can be

estimated from consideration of the distribution of

222

Rn.

“The section of the

238

U decay series that is most directly relevant to

understanding

210

Po is below

226

Ra, as shown in Fig. [7.5]. For each nuclide,

the sum of dissolved and adsorbed atoms, that is the atoms that have

been released from aquifer solids and are chemically exchanging, can be

considered together as the mobile pool. Further, in most cases, radioisotope

concentrations will have reached a steady state concentration, where the

supply of the isotope from within the minerals and from decay of mobile

atoms of the parent will together equal the decay rate (that is, the activity)

of the mobile pool of the isotope. For the short-lived nuclides below

222

Rn,

it is generally the case (see Porcelli and Swarzenski, 2003 [Ref. [7.24]])

that for very short-lived nuclides, recoil is much more important as an

input than weathering, which can then be ignored. Also, the decay rates

of short-lived adsorbed and dissolved nuclides are generally much greater

than precipitation rates in groundwaters, and so this also can generally be

ignored

(Porelli, 2008 [Ref. [7.4]]). Of course, there are circumstances

when these factors do need to be considered, e.g. in rapidly weathering

horizons, or preferential rapid weathering of minerals rich in U (and so also

rich in all daughter nuclides) minerals, or where abrupt changes in aquifer

chemistry cause rapid precipitation. However, in the absence of evidence

for these, weathering and precipitation can be ignored.”

The models have been applied to the available data to obtain recoil supply

rates for

210

Po, as well as K

values for both

210

Pb and

210

Pb [7.3]. Porcelli [7.3]

states that the distribution of

210

Po can be evaluated by starting with

222

Rn. The

radionuclide

222

Rn does not precipitate or adsorb, so that groundwater (

222

Rn)

119

FRESHWATER AND GROUNDWATER SYSTEMS

activity is equivalent to

222

Rn recoil plus the production from mobile

226

Ra [7.4],

which is otherwise expressed as [7.3]:

( ) ( ) ( )

( )

222 226 226

222 266

W RW

Rn Ra Ra Kbe=+

(7.1)

where

is the ratio of rock mass to water mass per unit volume;

222

is the fraction of

226

Ra decays that recoil daughter atoms to the

groundwater;

(

226

Ra)

is the concentration of

226

Ra in the aquifer rock;

(

226

Ra)

is the concentration of dissolved

226

Ra;

and (K

226

) is the ratio of atoms of

226

Ra adsorbed to atoms in solution.

Porcelli [7.3] continues:

“The

222

Rn in solution will produce

218

Po, which will largely adsorb. In

addition, there will be recoil of

218

Po into solution from decay of

222

within the minerals. The decay rate of mobile

218

Po (adsorbed and dissolved)

will equal the input rates. Mobile

218

Po will then produce mobile

214

Po,

along with further recoil of

214

Pb, so that mobile

214

Pb activity will be equal

to that of dissolved

222

Rn plus the addition of two recoil fluxes. The same

will apply to

210

Pb (produced by decay of

214

Pb to the very short-lived

214

Bi,

then the very short-lived

214

Po), which will have an activity of adsorbed

and dissolved atoms of

222

Rn plus 3 recoil fluxes. Since

210

Bi and

210

Po are

Source: Figure 3 of Ref. [7.3].

FIG. 7.5. The

238

U decay series from

226

Ra to

206

Pb, showing the recoil fluxes from within the

aquifer solid and so supplying the mobile pool of dissolved and adsorbed radionuclides in the

aquifer (reproduced with permission courtesy of Elsevier).

120

CHAPTER 7

produced by low energy beta decay, there are no recoil contributions to

these nuclides (see Fig. [7.5]). Therefore, the activity of mobile

210

Pb,

210

and

210

Po will be equal.”

Combining these effects together gives

[7.3]:

( )

( ) ( )

( )

226 222

210

3 Ra Rn

be +

(7.2)

The activity of mobile

210

Pb is equal to that of mobile

210

Po, so that [7.3]:

( )

210

(7.3)

If the production of

222

Rn is largely from recoil, so that the term for

production from adsorbed

226

Ra can be neglected, then [7.3]:

( ) ( )

222 226

222

Rn Rabe=

(7.4)

and

( ) ( )

210

210 222

Pb Rn

(7.5)

Porcelli [7.3] concludes by stating that there may be some circumstances in

which there is sufficient

226

Ra adsorbed to dominate

222

Rn production, which will

reduce K

210Pb

by a factor of four. However, the range in K

210Pb

is substantially

greater and so this will generally not be a major consideration. Moreover [7.3]:

“As discussed above, the values for K

210Po

and K

210Pb

are equal to the

number of atoms adsorbed divided by the number of atoms in solution

within any volume of aquifer (i.e. in units of mol/mol). In order to translate

this into the more familiar values of atoms per kg of rock divided by atoms

per liter of groundwater (i.e. L/kg), the porosity of the aquifer must be

known. Since this is often not reported or not well constrained, the values

calculated below are for moles adsorbed on aquifer solids relative to moles

in solution, or mol/mol.”

121

FRESHWATER AND GROUNDWATER SYSTEMS

Further information can be found in Ref. [7.3], including site specific

adsorption coefficients for groundwater in Australia, Brazil, Scandanavia and the

United States of America.

7.3. SURFACE WATER

The freshwater environment encompasses lotic (moving) water, such as

rivers and streams, and lentic water bodies (standing), such as lakes and ponds.

The concentrations of

210

Po within moving water depend on the environmental

conditions in the catchment area and reflecte both natural and anthropogenic

sources. The concentrations throughout the year is expected to vary, since

precipitation carries particulates from the surrounding areas into the water body.

Polonium in water can transfer to biota such as plants, phytoplankton,

zooplankton, invertebrates, fish, water based amphibians, crustaceans, mammals

and birds with dietary components from the aquatic environment (see Fig. 7.6).

FIG. 7.6. Conceptual model of an aquatic ecosystem.

122

CHAPTER 7

The interactions among the various abiotic and biotic environments

depends on the location. The International Union of Radioecology (IUR) suggests

that interaction matrices provide a convenient way to consider information on

relevant processes in the aquatic environment in a synthesized and structured

way

[7.25]. Each environment has its own unique chemical, physical and

biological characteristics. Nonetheless, the generic interaction matrices for the

aquatic environment provided by the IUR

[7.25] and the associated generic

discussions provide a convenient checklist for initial discussions of the relevant

processes.

The aquatic environment is characterized by the flow regime, water and

sediment quality, and local species, including primary producers, and primary

and secondary consumers. The IUR identifies the following main components for

the aquatic environment:

(a) The atmosphere, which is the region above the surface;

(b) Water, including surface water bodies (e.g. rivers, streams and lakes);

suspended in water;

(d) Deposited matter (sediment), which is particulate matter (inorganic and

organic) deposited at the bottom of rivers, streams and lakes, including

interstitial water;

(e) Primary producers, which are autotrophic organisms, such as phytoplankton,

macrophytes and aquatic plants;

(f) Primary consumers, which are animals that feed on primary producers

(e.g. zooplankton and macrobenthos);

(g) Secondary consumers, which are animals that feed on primary consumers

(e.g. omnivorous fish);

(h) Decomposers, which comprise microflora and protozoa.

7.3.1. Sources in surface fresh water

The sources of lead and polonium in surface waters include:

— Atmospheric deposition of resuspended dusts or aerosols;

— Fluvial inputs;

— In situ decay from progenitors (particularly

226

Ra in sediment; depending on

mixing rates, the interceding

222

Rn is often likely to decay before diffusing

to the water surface);

— Wash-off from the terrestrial environment;

— Anthropogenic inputs (e.g. phosphate ore processing operation and uranium

mining).

123

FRESHWATER AND GROUNDWATER SYSTEMS

Radionuclides dispersed in the environment can be deposited into surface

water or onto surfaces within the watershed, which, in the latter case, can

represent a long term source of radionuclides for freshwater ecosystems within

the affected region.

7.3.2. Levels of

210

Po in the freshwater environment

The activity of

210

Po in normally oxic waters is in the range of 1–5 mBq/L

and is higher, up to 17 mBq/L, in seasonally anoxic ponds [7.26]. Table 7.1

provides a summary of

210

Po measurements in the freshwater environment of

rivers and lakes. The amount of data for freshwater biota is less than for the

marine environment (see Table 7.2).

A similar pattern in freshwater biota has been observed when compared to

marine biota (see Chapter 8), in that higher concentrations have been reported

in the soft tissue of molluscs than in shells, whereas

210

Pb has been detected

in higher levels in the shells. A mollusc shell is largely composed of inorganic

substances, such as calcium carbonate, while the exoskeleton of prawn consists

of organic substances, such as chitin. As

210

Po has an affinity for organic moieties,

this accounts for the relatively higher concentration of

210

Po in the exoskeleton of

prawn [7.28].

7.3.3. Behaviour of

210

Po in standing water bodies

The behaviour of

210

Po in a water body is influenced by a combination

of physical and chemical processes. Within a standing water body (e.g. lake),

there is uptake of

210

Pb and

210

Po by particulates (particularly biomass) in the

water column. As these particulates settle to the bottom sediment, polonium is

scavenged from the water column. Within a few months, equilibrium between

210

Pb and

210

Po can be established in the sediment. However, changing redox

conditions in the sediment and water column also influence the polonium levels.

The chemical form (speciation) of sulphur in sediment is redox dependent and

can, in turn, affect

210

Pb and

210

Po behaviour (lead forms sulphides, e.g. PbS).

The cycling of polonium in sediment can be affected by microbial activity in the

bottom sediment, as the release of polonium can be enhanced in sulphur enriched

bacteria

[7.42]. Polonium has also been shown to become volatile in both fresh

and marine waters by the action of microorganisms

[7.43, 7.44] (see Section 3.2).

In addition to cycling in sediment, depending on the characteristics

of the lake, there can be zones within the water column that become anoxic

during periods of the year, which influences the behaviour of polonium in the

water column. Polonium has several oxidation states (−2, +2, +4 and +6), of

which the insoluble tetravalent state Po(IV) is the most stable in oxic aqueous

124

CHAPTER 7

solutions [7.45]. However, it can hydrolyse and form hydroxides. Polonium is

influenced by the cycling of iron and manganese

[7.26, 7.36, 7.46]. As oxygen

levels decrease, iron and manganese are reduced and enter into solution, bringing

with them transition metals, including

210

Pb and

210

Po, which are adsorbed onto

TABLE 7.1. BASELINE CONCENTRATIONS OF

210

Po IN SURFACE WATER

AND SEDIMENTS

Location

210

Po in water

(mBq/L)

210

Po in sediments

(Bq/kg)

Ref.

Kaveri River, India

0.77–1.27 14.4–26.5 [7.28]

Mutharasanallur Pond, India

1.4 60 [7.29]

Rostov, Russian Federation

1.8 × 10

−3

—

[7.30]

Vistula River, Poland 1.94–3.21

—

[7.31]

Tributaries of the Vistula River, Poland 2.15–6.03

—

[7.31]

Oder River, Poland 1.46–2.39

—

[7.32]

Tributaries of the Oder River, Poland 1.02–3.64

—

[7.32]

Saskatchewan, Canada (control site) —

240

[7.33]

Saskatchewan, Canada (unimpacted lake

near the McArthur River site)

2 50

[7.34]

Crystal Lake, Wisconsin, USA 1.6 —

[7.35]

Bickford Reservoir, Massachusetts, USA 1.3 —

[7.36]

Four Norwegian lakes 1.6–2 —

[7.37]

South Alligator River, Australia

(control site)

1.94 filtered

2.51 particulate

—

[7.38]

Fresh weight unless otherwise stated.

Adapted from Ref. [7.27].

—: data not available.

May be influenced by anthropogenic sources.

Bq/kg (DW).

125

FRESHWATER AND GROUNDWATER SYSTEMS

TABLE 7.2. CONCENTRATIONS OF

210

Po IN FRESHWATER BIOTA

SPECIES

Sample Part

210

(Bq/kg)

Concentration ratios

(L/kg)

Ref.

Plankton

19–29 (2.1–2.5) × 10

[7.28]

Water hyacinth

(Eichhornia crassipes)

Shoot

Root

2.3–6.5

6.7–31

(2.7–4.6) × 10

(0.73–2.2) × 10

[7.28]

[7.29]

Bivalve mollusc

(Lamellidens marginalis)

Shell

Soft tissue

1.2–4.5

53–106

(0.86–5.3) × 10

(3.8–8.4) × 10

[7.28]

[7.29]

Bivalve mollusc

(Anodonta cygnea)

Shell

Soft tissue

0.1

19.3

—

6.2 × 10

[7.39]

Bivalve mollusc

(Anodonta sp.)

Soft tissue

Shell

Whole

5.7

2.5

4.7

2.1 × 10

9.4 × 10

1.7 × 10

[7.37]

Bivalve mollusc

(Velesunio angasi)

Soft tissue 302–416

4.7 × 10

[7.38]

Crustacean

(crab, prawn)

Exoskeleton

Muscle

8.6–16.6

12–20

(0.81–1.2) × 10

(0.93–1.6) × 10

[7.28]

[7.29]

Snail

(Pila virens)

Shell

Soft tissue

0.2–3.9

33–46

(0.14–5.3) × 10

(2.6–4.2) × 10

[7.28]

[7.29]

Fish

(Mystus vittatus,

Orechromis mossambicus

Puntius chola)

Bone

Muscle

1.3–17

1.9–27

(0.13–1.2) × 10

(0.22–1.9) × 10

[7.28]

[7.29]

Fish

(perch, bream)

—

0.36–0.43 —

[7.30]

Fish

Bone 0.37–0.43

—

[7.40]

Fish Whole

Edible

1.0–6.5

0.08–1.9

(0.63–9.3) × 10

—

[7.37]

Fish

(Cyprinus carpio,

Sander lucioperca)

Bone

Muscle

Liver

8–14

2–7

11–183

—

[7.41]

Fresh weight unless otherwise stated.

Adapted from Ref. [7.27].

Bq/kg (DW).

—: data not available.

L/kg (DW).

126

CHAPTER 7

oxides. The insoluble Po(IV) is also reduced to Po(II) near the same redox

potential that Mn(IV) is reduced to Mn(II)

[7.36. 7.46]. The dissolved ions diffuse

upwards and, as they become oxidized again, they form precipitates. Thus,

changing redox conditions can have a significant effect on the levels of polonium

in a water body. As a result, much higher levels of

210

Po have been measured in

the anoxic zone of a water body (hypolimnion) compared to the epilimnion, and

an enrichment of

210

Po compared to

210

Pb is observed [7.26, 7.36, 7.46].

Under certain conditions, polonium can diffuse out of the sediments and,

the sediments can act as a sink for polonium.

7.3.4. Quantification of water and sediment quality

The interaction of radionuclides with bottom sediments and suspended

particles is controlled by many environmental processes, such as sedimentation

and resuspension (see Fig.

7.7).

Radionuclide

in water (dissolved)

Radionuclide

particulate form (fast

exchange process)

Radionuclide

particulate form (slow

exchange process)

Source: Figure 2 of Ref. [7.47].

Note: The fast processes are marked in the subscript by ‘f’, while the slow processes are

marked by ‘s’.

FIG. 7.7. Schematic structure of contaminant fluxes and environmental compartments

involved in the processes of radionuclide migration to and from bottom sediment.

127

FRESHWATER AND GROUNDWATER SYSTEMS

Monte et al. [7.47] describe three active compartments:

— Dissolved radionuclide in water (C);

— Particulate radionuclide: rapid exchange component (D

);

— Particulate radionuclide: slow exchange component (D

Accordingly

[7.47]:

“The fluxes (Bq·s

−1

) from a compartment can be calculated as the product

of the total amount of radionuclide in the compartment (Bq) multiplied by

the ‘rates’ (s

−1

) of migration (K

, K

“The seven radionuclide fluxes can be schematized as follows:

(a) Radionuclide fluxes from dissolved form to particulate form and vice

versa — rapid exchange processes (K

·C and K

·D

);

(b) Radionuclide fluxes from D

to D

and vice-versa (K

·D

and K

·D

);

and vice-versa (K

·C and

·D

);

(d) Radionuclide irreversible burial in inactive sediments (K

·D

).”

Under oxic conditions, an assessment of the interaction of dissolved

radionuclides with solid particles in suspension or deposited, is often based on

the K

concept, assuming a presence of an equilibration between the dissolved

, Bq/m

) and the adsorbed phases (C

, Bq/kg) of a radionuclide [7.47]:

(7.6)

There are very limited data available to determine a K

for polonium.

Ciffroy et al. [7.48] did not find a sufficient database of information. One study of

polonium in a pond derived an overall K

of 4.3 × 10

L/kg (FW) [7.29], whereas

values in a river were estimated at 1.5 × 10

L/kg (FW) for suspended matter

to dissolved and 3

× 10

L/kg (FW) for bottom sediment to dissolved [7.49].

Application of these values should be treated with caution, as they are based on

limited information and K

values can vary widely.

7.3.5. Freshwater aquatic biota

The accumulation of

210

Po by aquatic biota is a dynamic process, with

many pathways of exposure being involved. However, it is often simplified by

128

CHAPTER 7

assuming that abiotic and biotic media have reached steady state conditions

and a concentration ratio (CR) can thus be established that relates the biota

concentration to the concentration in the abiotic reference medium (water or

sediment). These CRs can also be derived for those non-aquatic animals that

depend on the aquatic environment, including mammals and birds.

There are two basic sets of data on radionuclide accumulation in aquatic

organisms. The first approach is based on assessments of CRs for edible tissues

of freshwater biota. Within the second approach, intended mainly for assessments

of dose to biota, activity concentrations of radionuclides are calculated (or

measured) for the whole organism. Yankovich et

al. [7.50] describe methods for

converting tissue concentrations to whole body values for a variety of species

using CRs. The empirically derived CRs for lead and polonium are summarized

in Table

7.3.

Until fairly recently, the Handbook of Parameter Values for the Prediction

of Radionuclide Transfer in Temperate Environments

[7.65] provided the

basis for environmental assessments. Since that publication was released, new

datasets have become available, and an update of that publication was considered

appropriate. Two large scale reviews, in the framework of the international

IAEA projects, Environmental Modelling for Radiation Safety (EMRAS

I and

EMRAS

II), compiled the available data covering both transfers to human

foodstuffs and the available quantitative data on the transfer of radionuclides to

wildlife. The update was accomplished with the publication of the Handbook of

Parameter Values for the Prediction of Radionuclide Transfer in Terrestrial and

Freshwater Environments

[7.66]. Much of the earlier data on biological uptake

for many radionuclides was often on specific tissues and not the whole organism.

The recent reviews were much wider in context, based on many sources and,

hence, many more data for polonium were compiled. The CRs for

210

Pb and

210

are presented in Table 7.3 for broad wild biota groups and comprise a subset from

Handbook of Parameter Values for the Prediction of Radionuclide Transfer to

Wildlife

[7.67] — an overview of which is provided in Ref. [7.51].

Although Ref. [7.66] represents a key reference for radioecologists,

modellers and authorities by providing data for environmental impact

assessments, it includes just a few freshwater CRs for lead and polonium. Hence,

it has limited applicability for the impact assessment of naturally occurring

radioactive material on freshwater species and their contamination of freshwater

food.

From the available tissue data (see Table 7.2), the distribution of

210

Po within an organism shows higher concentrations in the kidney, liver

and other organs in the digestive system. A higher accumulation of

210

Po is

found in the soft tissue of molluscs than in shells, whereas

210

Pb has been

detected at higher levels in shells. In the soft tissue of an organism,

210

Po is

129

FRESHWATER AND GROUNDWATER SYSTEMS

TABLE 7.3.

210

Pb AND

210

Po CONCENTRATION RATIO (CR

WO-WATER

) VALUES FOR WILDLIFE GROUPS IN

FRESHWATER ECOSYSTEMS

Wildlife group

wo-water

Ref.

AM ASD GM GSD Min. Max.

210

Amphibian 2 5.3 —

—

1.7 8.9 [7.52]

Crustacean 5 3.9 × 10

4.7 × 10

2.5 × 10

2.6 —

—

[7.53]

Fish

Benthic feeding

Forage

Piscivorous

379

148

201

2.5 × 10

1.8 × 10

2.6 × 10

3.5 × 10

7.0 × 10

6.3 × 10

6.2 × 10

7.8 × 10

8.7 × 10

4.8 × 10

9.9

1.4 × 10

4.3

5.0

4.0

3.8

2.0

3.2

2.0

8.3

7.5 × 10

3.5 × 10

5.7 × 10

[7.52–7.58]

[7.52–7.55, 7.58]

[7.52, 7.53, 7.55–7.57]

[7.52–7.55]

Mollusc (bivalve) 32 6.0 × 10

—

2.3 × 10

4.0 1.1 × 10

2.9 × 10

[7.38, 7.54, 7.59, 7.60]

Reptile 12 4.4 × 10

6.2 × 10

2.5 × 10

2.9 1.3 × 10

1.9 × 10

[7.61]

Vascular plant 21 6.2 × 10

7.0 × 10

4.1 × 10

2.5 1.3 × 10

1.9 × 10

[7.52, 7.60]

130

CHAPTER 7

TABLE 7.3.

210

Pb AND

210

Po CONCENTRATION RATIO (CR

WO-WATER

) VALUES FOR WILDLIFE GROUPS IN

FRESHWATER ECOSYSTEMS (cont.)

Wildlife group

wo-water

Ref.

AM ASD GM GSD Min. Max.

210

Crustacean 12 8.3 × 10

7.0 × 10

6.3 × 10

2.1 1.2 × 10

1.6 × 10

[7.28, 7.29, 7.53]

Fish

Benthic feeding

Forage

Piscivorous

203

2.0 × 10

1.6 × 10

7.6 × 10

1.3 × 10

6.6 × 10

4.4 × 10

1.2 × 10

6.7 × 10

5.9 × 10

5.7 × 10

4.2 × 10

2.6 × 10

4.8

4.2

3.0

6.1

4.9 × 10

6.3 × 10

1.3 × 10

4.9 × 10

3.7 × 10

1.9 × 10

2.6 × 10

3.7 × 10

[7.28, 7.29, 7.49, 7.53, 7.54, 7.62]

[7.28, 7.29, 7.49, 7.53, 7.54]

[7.28, 7.29, 7.53, 7.62]

[7.53, 7.54, 7.63]

Mollusc

Bivalve

147

141

1.2 × 10

1.3 × 10

5.2 × 10

4.9 × 10

1.1 × 10

1.2 × 10

1.5

1.7 × 10

[7.28, 7.29, 7.38, 7.54, 7.62, 7.64]

Reptile 7 3.6 × 10

2.3 × 10

3.1 × 10

1.8 1.5 × 10

7.3 × 10

[7.61]

Vascular plant 31 2.0 × 10

1.5 × 10

1.6 × 10

2.0 5.5 × 10

4.6 × 10

[7.28, 7.29, 7.62]

Source:

See Ref. [7.51].

Note: CR

wo-water

is the activity concentration in the whole organism (in Bq/kg, FW) divided by the activity concentration in the water

(in Bq/kg, DW). AM — arithmetic mean; ASD — arithmetic standard deviation; CR — concentration ratio; GM — geometric mean;

GSD — geometric standard deviation.

—: data not available.

131

FRESHWATER AND GROUNDWATER SYSTEMS

typically found at higher levels than

210

Pb. Since

210

Po concentrations in soft

tissue are normally higher than those in bones and comprise most of the edible

tissues of freshwater species, the data in Table 7.3 can also be used to assess

210

Po concentrations in human foodstuffs. For polonium, the ratio of whole

body to tissue is expected to be approximately 1.1 for freshwater fish and 2 for

mammals [7.50]. However, there is a paucity of data for most species. The use

of CRs in Table 7.3 represents a cautious approach to estimating

210

Po in edible

tissues when sufficient data are unavailable. Additional reasons supporting the

use of

210

Po CRs for assessment of human foodstuff contamination include the

following [7.67]:

(a) For aquatic ecosystems, the whole organism CR

wo-media

values for bivalve

molluscs, large crustaceans and marine gastropods do not include the shell.

This is consistent with commonly used dosimetric approaches.

(b) For vertebrate wildlife groups, whole organism CR

wo-media

values typically

do not include the gastrointestinal tract contents, although there may be

some exceptions, such as when animals have been monitored live and in

the case of small fish.

Table 7.3 also provides CRs for

210

Pb, since it serves as a source of

210

in wildlife. Such information is also often valuable for modelling purposes. The

data in Table 7.3 clearly show that

210

Po CRs tend to be one to two orders of

magnitude greater than those for

210

Pb, thus demonstrating higher mobility in

freshwater ecosystems. The highest

210

Po CRs are in bivalve molluscs, followed

by crustaceans and some foraging species, indicating the sensitivity of some

freshwater organisms to contamination by polonium.

In summary,

210

Po in groundwater occurs over a wide range of

concentrations. Once the differences due to release from aquifer minerals of

222

Rn by recoil are considered, there are very large differences due to adsorption

210

Po onto surfaces. Bulk adsorption coefficients are difficult to predict

because of wide variations in grain size and mineralogy over even very small

spatial scales. Therefore, while earlier studies can provide some guidance on

what might be expected, each aquifer needs to be characterized individually.

In order to do so, accompanying data on

222

Rn are valuable for characterizing

adsorption. Limited data also suggest that the concentrations of

210

Po on particles

and colloids can be important under some circumstances. Future studies on

the chemistry of polonium in dynamic groundwater regions, such as oxic and

anoxic transitions and areas of salt water intrusion, are necessary. The behaviour

of polonium within the vadose zone, which is most vulnerable to anthropogenic

inputs, is also poorly understood.

132

CHAPTER 7

The levels of

210

Po in surface water in freshwater systems are influenced by

in situ decay, surface runoff, fluvial inputs and atmospheric deposition. Within

a standing water body,

210

Po is removed through the settling of particulates, and

its behaviour in the sediment is influenced by the cycling of iron and manganese

as the redox conditions change in the sediment. This is also an important process

if there are seasonal anoxic zones within the water body. More research is still

needed on the partitioning of

210

Po onto particulates in the surface water column

to fully understand the cycling of

210

Po and how it relates to manganese, iron and

sulphur. A more comprehensive database of information needs to be developed to

better understand the dynamics of

210

Po in the freshwater food web.

REFERENCES TO CHAPTER 7

[7.1] MONTE, L., et al., Review and assessment of models used to predict the fate of

radionuclides in lakes, J. Environ. Radioact.

69 (2003) 177–205.

[7.2] INTERNATIONAL ATOMIC ENERGY AGENCY, Quantification of Radionuclide

Transfer in Terrestrial and Freshwater Environments for Radiological Assessments,

IAEA-TECDOC-1616, IAEA, Vienna

(2009).

[7.3] PORCELLI, D., A method for determining the extent of bulk

210

Po and

210

Pb adsorption

and retardation in aquifers, Chem. Geol. 382 (2014) 132–139.

[7.4] PORCELLI, D., “Investigating groundwater processes using U- and Th-series

nuclides”, U/Th Series Radionuclides in Aquatic Systems (KRISHNASWAMI, S.,

COCHRAN, J.K., Eds), Radiation in the Environment, Vol.

13, Elsevier, Amsterdam

(2008)

105–153.

[7.5] KIGOSHI, K., Alpha-recoil thorium-234: Dissolution into water and the uranium-234/

uranium-238 disequilibrium in nature, Science 173 (1971) 47–48.

[7.6] FLEISCHER, R.L., RAABE, O.G., Recoiling alpha-emitting nuclei: Mechanisms for

uranium-series disequilibrium, Geochim. Cosmochim. Acta

42 (1978) 973–978.

[7.7] FOCAZIO, M.J., et al., Occurrence of Selected Radionuclides in Ground Water Used

for Drinking Water in the United States: A Reconnaissance Survey, 1998,

Water-Resources Investigations Report

00-4273, United States Geological Survey,

Reston, VA

(2001).

[7.8] RUBERU, S.R., LIU, Y.-G., KUSUM PERERA, S., Occurrence and distribution of

210

Pb and

210

Po in selected California groundwater wells, Health Phys. 92

(2007)

432–441.

[7.9] SALONEN, L., HUIKURI, P., “Elevated levels of uranium series radionuclides in

private water supplies in Finland”, High Levels of Natural Radiation and Radon Areas:

Radiation Dose and Health Effects (Proc. 5th Int. Conf. Munich, 2000), Vol.

II: Poster

Presentations (2002)

87–91.

[7.10] DICKSON, B.L., HERCZEG, A.L., Naturally occurring radionuclides in acid: Saline

groundwaters around Lake Tyrrell, Victoria, Australia, Chem. Geol.

96 (1992) 95–114.

133

FRESHWATER AND GROUNDWATER SYSTEMS

[7.11] VAARAMAA, K., LEHTO, J., ERVANNE, H., Soluble and particle-bound

234,238

226

Ra and

210

Po in ground waters, Radiochim. Acta 91 (2003) 21–28.

[7.12] LEHTO, J., KELOKASKI, P., VAARAMAA, K., JAAKKOLA, T., Soluble and

particle-bound Po-210 and Pb-210 in groundwaters, Radiochim. Acta

(1999)

149–155.

[7.13] VESTERBACKA, P., HÄMÄLÄINEN, K., LEHTO, J., The effect of water treatment

on the presence of particle-bound

210

Po and

210

Pb in groundwater, Radiochim. Acta 93

(2005)

291–296.

[7.14] MORROW, P.E., DELLA ROSA, R.J., CASARETT, L.J., MILLER, G.J.,

Investigations of the colloidal properties of polonium-210 solutions by using molecular

filters, Radiat. Res. Suppl.

5 (1964) 1–15.

[7.15] BURNETT, W.C., COWART, J.B., CHIN, P.A., “Polonium in the surficial aquifer of

West Central Florida”, Radon, Radium, and Other Radioactivity in Ground Water:

Hydrogeologic Impact and Application to Indoor Airborne Contamination (Proc.

NWWA Conf. Somerset, NJ, 1987) (GRAVES, B., Ed.), Lewis Publishers, Chelsea, MI

(1987)

251–269.

[7.16] SEILER, R.L.,

210

Po in Nevada groundwater and its relation to gross alpha radioactivity,

Ground Water

49 (2011) 160–171.

[7.17] COWART, J.B., BURNETT, W.C., CHIN, P.A., HARADA, K., “Occurrence of

210

in natural waters in Florida”, Trace Substances in Environmental Health: XXI (Proc.

Univ. Missouri’s 21st Conf. St. Louis, MO, 1987) (HEMPHILL, D.D., Ed.), University

of Missouri Press, Columbia, MO (1987) 172–185.

[7.18] KRISHNASWAMI, S., GRAUSTEIN, W.C., TUREKIAN, K.K., DOWD, J.F.,

Radium, thorium and radioactive lead isotopes in groundwaters: Application to the

situ determination of adsorption–desorption rate constants and retardation factors,

Water Resour. Res.

18 (1982) 1663–1675.

[7.19] DAVIDSON, M.R., DICKSON, B.L., A porous flow model for steady state transport

of radium in groundwater, Water Resour. Res.

22 (1986) 34–44.

[7.20] KU, T.-L., LUO, S., LESLIEE, B.W., HAMMOND, D.E., “Decay-series disequilibria

applied to the study of rock-water interaction and geothermal systems”, Uranium-

series Disequilibrium: Applications to Earth, Marine, and Environmental Sciences

(IVANOVICH, M., HARMON, R.S., Eds), Clarendon Press, Oxford (1992)

631–668.

[7.21] LUO, S.D., KU, T.L., ROBACK, R., MURRELL, M., McLING, T.L., In-situ

radionuclide transport and preferential groundwater flows at INEEL (Idaho):

Decay-series disequilibrium studies, Geochim. Cosmochim. Acta

64 (2000) 867–881.

[7.22] TRICCA, A., PORCELLI, D., WASSERBURG, G.J., Factors controlling the

groundwater transport of U, Th, Ra, and Rn, J. Earth Syst. Sci.

109 (2000) 95–108.

[7.23] TRICCA, A., WASSERBURG, G.J., PORCELLI, D., BASKARAN, M., The transport

of U- and Th-series nuclides in a sandy unconfined aquifer, Geochim. Cosmochim.

Acta

65 (2001) 1187–1210.

[7.24] PORCELLI, D., SWARZENSKI, P.W., The behavior of U- and Th-series nuclides in

groundwater, Rev. Mineral. Geochem.

52 (2003) 317–361.

134

CHAPTER 7

[7.25] INTERNATIONAL UNION OF RADIOECOLOGY, Recommendations for Improving

Predictions of the Long-term Behaviour of

Cl,

Tc,

237

Np and

238

U, IUR

Report

6, IUR, Warrington (2006).

[7.26] KIM, G., KIM, S.-J, HARADA, K., SCHULTZ, M.K., BURNETT, W.C., Enrichment

of excess

210

Po in anoxic ponds, Environ. Sci. Technol. 39 (2005) 4894–4899.

[7.27] COPPIN, F., ROUSSEL-DEBET, S., Comportement du

210

Po en milieu terrestre: revue

bibliographique, Radioprot. 39 (2004) 39–58.

[7.28] SHAHEED, K., SOMASUNDARAM, S.S.N., SHAHUL HAMEED, P., IYENGAR,

M.A.R., A study of polonium-210 distribution aspects in the riverine ecosystem of

Kaveri, Tiruchirappalli, India, Environ. Pollut. 95 (1997) 371–377.

[7.29] SHAHUL HAMEED, P., SHAHEED, K., SOMASUNDARAM, S.S.N., A study on

distribution of natural radionuclide polonium-210 in a pond ecosystem, J. Biosci.

(1997)

627–634.

[7.30] LADINSKAYA, L.A., PARFENOV, Yu.D., POPOV, D.K., FEDOROVA, A.V.,

210

and

210

Po content in air, water, foodstuffs, and the human body, Arch. Environ.

Health

27 (1973) 254–258.

[7.31] SKWARZEC, B., JAHNZ, A., The inflow of polonium

210

Po from Vistula river

catchments area, J. Environ. Sci. Health A

42 (2007) 2117–2122.

[7.32] SKWARZEC, B., TUSZKOWSKA, A., Inflow of

210

Po from the Odra River catchment

area to the Baltic Sea, Chem. Anal.

53 (2008) 809–820.

[7.33] THOMAS, P., LIBER, K., An estimation of radiation doses to benthic invertebrates

from sediments collected near a Canadian uranium mine, Environ. Int.

(2001)

341–353.

[7.34] ENVIRONMENT CANADA, Priority Substances List Assessment Report: Releases

of Radionuclides from Nuclear Facilities (Impact on Non-Human Biota), Government

of Canada, Ottawa

(2003).

[7.35] TALBOT, R.W., ANDREN, A.W., Seasonal variations of

210

Pb and

210

Po concentrations

in an oligotrophic lake, Geochim. Cosmochim. Acta

48 (1984) 2053–2063.

[7.36] BENOIT, G., HEMOND, H.F., Polonium-210 and lead-210 remobilization from lake

sediments in relation to iron and manganese cycling, Environ. Sci. Technol.

(1990)

1224–1234.

[7.37] GJELSVIK, R., BROWN, J. (Eds), Po-210 and Other Radionuclides in Terrestrial and

Freshwater Environments, NKS-181, Nordic Nuclear Safety Research,

Roskilde (2009).

[7.38] RYAN, B., BOLLHÖFER, A., MARTIN, P., Radionuclides and metals in freshwater

mussels of the upper South Alligator River, Australia, J. Environ. Radioact.

(2008)

509–526.

[7.39] STEPNOWSKI, P., SKWARZEC, B., A comparison of

210

Po accumulation in molluscs

from the southern Baltic, the coast of Spitsbergen and Sasek Wielki Lake in Poland, J.

Environ. Radioact.

49 (2000) 201–208.

[7.40] PARFENOV, Yu.D., Polonium-210 in the environment and in the human organism, At.

Energy Rev.

12 (1974) 75–143.

[7.41] LaROCK, P., HYUN, J.-H., BOUTELLE, S., BURNETT, W.C., HULL, C.D., Bacterial

mobilization of polonium, Geochim. Cosmochim. Acta 60 (1996) 4321–4328.

135

FRESHWATER AND GROUNDWATER SYSTEMS

[7.42] SKIPPERUD, L., JØRGENSEN, A.G., HEIER, L.S., SALBU, B., ROSSELAND,

B.O., Po-210 and Pb-210 in water and fish from Taboshar uranium mining Pit Lake,

Tajikistan, J. Environ. Radioact.

123 (2013) 82–89.

[7.43] MOMOSHIMA, N., SONG, L.-X., OSAKI, S., MAEDA, Y., Formation and emission

of volatile polonium compound by microbial activity and polonium methylation with

methylcobalamin, Environ. Sci. Technol.

35 (2001) 2956–2960.

[7.44] MOMOSHIMA, N., SONG, L.-X., OSAKI, S., MAEDA, Y., Biologically induced Po

emission from fresh water, J. Environ. Radioact.

63 (2002) 187–197.

[7.45] FIGGINS, P.E., The Radiochemistry of Polonium, NAS-NS 3037, National Academy

of Sciences, Washington, DC (1961).

[7.46] BALISTRIERI, L.S., MURRAY, J.W., PAUL, B., The geochemical cycling of stable

Pb,

210

Pb, and

210

Po in seasonally anoxic Lake Sammamish, Washington, USA,

Geochim. Cosmochim. Acta

59 (1995) 4845–4861.

[7.47] MONTE, L., PERIAÑEZ, R., BOYER, P., SMITH, J.T., BRITTAIN, J.E., “Physical

processes in freshwater ecosystems”, Quantification of Radionuclide Transfer in

Terrestrial and Freshwater Environments for Radiological Assessments,

IAEA-TECDOC-1616, IAEA, Vienna (2009)

419–434.

[7.48] CIFFROY, P., DURRIEU, G., GARNIER, J.-M., “Distribution of radionuclides

between solid and liquid phases in freshwaters”, ibid., pp

441–471.

[7.49] CARVALHO, F.P., OLIVEIRA, J.M., LOPES, I., BATISTA, A., Radionuclides from

past uranium mining in rivers of Portugal, J. Environ. Radioact.

98 (2007) 298–314.

[7.50] YANKOVICH, T.L., et al., Whole-body to tissue concentration ratios for use in biota

dose assessments for animals, Radiat. Environ. Biophys.

49 (2010) 549–565.

[7.51] HOWARD, B.J., et al., The IAEA handbook on radionuclide transfer to wildlife, J.

Environ. Radioact.

121 (2013) 55–74.

[7.52] YANKOVICH, T.L., Compilation of Concentration Ratios for Aquatic Non-human

Biota Collected by the Canadian Power Reactors Sector, CANDU Owners Group

Inc.

(2010).

[7.53] MARTIN, P., HANCOCK, G.J., JOHNSTON, A., MURRAY, A.S., Bioaccumulation

of Radionuclides in Traditional Aboriginal Foods from the Magela and Cooper Creek

Systems, Research Report

11, Supervising Scientist for the Alligator Rivers Region,

Australian Government Publishing Services, Canberra

(1995).

[7.54] BROWN, J.E., GJELSVIK, R., HOLM, E., ROOS, P., SAXEN, R., OUTOLA, I.

(Eds), Filling Knowledge Gaps in Radiation Protection Methodologies for Non-human

biota: Final Summary Report, NKS-187, Nordic Nuclear Safety Research,

Roskilde (2009).

[7.55] AHMAD, M.K., ISLAM, S., RAHMAN, M.S., HAQUE, M.R., ISLAM, M.M., Heavy

metals in water, sediment and some fishes of Buriganga River, Bangladesh, Int. J.

Environ. Res.

4 (2010) 321–332.

[7.56] MALIK, N., BISWAS, A.K., QURESHI, T.A., BORANA, K., VIRHA, R.,

Bioaccumulation of heavy metals in fish tissues of a freshwater lake of Bhopal,

Environ. Monit. Assess.

160 (2010) 267–276.

136

CHAPTER 7

[7.57] MOHAMED, F.A.S., Bioaccumulation of selected metals and histopathological

alterations in tissues of Oreochromis niloticus and Lates niloticus from Lake Nasser,

Egypt, Glob. Vet.

2 (2008) 205–218.

[7.58] ÖZTÜRK, M., ÖZÖZEN, G., MINARECI, O., MINARECI, E., Determination of

heavy metals in fish, water and sediments of Avsar Dam Lake in Turkey, Iran. J.

Environ. Health Sci. Eng.

6 (2009) 73–80.

JOHNSTON, A., MURRAY, A., MARTEN, R., MARTIN, P., PETTERSON, H.,

“Uranium series radionuclide concentrations in significant Aboriginal foods”, Alligator

Rivers Region Research Institute, Research Report

1983–84, Supervising Scientist for

the Alligator Rivers Region, Australian Government Publishing Services, Canberra

(1984)

43–44.

[7.59] ENGDAHL, A., TERNSELL, A., HANNU, S., Oskarshamn Site Investigation:

Chemical Characterisation of Deposits and Biota, SKB

P-06-320, Swedish Nuclear

Fuel and Waste Management, Stockholm

(2006).

[7.60] WOOD, M.D., BERESFORD, N.A., SEMENOV, D.V., YANKOVICH, T.L.,

COPPLESTONE, D., Radionuclide transfer to reptiles, Radiat. Environ. Biophys.

(2010)

509–530.

[7.61] SHAHUL HAMEED, P., ASOKAN, R., IYENGAR, M.A.R., KANNAN, V., The

freshwater mussel Parreysia favidens (Benson) as a biological indicator of

polonium-210 in a riverine system, Chem. Ecol.

8 (1993) 11–18.

[7.62] CANADA NORTH ENVIRONMENTAL SERVICES, Shea Creek Project Area,

Environmental Baseline Investigation 2007–2009, Draft Report (2010).

[7.63] JOHNSTON, A., Radiation Exposure of Members of the Public Resulting from

Pperation of the Ranger Uranium Mine, Technical Memorandum

20, Supervising

Scientist for the Alligator Rivers Region, Australian Government Publishing Services,

Canberra

(1987).

[7.64] 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).

[7.65] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer in Terrestrial and Freshwater Environments,

Technical Reports Series No.

472, IAEA, Vienna (2010).

[7.66] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer to Wildlife, Technical Reports Series

No. 479, IAEA, Vienna (2014).

137

Chapter 8

210

Pb AND

210

Po IN MARINE SYSTEMS

8.1. INTRODUCTION TO THE MARINE ENVIRONMENT

Two thirds of the planet is covered with sea water. Near the continents

(neritic province), waters located between the shoreline and the shelf break, or

continental margin, are called seas and are fringes of the immense blue waters

beyond continental margins that make up the oceans (oceanic province). Oceans

spread over all latitudes and longitudes, and their depths range from the shoreline

to about 11

000 m in a few hadal trenches. The average depth of the North

Atlantic, for example, is about 4500

Despite the common chemical composition of sea water, oceans are by no

means a uniform system. The combination of water temperature, salinity, sunlight

and dissolved nutrients, such as chloride, sulphate, nitrate and other salts, creates

different biotopes that are home to many species in marine ecosystems.

Biota species and biomass are unevenly distributed in the oceans. On the

sea-floor, the biomass of benthic organisms decreases from the coastal waters to

abyssal and hadal regions. In the pelagic domain, biomass over the continental

shelf is much higher than in the open ocean. In specific areas of the continental

margin, the upwelling of nutrient loaded deep water enhances phytoplankton

primary productivity and contributes to the abundance of fishing resources.

8.2. POLONIUM SOURCES IN THE MARINE ENVIRONMENT

In sea water, polonium is present as a naturally occurring radionuclide and

it is a companion of many other essential and non-essential dissolved elements.

Polonium enters the marine environment from atmospheric deposition of

210

and

210

Po at the ocean surface, from the in situ radioactive decay of

226

dissolved in sea water, from the decay of

222

Rn gas exhaled from the sea-floor,

and from river and anthropogenic discharges. Dissolved polonium ions rapidly

adsorb onto suspended particles and are accumulated by marine organisms. The

downward movement of those particles and planktonic faecal pellets leads to

the depletion of polonium in surface waters, particularly in areas of enhanced

primary and secondary productivity.

The atmospheric depositions of

210

Po originates from the decay of

atmospheric radon, exhaled predominantly from landmasses and from

138

CHAPTER 8

resuspended dust. Most of the

210

Po measurements in surface air and in

atmospheric depositions were made simultaneously for

210

Pb and

210

Po, and

conclude that

210

Po concentrations in aerosols are minor and usually about

0.1

times

210

Pb concentrations (see Chapter 5). Average concentrations of

210

and

210

Po in the air, as well as atmospheric deposition fluxes, vary with latitude

and rainfall and are largely controlled by radon, exhalation and atmospheric

concentrations, which are dependent on land mass distributions, and are

modulated by global atmospheric circulation (see Chapter

5). Although direct

measurements of

210

Pb and

210

Po deposited at the ocean surface have not been

made,

210

Pb transport by global atmospheric circulation has been modelled, and

210

Pb deposition at the ocean surface has been estimated for specific areas to a

reasonable degree of accuracy

[8.1–8.3].

Ingrowth of

210

Po from

210

Pb–

210

Bi radioactive decay in sea water takes

place everywhere in the ocean and depends on

222

Rn concentration in sea

water, which in turn is the sum of

222

Rn from in situ decay of dissolved

226

and

222

Rn diffusing from the sea-floor following

226

Ra decay in sediments. The

resulting

210

Po concentrations in sea water also greatly depend on

210

Po binding

to suspended particulate matter and particle scavenging in the water column.

Polonium-210 also enters coastal areas with river discharges which feed the

seas with chemical elements, including radionuclides of the natural radioactive

series, in dissolved and solid phases. Owing to the low water solubility of

polonium and its high partitioning into the solid phase (K

), most polonium

discharged by rivers is associated with suspended particulate matter and bottom

sediment discharges

[8.4].

The most relevant, direct

210

Pb and

210

Po anthropogenic discharges into the

sea relate to non-nuclear industries, such as phosphate ore processing, oil and

gas exploitation, and heavy mineral sands mining

[8.5]. The better investigated

case studies relate to phosphate ore processing for phosphoric acid and phosphate

fertilizer manufacturing. Research conducted in several countries has shown

that naturally occurring concentrations of

210

Po in estuaries and coastal waters

receiving such phosphatic discharges were locally enhanced by factors in the

range of 10–1000

[8.2, 8.6–8.10]. Often neglected but locally important, past

radioactive waste dumping in the sea contained non-negligible amounts of

210

and

226

Ra [8.11, 8.12].

8.3. POLONIUM IN THE OCEANS

Polonium concentrations in dissolved and particulate phases have been

reported for coastal waters and for the open ocean (see Refs [8.13–8.15]).

Generally, polonium in coastal waters is largely associated with suspended

139

MARINE SYSTEMS

particulate matter, with a minor fraction in the dissolved phase (<35%), probably

due to the higher suspended particle loads and more intensive mixing than in open

ocean waters, where polonium is mostly in the dissolved phase (see Table

8.1). In

coastal sea water,

210

Pb is also mainly associated with suspended matter, although

to a lesser extent than

210

Po. Neither concentration shows significant seasonal

variation; thus, atmospheric depositions do not seem substantially to modify their

concentrations in sea water throughout the year.

In coastal waters, concentrations of dissolved

210

Po are around

0.5

Bq/m

, and usually the Po:Pb ratio is near one in the dissolved phase. In

suspended particulate matter, however,

210

Po concentrations are generally higher

than

210

Pb. Partitioning coefficients (K

) for

210

Pb and

210

Po between suspended

particulate matter and water (soluble fraction) have been reported for a few

environments, and indicate high K

values for

210

Po in the range of (2–7) × 10

which are on average five

times higher than the K

values for

210

Pb. The strong

partitioning of these two radionuclides into the particulate phase, higher for

TABLE 8.1. PARTITIONING COEFFICIENT (K

) OF

210

Pb AND

210

Po IN THE

MARINE ENVIRONMENT BASED ON IN SITU DETERMINATIONS

Region

Dissolved

(Bq/m

)

Particulate

(Bq/m

)

dissolved

Ref.

210

NW Mediterranean

Sea, off Spain

1.27 ± 0.31 0.049 ± 0.011 96 (1.0 ± 0.57) × 10

[8.16]

NW Mediterranean

Sea, off Monaco

1.62 ± 0.07 0.27 ± 0.02 86 (6.4 ± 0.5) × 10

[8.17]

NE Atlantic Ocean,

Portuguese surf

zone

0.54 ± 0.28 1.11 ± 0.73 33 (1.7 ± 1.0) × 10

[8.15]

210

NW Mediterranean

Sea, off Spain

1.09 ± 0.37 0.27 ± 0.083 79 (7.8 ± 5.6) × 10

[8.16]

NW Mediterranean

Sea, off Monaco

0.62 ± 0.03 0.25 ± 0.03 71 (1.6 ± 0.2) × 10

[8.17]

NE Atlantic Ocean,

Portuguese surf

zone

0.57 ± 0.13 1.87 ± 0.73 23 (2.8 ± 0.9) × 10

[8.15]

NE Atlantic Ocean,

French surf zone

0.47 ± 0.21 0.94 ± 0.37 33 (1.2 ± 0.62) × 10

[8.13]

140

CHAPTER 8

210

Po than

210

Pb, contributes to shaping the distribution of

210

Po in the marine

environment.

In the open ocean, the water column is stratified and vertical mixing is

reduced and mostly occurs in the epipelagic layer. Water stratification allows for

the formation of gradients of dissolved salts and relatively stable vertical profiles

of their concentrations. In the oceanic water column,

210

Po is produced from

radon decay in the atmosphere, the decay of dissolved radon and radon exhaled

from the ocean floor. Radon concentrations in the water column are lowest in the

surface layer and greatest in the deep-sea water layer. Radon distribution mainly

depends on diffusion in the water column and marine currents.

The isotopes

210

Pb and

210

Po are also introduced to the ocean from

atmospheric depositions, and can be detected in the upper layer of the ocean,

where a measurable excess of

210

Pb and

210

Po over dissolved

226

Ra concentrations

are reported

[8.16, 8.18]. Adsorption of

210

Pb and

210

Po onto particulates,

followed by uptake by phytoplankton and zooplankton, allows for the removal

of these radionuclides from the epipelagic layer with the downward flux of

biogenic particles. Some of the particles falling from the upper oceanic layer are

consumed in intermediate ocean layers and the chemical elements are recycled,

contributing to supporting the abundant marine life in the oceanic mesopelagic

layer. Particles, in their settling path through the deeper ocean layers, may further

adsorb and remove soluble

210

Pb and

210

Po from the water column, creating an

imbalance between

226

Ra (not adsorbed onto particles) and its particle reactive

daughters, and thus prevent the formation of secular radioactive equilibrium.

As a consequence of this radionuclide scavenging in the deep-sea layers,

210

and

210

Po are deficient in comparison to

226

Ra activity concentrations in all

oceans

[8.19–8.22]. Although very fine particles (Stokes particles, such as clay

particles) may settle with very low velocities of about 900

m/a, some of the

biogenic particulate materials generated in the upper layer of the ocean, especially

larger faecal pellets and carcasses of zooplankton organisms, sink rapidly and

may reach the abyssal sea-floor in days or weeks, adding

210

Pb and

210

Po to the

top sediment layer (see Fig.

8.1) [8.16, 8.23–8.25].

Figure 8.2 displays the vertical profiles of

210

Po along with those of

210

and

226

Ra in the north-eastern Atlantic Ocean water column. Similar profiles have

also been reported for other oceans

[8.17, 8.26, 8.27]. While the concentration

of dissolved

210

Pb can reach around 2.5 Bq/m

in the upper layer of the ocean,

dissolved

210

Pb in deeper layers barely reaches 1 Bq/m

. Dissolved

210

Po in

the upper layer is around 1

Bq/m

; it increases in the mesopelagic layer and

decreases again, and remains at concentrations much lower than those of

210

Pb all

the way down to the abyssal sea-floor. In particulate matter, the

210

Po:

210

Pb ratio

is generally consistently greater than 1.

141

MARINE SYSTEMS

The mean residence time of

210

Po in ocean water layers was calculated for

several oceans and values are about 0.5–1

years in the upper layer and somewhat

longer, around 2 years, in the soluble phase of deep-sea layers

[8.16–8.18,

8.23,

8.25]. Accumulation of

210

Po deposited onto the oceanic bottom sediments

has generally not been measured, as all available reports focus on

210

measurements

[8.15]. The

210

Po flux arriving at the abyssal sea-floor in the

north-eastern Atlantic Ocean was estimated as twice that of the

210

Pb flux and,

in turn, the settling

210

Pb flux at the abyssal sea-floor was computed as twice that

of the

210

Pb atmospheric flux entering at the surface of the ocean. Furthermore,

Source: Figure 2 of Ref. [8.23].

Note: d — dissolved phase; Fd — deposition flux on the sea-floor; Ia — atmospheric input

210

Pb (Bq·m

−2

·a

−1

); T — mean residence time (years); p — particulate phase;

Part. — radionuclide flux associated to the particle flux (Bq·m

−2

·a

−1

FIG. 8.1. Box model of the water column in the north-eastern Atlantic Ocean, with inventories

and mean residence times of

210

Pb and

210

Po, and deposition fluxes at the sea-floor.

142

CHAPTER 8

while the

210

Po:

210

Pb ratio in atmospheric particle depositions at the surface of

the ocean is around 0.1, in the sediment depositions onto the abyssal sea-floor of

the north-eastern Atlantic Ocean, the

210

Po:

210

Pb ratio is around 2 (see Fig. 8.1).

8.4. ACCUMULATION AND TURNOVER OF

210

Pb AND

210

IN MARINE ORGANISMS

Carvalho

[8.28] reports that:

“Polonium (

210

Po) and radioactive lead (

210

Pb) in marine organisms

have attracted the attention of scientists because of their relatively high

concentrations in comparison with those in terrestrial organisms and also

because

210

Po concentrations in marine biota often are much enhanced

in comparison to those of

210

Pb, its radioactive grandparent [8.29, 8.30].

The first determinations of

210

Po in marine biota had shown high

concentrations of largely unsupported

210

Po [i.e. above the

210

concentration] in plankton, 110

Bq·kg

−1

(dry weight), and soon other

reports added information on unsupported

210

Po in marine fish, crustacean

and whales

[Refs [8.29, 8.31–8.33]]. Early measurements did record

also high

210

Po concentrations in some marine species, such as the

lanternfish, and in certain internal organs such as the pyloric caecca of

tuna fish

[Refs [8.34, 8.35]]. Pioneer data were reviewed by Cherry and

Shannon

[Ref. [8.29]] who highlighted the importance of

210

Po as a major

Source: Based on Ref. [8.23].

FIG. 8.2. Profiles of

226

Ra,

210

Pb and

210

Po in the water column of the north-eastern Atlantic

Ocean.

143

MARINE SYSTEMS

internal source of the radiation dose received in the tissues of marine

organisms.

“Further work revealed the distribution of

210

Po in crustacean tissues,

including the euphausiid Meganicthyphanes norvegica, and the

role of zooplankton in the removal of

210

Po from the ocean surface

layer

[Refs [8.36–8.38]].”

Early work on polonium distribution in marine crustaceans highlighted

elevated concentrations in the hepatopancreas and suggested that food could be

the main source of polonium in invertebrates and the release of faecal pellets its

main excretion route

[8.36, 8.39–8.43].

The understanding of accumulation mechanisms and the biokinetics of

many radionuclides in marine biota has been the goal of extensive experimental

work carried out over decades in many radioecology laboratories. However, in

contrast to, for example, fission products and transuranium elements, not much

experimental work has been performed on the biokinetics of polonium in marine

species and aquatic species in general (see the early works in Refs

[8.44, 8.45]).

This is understandable because, as a pure alpha emitter, polonium is not

easily amenable to measurement by direct spectrometric techniques and even

its measurement by total alpha counting is not possible without some type of

relatively expensive and time consuming method. This difficulty was likely

the main reason for the long persistent gap in knowledge on polonium uptake

pathways, metabolism and turnover rates in marine biota, leaving polonium

biokinetics open to guess work or assumptions by inference from indirect

evidence.

The development of alpha spectrometry equipment and the production of

artificial polonium isotopes allowed the use of double tracer techniques, involving

three polonium isotopes, to investigate the role of the food ingestion pathway

(food labelled with

208

Po) and sea water (labelled with naturally occurring

and added

210

Po), and, using

209

Po as an isotopic tracer, for determination of

radiochemical yield and internal analytical quality control of every sample

[8.46].

Using such a double tracer technique, it is possible to quantify food and

water as polonium sources, as well as polonium elimination by excretion and

radioactive decay. This allows a mathematical description of the kinetics

of polonium uptake and elimination in marine organisms. Carvalho and

Fowler [8.46] model the polonium biokinetics using the following mass balance

equation:

( )

ww f

I WC AFC k Q

l= + -+

(8.1)

144

CHAPTER 8

where

is the activity of polonium in the organism at time t (Bq);

is the uptake rate constant fom water (mL/g organism d

−1

) by all processes,

namely drinking and surface adsorption;

is the weight of the organism (g);

is the polonium concentration in water (Bq/mL);

A is the digestive assimilation efficiency or absorption efficiency of polonium

from food;

is the amount of food ingested by an organism per day (g/d);

is the concentration of polonium in food (mBq/g);

K is the elimination rate constant of polonium (d

−1

) from the organism;

and λ is the radioactive disintegration rate constant of

210

Po (λ

210Po

= 0.005 01 d

−1

);

208Po

= 0.000 656 d

–1

). Each term in this equation can be considered separately

for assessing and describing the kinetics of polonium accumulation in the

organism from water and food only.

Expressed in terms of activity concentration, the polonium concentration in

an organism in steady state equilibrium of exchange with the environment (C

) is

described by

[8.46]:

( )

kWl

æö

èø

(8.2)

In a set of laboratory experiments using shrimp, prawn and fish, Carvalho

and Fowler [8.47] show that polonium absorbed in the internal tissues of biota

is almost exclusively taken up from the ingested food and absorbed through the

digestive pathway.

In shrimp, it was experimentally verified that, in spite of the presence

of dissolved

210

Po in sea water, there was only minor absorption of

210

Po ions

from water. Furthermore, this minor absorption did not take place through gill

epithelium, as it does for many other ionic elements, but rather the tiny amounts

210

Po absorbed from sea water were due to the drinking water reflex of marine

organisms for osmotic balance, and so this polonium was absorbed through the

gut wall in the same manner as polonium from food

[8.46, 8.47].

Furthermore, it was demonstrated that crustaceans, such as shrimp and

prawn, in sea water containing dissolved

210

Pb and

210

Po, easily accumulated

210

Po from ingested food into their internal organs, while

210

Pb absorption from

food was minor. In addition, dissolved

210

Pb in sea water was adsorbed onto

exposed surfaces, such as the exoskeleton and gills, without significant absorption

into internal organs. This differentiated uptake, facilitating gut absorption of

145

MARINE SYSTEMS

food-borne

210

Po in comparison to food-borne

210

Pb, explains the increased Po:Pb

ratio (>1) systematically found in internal tissues of marine biota.

Moreover, Carvalho and Fowler [8.46] show that the ingestion pathway

accounted for over 97% of

210

Po intake in shrimp, and for over 99% in fish

internal tissues. The absorption of polonium from

208

Po labelled food closely

matches the gut absorption efficiency of protein from diet in prawn and fish.

Once absorbed, this

208

Po is rapidly distributed in internal tissues according to

the distribution pattern of naturally occurring

210

Po. Further experiments show

that polonium absorbed though the digestive pathway and accumulated in the

liver of teleost fish binds to proteins, such as ferritin and metallothioneins, that

strongly bind

210

Po and are able to retain significant amounts of this radioelement

in internal tissues

[8.48].

Experiments carried out with phytoplankton cells verify their rapid uptake

210

Po from liquid culture media. Using these polonium labelled cells to feed

mussels, it can be shown that

210

Po is rapidly absorbed and accumulates in the

soft tissue of mussels. It can be concluded that filter feeding molluscs, such as

mussels, in spite of filtering large volumes of sea water containing dissolved

210

Po, accumulate

210

Po in internal tissues mainly from ingested food [8.49].

Experiments using the same polonium double tracer technique also show

that

210

Po dissolved in the liquid media of microalgal cultures can be rapidly

adsorbed onto and absorbed into phytoplankton cells. In addition, through

ingestion of these cells,

210

Po was transferred to herbivorous (planktivorous)

zooplankton

[8.50–8.52].

From kinetic studies on

210

Pb and

210

Po in marine shrimp, Carvalho and

Fowler [8.47] conclude that

210

Po has a biological half-life of 10–11 days

in whole body shrimp (T

1/2

= 7 days in the hepatopancreas; T

1/2

= 28 days in

muscle). At 4–11 days,

210

Pb has a shorter half-life in whole body shrimp and

a much shorter half-life in internal organs (T

1/2

= 4 days in the hepatopancreas;

1/2

= 2 days in muscle), which may further contribute to higher

210

Po:

210

Pb ratios

in marine biota.

The distribution of polonium in internal organs and tissues has been reported

for many marine organisms, including molluscs, crustaceans, fish and mammals.

Higher

210

Po concentrations are systematically found in the liver and digestive

tract, and lower ones in muscle tissue. For example,

210

Po concentrations in fish

tissues vary enormously from species to species (see Fig. 8.3), but there is a

general pattern of distribution: lower concentrations are consistently measured in

muscle tissue and higher concentrations are found in the gut and liver, mostly in

relation to the digestive organs (see Section 8.5.2).

This distribution pattern is common to teleost (bony fish) and

elasmobranchs (cartilaginous fish), and cartilaginous fish generally display

lower

210

Po concentrations than teleost fish — most probably on account of

146

CHAPTER 8

Source: See Ref. [8.53].

FIG. 8.3. Polonium-210 in fish organs of a teleost fish (above) and a cartilagineous fish

(below).

their physiology and the biochemical composition of their tissues [8.54, 8.55].

In several species of teleost fish and marine mammals,

210

Po variations in the

muscle tissue were correlated with red muscle content

[8.56]. As red muscle has

a higher myoglobin content, this association with

210

Po was considered a likely

biochemical association.

High

210

Po concentrations were reported in 1982 for the hepatopancreas

of some mid-water pelagic shrimp, giving rise to high absorbed radiation

doses

[8.57]. Surprisingly high

210

Po concentrations were also measured in

147

MARINE SYSTEMS

the gonads of several fish species [8.54]. This was assumed to relate to the

accumulation of proteins in the egg, but high

210

Po in eggs may expose the

genetic material to relatively high radiation doses. For example, in the case

of the common sardine, Carvalho [8.54] estimate that effective absorbed dose

from

210

Po could reach 170 mSv/a in the gonads and even higher in some other

tissues. It has been suggested that this could merit further investigation, as this

dose might have biological effects at a genetic level and might play a role in the

genetic variability of organisms.

The search for genotoxic effects (micronuclei and DNA strand breaks)

induced by

210

Po was undertaken in mussels from the Brazilian coast, with

naturally occurring

210

Po concentrations averaging 155.6 ± 7.9 Bq/kg (FW) in

their soft tissue. At this level of

210

Po concentrations, the internal dose rate was

0.02

mGy/d (7.3 mGy/a) and no genotoxic effects were observed. However, this

dose rate was lower than the suggested dose threshold for inducing effects —

mGy/d [8.58]. High

210

Po concentrations are also reported in marine mammals

at levels that are unusual by human standards (see Appendix

III).

8.5. POLONIUM IN MARINE BIOTA IN SEVERAL OCEAN REGIONS

The

210

Pb and

210

Po activity concentrations reported in marine biota are

spread over several orders of magnitude. The radionuclide levels do not relate

to seawater depths, ecosystems or geographical locations

[8.28]. This section

reviews

210

Po data pertinent to several ocean regions and many taxonomic

groups, with comments on ecology and trophic relationships in order to aid data

interpretation.

8.5.1. Intertidal zone

Carvalho

[8.28] reports that:

“The intertidal rocky shore and also soft bottom areas (e.g., in estuaries)

of the Atlantic Ocean and of the Mediterranean Sea are zones of dense

biological habitats with high biomass and typical populations such as

those of seaweeds, mussel and oyster beds, and characteristic fish species

(e.g., Blenniidae). Similar dense populations may be found on the rocky

shores of most temperate regions [Refs [8.59–8.61]].”

In the intertidal zone,

210

Po accumulated in macrophytic algae comes

from

210

Po dissolved in sea water. The first likely uptake mechanism is surface

adsorption by simple contact. Adsorption of dissolved radionuclides onto the

148

CHAPTER 8

surface of algal blades by diffusion mechanisms across the blade surface boundary

layer also apply to transuranic elements, such as plutonium and americium, with

similar ‘piston velocities’, and it is not considered an active biological uptake

mechanism (for further information, see the early work in Refs [8.62–8.64]).

Carvalho [8.28] reports that the

210

Po:

210

Pb ratios in macrophytic algae are

consistently higher than unity, often ranging from three to ten. This seems to be

due to preferential adsorption of

210

Po onto the algal surface. Analyses of

210

and

210

Po distribution in seaweed in the 1980s confirmed that most

210

Po is bound

or adsorbed onto external surfaces

[8.65]. Later work also confirmed passive

adsorption of metals to microalgal cell walls and fast kinetics of

210

Po and

241

sorption in marine phytoplankton and zooplankton cells

[8.66–8.68].

Detailed studies report

210

Po concentrations in seaweed averaging

4.8

Bq/kg (FW) (range: 1.6–9.1 Bq/kg) in the coastal area of Portugal in the

north-eastern Atlantic Ocean

[8.28, 8.54]. Carvalho [8.28] reports that differences

among algae classes are small, and Fucus sp. (Phaeophyceae), often used as

a bioindicator species, is the seaweed with the highest

210

Po concentration

(9.1

Bq/kg, FW). Concentrations of

210

Po in Fucus vesiculosus at Pirou, the

French Channel, are 13.7

Bq/kg (FW) [8.13]. Concentrations in the range

of 11–26

Bq/kg (FW) are reported for several seaweed species on the Tamil

Nadu coast, India, in a tropical environment

[8.69]. Since

210

Po accumulation

in seaweed is operated by sorption mechanisms, those slight variations in

210

concentrations in seaweed might be due to the sorption characteristics of algal

coatings. Furthermore, adsorption of radionuclides might also be affected

by water temperature, water mixing and suspended particulate matter load,

especially in the surf zone.

Polonium-210 accumulation in animal species is very different to that in

seaweed. Different mollusc species living in the same intertidal rocky shore can

display very different

210

Po concentrations. For example, Carvalho [8.28] finds

the following concentrations in molluscs at the north-eastern Atlantic coast:

— Limpet (Patella aspera): A herbivore mollusc feeding on seaweed grasped

from the rock surface — 11.6

Bq/kg (FW);

— Mediterranean mussel (Mytilus galloprovincialis): A filter feeding mussel

whose diet is organic suspended particles — 132

Bq/kg (FW);

— Winkle (Littorina littorea): A carnivore gastropod — 283 Bq/kg (FW).

Although these molluscs all live in the same zone,

210

Po is concentrated

in their soft tissue at different levels, which reflects that these species occupy

distinct biotopes and trophic levels although they live together in the same

zone

[8.28]. Comparing and interpreting results for organisms from rocky shores

149

MARINE SYSTEMS

is generally straightforward. However, this might not be the case for results from

soft bottom organisms.

Intertidal, sediment dwelling bivalve molluscs which live in soft

bottoms at the temperate north-eastern Atlantic coast, such as the clams Tapes

decussatus and Ensis siliqua, exhibit dry weight concentrations of 760

Bq/kg and

225

Bq/kg, respectively. Another clam (Anadara granosa) from a similar

biotope in the tropical environment of Mumbai Harbour, India, exhibited

313

Bq/kg (DW) and another burrowing bivalve (Tonna dolium) on the Tamil

Nadu coast, India, exhibited 132

Bq/kg (FW) [8.69, 8.70]. Although living in

similar environments, these different species might have different feeding

strategies and diets, which would account for the differences. It should be

recognized that results of radionuclide analysis in sediment burrowing molluscs

may be affected by incidental parameters, such as the presence of sediment in the

gut that is sometimes difficult to remove. While some researchers remove the gut

sediment prior to analysis by dissection, washing or depuration, others do not.

Thus without a common methodology in sample preparation, comparing results

and interpretation can be problematic.

This difficulty is not present in every case, especially when studying

organisms that do not ingest bulk sediment to process organic matter as food.

Although all crab species inhabit sandy and muddy substrates, their feeding

strategies involve food selection instead of ingestion of bulk sediment

[8.71].

Extensive analysis of

210

Po in ten species of brachyuran crabs from the intertidal

zone of the Gulf of Mannar, on the Indian coast of the Bay of Bengal, show

that

210

Po concentration (FW) in crabs varies according to their food habits: it

is low in herbivores, 45.9

Bq/kg; and significantly higher in carnivores, up to

221.2

Bq/kg.

Heyraud et al. [8.72] examine whether mussel samples from different

latitudes might reflect the differences in atmospheric depositions of

210

Pb and

210

Po (see Refs [8.73–8.80] and Table 8.2.). As depositions in the southern

hemisphere are up to ten times lower than those in the northern hemisphere (see

Section

5.3), they hypothesize that

210

Pb and

210

Po in mussels might reflect that

difference. Their results show that mussels from European and South African

coasts display similar average radionuclide concentrations and no noticeable

latitude effect can be identified. Furthermore, mussels from the coasts of

Denmark and the United Kingdom with low seawater temperature and strong

seasonal temperature fluctuation, when analysed for

210

Pb and

210

Po, display an

average and range of

210

Po concentrations comparable to mussels from tropical

zones on the coast of Brazil and from the coast of the Mediterranean Sea.

Charmasson et al. [8.73] compare

210

Pb and

210

Po accumulated in filter

feeding mussels from intertidal areas of the Mediterranean coast and mussels

from deep-sea hydrothermal vents of the Menez Gwen field, in the Mid-Atlantic

150

CHAPTER 8

Ridge, the latter exposed to sulphide rich vent waters, and report similar

concentrations despite all the environmental differences. Results for coastal

clams and deep-sea clams from soft bottoms of the north-eastern Atlantic Ocean,

both sediment burrowing, suspension feeding bivalves, also reveal similar

210

and

210

Po concentrations [8.28]. These results suggest that, in spite of different

TABLE 8.2. ACTIVITY CONCENTRATIONS (AND RANGE) OF

210

Po AND

210

Pb IN MUSSEL (Mytilus spp.) SOFT TISSUES

Region DW:FW

Concentration (Bq/kg, DW)

Po:Pb

(range)

Ref.

210

Adriatic Sea, Italy 0.16 112 (50–142) —

—

[8.75]

Aegean Sea, Turkey 0.19 305 (52–1344) 28.0 (6–167) 12.1

(3.1–25.0)

[8.76]

English Channel

Pirou, French coast

0.17

296

12.2

25.4

[8.13]

Irish Sea, Ireland 0.12 222 (80–468) 11.9 (4–25) 15.9 [8.14]

Mediterranean Sea

Croatia

Toulon, France

—

0.17

54–460

203.1

—

41.3

—

8.6

[8.77]

[8.73]

NE Atlantic

Portugal open coast

(oligotrophic)

Portugal Tejo Estuary

(mesotrophic)

Ré Island, France

0.13

—

759 (460–1470 )

210

289–589

45 (23–96)

—

[8.80]

[8.74]

North Sea

United Kingdom

—

104–3124

2.8–284

—

[8.79]

South Atlantic

Brazil

Cape of Good Hope,

South Africa

—

1995

156

(54–460)

409 (170–1290)

—

5.7 (1.2–16.2)

—

86 (22–218)

[8.78]

[8.57]

[8.72]

—: data not available.

Fresh weight.

151

MARINE SYSTEMS

environmental conditions in distinct ecosystems, similar organisms with similar

feeding strategies and at the same trophic level in the food chain accumulate

210

Po up to comparable concentrations.

Coastal mussels, often of the family Mytillidae, are used worldwide

as sentinel organisms to monitor contamination in coastal areas following

the concept of the pilot International Mussel Watch programme of the United

Nations Educational, Scientific and Cultural Organization to report

210

Po in

mussels in many regions. In the Mediterranean Sea, for example,

210

Po has

been determined in mussels from the coast to assess the radiological risk from

radioactive atmospheric depositions (mainly

137

Cs) from the Chernobyl accident

compared to

210

Po. It was concluded that, in spite of widespread deposition of

137

Cs in the Mediterranean Sea, the radiation dose to mussels and to mussel

consumers would be a minor fraction of that from naturally occurring

210

Po. A

similar finding was deteremined in 1995 by the then IAEA Marine Environment

Laboratory, Morocco

[8.81]. However,

210

Po concentrations for mussels also

vary widely

[8.82].

Mussels have also been used as sentinel organisms to monitor the impact

of local discharges of radionuclides which include

210

Po, for example from

phosphate industries. In some coastal areas subject to these effluents,

210

enhancement can be observed in mussels and other marine molluscs, while in

other cases, surprisingly no

210

Po enhancement could be detected [8.8, 8.79].

The

210

Po concentrations in mussels in different regions around the world vary

across a wide range of values, and results from specific areas with anthropogenic

discharges often cannot be considered significantly different on a statistical basis

(see Table

8.2). An examination of published regional

210

Po datasets reveals

wide ranges of concentration in mussels from the same coastal region, collected

at the same time of the year and by the same research team. For example, the

concentrations of naturally occurring

210

Po in mussels from 27 locations around

Ireland were in the range of 80–468

Bq/kg (DW) [8.14], which indicates that

wide variations truly do exist among samples and are not artefacts. An interesting

finding is that

210

Po concentrations in mussels from the most pristine areas are

sometimes even higher than those in mussels from areas subject to anthropogenic

discharges of

210

Po, thus rendering coastal monitoring results difficult to interpret

(see Refs [8.13, 8.14, 8.83]).

This interest in the use of mussels as biomonitors led to a detailed

investigation of

210

Pb and

210

Po bioaccumulation in mussels, the effects of

environmental parameters and the physiological condition index of mussels on

210

Po concentration [8.84], and the assessment of the allometric effect on

210

concentration

[8.85]. One important finding is that the fluctuation of

210

Po in

mussel soft tissue throughout the year seems at first glance to be a consequence of

the seasonal

210

Pb–Po atmospheric deposition flux. Instead, Carvalho et al. [8.85]

152

CHAPTER 8

Note: Based on the data presented in Ref. [8.85].

FIG. 8.4. Variation of

210

Pb and

210

Po body burden in mussel soft tissue as a function of shell

length.

verify that it is an effect of the seasonal fluctuation of the physiological condition

of mussels. They report that during winter, when mussels have less fat, the

210

concentration per kilogram in the soft tissue is higher than in summer. In summer,

when the accumulation of lipids and ripe gonads increase the physiological

condition index of mussels,

210

Po activity concentration expressed on a mass

basis exhibit lower values. However, when

210

Po concentration is expressed per

mussel (

210

Po body burden),

210

Po values are nearly stable over the year, and the

main seasonal fluctuation is of lipid (fat) reserves, not of

210

Po [8.15]. Carvalho et

al. [8.85] describe the allometric effect on

210

Po concentration as very pronounced

in mussels and, for example, an increase in mussel shell length from 2.5

cm to

5.0

cm can correspond to a decrease of 50% in

210

Po concentration in mussel soft

tissue (see Figs

8.4 and 8.5).

These findings on mussels, based on the determinations of

210

Pb and

210

Po, lead Carvalho et al. [8.85] to recommendations that can be applied to all

contaminants and include the following:

(a) To select carefully mussels of the same size class to avoid allometric effects

210

Po concentrations;

(b) To determine the physiological condition index of mussels to correct for the

fat content;

minimize inter-individual variation;

(d) To allow for statistically valid comparison of results from different places.

153

MARINE SYSTEMS

An improved sampling methodology can achieve comparable results in

coastal monitoring using molluscs.

8.5.2. Coastal seas

Carvalho

[8.28] reports that:

“Amongst the large diversity of fish species in the waters above continental

shelf, the Clupeidae (sardine, herring, and sprats), the Scombridae

(mackerel) and Thunidae (tuna), have a special importance due to their

abundance compared with other fish and to their major place in the human

diet. While these are all pelagic fishes, other also very common fish species

are demersal [i.e. live near the sea-floor and feed on benthic organisms],

such as the sea hake (Merlucius merlucius), the pouting (Trisopterus luscus)

and the small benthic sharks, such as the spotted catshark (Scyliorhinus

canicula).”

With regard to

210

Po transfer from the marine environment to human

populations, the consumption of seafood from coastal seas is the main pathway

and therefore has received much attention in the assessment of radiation exposure

(see Refs [8.70, 8.86–8.90]).

Fish species from many coastal regions around the world have been

analysed, either on a whole body basis or in segregated organs and tissues, and

the

210

Po concentrations are spread over a very wide range of values which has

challenged interpretation (see Table

8.3). Attempts have been made to find an

y = - 209.87x + 1589.8

= 0.386

= - 8.5992x + 70.912

= 0.335

100

1000

10000

1 234 5

Shell length (cm)

Radionuclide conc (Bq/kg)

Po210

Pb210

Note: Based on the data presented in Ref. [8.85].

FIG. 8.5. Allometry of

210

Pb and

210

Po concentrations in mussels as a function of shell length.

154

CHAPTER 8

TABLE 8.3. ACTIVITY CONCENTRATIONS OF

210

Po AND

210

Pb IN FISH

TISSUE FROM COASTAL AREAS OF THE NORTH ATLANTIC

Species and region Tissue DW:FW

210

(Bq/kg)

210

(Bq/kg)

Po:Pb

Sardine

(Sardina pilchardus)

Madeira Island, Portugal

Muscle 0.23 66 1.0 66

Liver 0.28 2 140 6.2 345

Gonad 0.27 275 1.8 156

Bone 0.39 197 26 7.5

Stomach 0.27 510 2.5 199

Caecca 0.23 2 490 138 18

Intestine 0.24 28 000 100 281

Chub mackerel

(Scomber japonicas)

Madeira Island, Portugal

Muscle 0.26 19 0.63 30

Liver 0.56 1 035 7.3 141

Gonad 0.25 183 22 8

Bone 0.39 42 11 3.7

Blue jack mackerel

(Trachurus picturatus)

Madeira Island, Portugal

Muscle 0.28 8.9 0.7 13

Liver 0.28 615 134 4.6

Gonad 0.17 56 1.1 51

Bone 0.38 43 15 2.8

White sea bream

(Diplodus sargus)

Lisbon, Portugal

Muscle 0.27 0.52 0.20 2.6

Liver 0.31 28 14.5 1.9

Gut 0.28

39 1.13 34

Caecca 0.25 69 5.0 14

Bone 0.53 25 31 0.8

Cod

(Gadus morhua)

Newfoundland, Canada

Muscle 0.19 0.628 0.040 16

Liver —

6.59 0.08 81

Gonad —

1.87 0.014 14

Bone —

0.41 0.61 0.7

European hake

(Merlucius merlucius)

Sesimbra, Portugal

Muscle 0.19 6.4 0.15 43

Liver 0.64 10.8 1.51 7

Gonad 0.17 52.0 1.26 41

Bone 0.30 7.8 0.77 10

Source: See Refs [8.28, 8.83].

—: data not available.

155

MARINE SYSTEMS

explanation for

210

Po variation based on allometry (body size and metabolic

rates of species), diet composition, water depth, latitude and other environmental

parameters.

For example, Dahlgaard [8.83] attempts to relate

210

Po concentration levels

in fish with variation of environmental parameters in the Baltic Sea. Average

210

Po concentrations (FW) in frozen fish fillets are 0.35 Bq/kg (cod), 0.65 Bq/kg

(herring) and 0.96 Bq/kg (plaice). The differences in

210

Po between species are

statistically significant. However, an attempt to find a correlation between

210

concentrations in fish and the salinity gradient, which in the Baltic–North Sea

region is in the range of 8–35%, shows no statistically significant effect either on

fish or on mussels.

A 1988 study of herring, cod and flounder in the Baltic Sea (Gulf of

Gdansk, Poland) found that the highest concentrations were measured in the

digestive organs, in particular the intestine

[8.91]. Furthermore, the contribution

of the digestive organs to the total

210

Po accumulated in the fish body was

positively correlated with the degree of fullness of the stomach, while the

210

concentration in fish decreased if food was lacking in the stomach.

The

210

Po concentration (

210

Po whole body burden or

210

Po concentration

in muscle) can allow for ranking species into different

210

Po concentration levels.

Heyraud and Cherry [8.40] tentatively interpret this as being dependent on

the diet of each species. For example,

210

Po concentrations are much lower in

herbivorous fish (e.g. white sea bream, Diplodus sargus) than in planktivorous

fish (e.g. sardine, Sardina pilchardus)

[8.28, 8.54]. Research was also published

in 1989 on the effect of diet composition on

210

Po concentrations in fish, in an

attempt to apply

210

Po as an indicator of diet, at least in clupeoid fish [8.92].

While some parameters play a role in constraining

210

Po concentrations, none

is sufficient to provide a key explanation for such a wide range of concentrations

determined in fish and other biota living in the same environment. A different

approach, rooted in previous findings, has been proposed for interpreting

210

Po concentrations in marine biota on the basis of food chains in the ocean

and the trophic position of organisms (see Refs [8.28, 8.46, 8.48, 8.54]). This

interpretation is based on evidence provided by a comparison of

210

Po levels in

species that are taxonomically related, often with very similar feeding strategies

and especially occupying the same trophic level in the food chain

[8.28].

One example is that the

210

Po concentration level in the muscle of

planktivorous sardines (Sardina pilchardus) from the north-eastern Atlantic

Ocean (see Table 8.3) is comparable to that in similar fish of the tropical sea

environment off the Tamil Nadu coast, India, where

210

Po concentrations (FW) in

planktivorous fish muscle are 190

Bq/kg for Sardinella longiceps and 116Bq/kg

for Eleuteronema tetradactylum

[8.28, 8.69]. Similarly high

210

Po concentrations

are reported for the anchovy (Engraulis encrasicolous) in the Adriatic Sea, and

156

CHAPTER 8

for the anchovy and the sprat (Sprattus sprattus) in the Black Sea — all of which

are planktivorous fish species that consistently exhibit

210

Po concentrations

higher than all other fish from the same areas

[8.75, 8.93].

Another example is the similar

210

Po concentration levels in mesopelagic

fish species, such as the lantern fish and hatchet fish, from the Pacific Ocean

and the North Atlantic Ocean, but with a similar diet based on small pelagic

zooplankton crustaceans and occupying the same trophic level (see Refs [8.28,

8.54, 8.94]).

A last example, reported by Carvalho

[8.28], is the

210

Po concentration in

large predators of coastal seas, such as dolphins (Delphinus delphis), with average

210

Po concentrations of 49 Bq/kg in the dorsal muscle and 107 Bq/kg in the liver,

which are generally higher than in fish

[8.95, 8.96]. The diet of these marine

mammals (Delphinidae) is exclusively based on seafood and they are at the top of

food chains in coastal seas. Dolphins concentrate

210

Po exclusively from the diet

(i.e.

210

Po transferred along the food chain) — there is practically no absorption

from water. Interestingly, other Delphinidae from the tropical waters of Brazil

in the southern hemisphere and from temperate waters off the Portuguese coast

exhibit very similar

210

Po concentrations, reinforcing the absence of an influence

from geographical and climatic regions and the key role of trophic level and diet

210

Po accumulation (see Appendix III).

Some trophic relationships and

210

Po concentration levels are easy to

interpret and correlate, such as

210

Po transfer from phytoplankton to herbivorous

zooplankton, from plankton to planktivorous fish, or from fish to marine

mammals, as described above; while others might be more challenging. For

example, Hassona et al. [8.97] report

210

Po concentrations (FW) for coral reef

fish species from the Red

Sea, a tropical environment with complex food webs in

the sublittoral zone, according to their general dietary habits:

— Carnivorous species: 1.8 Bq/kg (range: 0.25–6.42 Bq/kg);

— Herbivores: 2.2 Bq/kg (range: 1.52–3.80 Bq/kg);

— Omnivorous fish: 2.5 Bq/kg (range: 0.72–5.04 Bq/kg).

Although Hassona et al. [8.97] interpret

210

Po concentrations as a reflection

of the dietary habit of the species, differences in

210

Po concentrations among

groups are minor.

Eventually, and particularly in reef ecosystems, one aspect that might need

refinement and further detailed work is the understanding and classification

of the dietary habits of fish and a proper relationship with

210

Po accumulation.

There are at least four types of alimentary canal in marine herbivorous fish which

can be modelled using analogies with chemical reactors

[8.98]. Herbivores can

consume food of different qualities, with different digestion rates and different

157

MARINE SYSTEMS

transit times; and they can process food in non-acidic to highly acidic stomachs,

which can modify

210

Po uptake. Furthermore, carnivorous fish in coral reef

ecosystems involuntarily ingest large amounts of coral fragments while capturing

their prey. Studies have shown that sea-birds that find food in the sea exhibit

210

Po concentrations that reflect the food habits and

210

Po concentrations in their

prey

[8.99, 8.100].

8.5.3. Epipelagic zone

In the oceanic province,

210

Pb and

210

Po concentrations in the upper layer

of the ocean are controlled mainly by atmospheric deposition and, to a much

lesser extent, by the decay of

226

Ra dissolved in sea water [8.28]. A 1988 study

found the average

210

Po concentration in this oceanic upper layer to be around

1 mBq/L [8.22]. The upper layer of the ocean, or epipelagic zone, contains a

very significant biomass of phytoplankton and zooplankton

[8.59, 8.61].

Phytoplankton exhibits

210

Po concentration factors of around 10

–10

and Po:Pb

ratios of around 3, which are lower than those observed in some zooplankton

groups (see Refs [8.101–8.103]).

Investigations made of

210

Po accumulation in zooplankton report a high

concentration in euphausiids (krill) and that

210

Po concentration increases from

phytoplankton to macrozooplankton to mesozoopklankton

[8.39, 8.54, 8.102].

The concentrations of

210

Pb and

210

Po in zooplankton faecal pellets and their

role in the rapid removal of these radionuclides from the epipelagic zone have

triggered research on the vertical transport of elements carried by relatively large

biogenic particles in the water column (see Refs [8.24, 8.26, 8.27, 8.36–8.38]).

However, studies on plankton always face the difficulty that plankton samples

obtained with towed nets are based on the size of organisms, being a mixture

of species rather than plankton organisms sorted by phylogenetic group,

ecological niche or trophic level. Organism size is not an indication of trophic

level. Furthermore, even small planktonic organisms have diverse strategies of

food selection and may occupy different trophic levels regardless of their similar

size

[8.104].

There are data obtained by analysing plankton sorted by species or genus

rather than by analysing bulk plankton samples. Results on

210

Po based on species

separation has increased the understanding of

210

Po accumulation from sea water

to phytoplankton to mesozooplankton to krill (Eupasia superba) in the Antarctic

Ocean

[8.105]. Furthermore, a comparison of

210

Po datasets from Antarctica,

the Mediterranean and the north-eastern Atlantic Ocean reveals that the range

210

Po concentrations are similar and unrelated to seawater temperature or

geographical region (see Refs [8.54, 8.101, 8.102, 8.105]).

158

CHAPTER 8

Reporting on the north-eastern Atlantic Ocean, Carvalho [8.28] finds that:

“The groups of small zooplanktonic crustaceans such as ostracods,

amphipods, copepods, euphausiids and mysids, displayed the highest

210

activity concentrations in marine organisms ranging from 84

Bq·kg

−1

euphausiids [generally carnivores] to 2900

Bq·kg

−1

in amphipods [mostly

planktivorous herbivores]. Other important zooplanktonic groups, such as

the chaetognaths (Sagitta spp.), pteropods (Creseis spp., Diacria trispinosa)

and pelagic polychaetes contained

210

Po concentrations in that range and

Po:Pb ratios also higher than jellyfish, generally between 5 and 10 and

occasionally higher.

“Planktonic fish larvae (Mycthopidae, Alepochephalidae) and cephalopod

larvae (Loligonodae, Ommastrephidae) displayed

210

Po concentrations

similar to other zooplankton taxa as reported above, but with Po:Pb around

6 and not different from their adult fish stages while this ratio went up to

157 in crustaceans (mysids).”

Carvalho

[8.28] reports that large predator teleosts of the north-eastern

Atlantic Ocean epipelagic zone, such as tuna (Thynnus obesus and other tuna

species), feed upon clupeoid fish, squid and zooplankton, while the blue marlin

(Makaira nigricans) feeds upon clupeoid fish, and the oil-fish (Ruvettus pretiosus)

eats several pelagic fish species (see Fig.

8.6). Polonium-210 concentration in

tuna muscle is 3.0

± 0.1 Bq/kg, with a Po:Pb ratio of 6.6, in blue marlin muscle

0.4

± 0.2 Bq/kg and in oil-fish 0.7 ± 0.02 Bq/kg, thus also much lower than in

tuna.

Carvalho

[8.28] finds that:

“Other oceanic large predators on the top of food chains, such as the sperm

whale (Physeter catodon) feeding upon deep water cephalopods, contained

210

Po and

210

Pb concentrations in muscle tissue roughly comparable to

tuna.”

One interesting question is raised by

210

Po concentrations in the same or

closely related species, occupying the same trophic level in different oceans.

For example,

210

Po concentrations in tuna muscle from the Atlantic Ocean, the

Pacific Ocean and the Mediterranean Sea are within one order of magnitude.

However, taking into account the uncertainty of determinations, they are

significantly different and are in the range of 3–53.3

Bq/kg (FW) (see Table 8.4).

As tuna are at the top of marine food chains and do not absorb

210

Po from water,

different

210

Po concentrations in tuna must be a consequence of different

210

159

MARINE SYSTEMS

Source: See Fig. 2 of Ref. [8.28].

FIG. 8.6. Polonium-210 in pelagic food chains in the ocean (reproduced with permission

courtesy of Elsevier).

TABLE 8.4. ACTIVITY CONCENTRATIONS (AND RANGES) OF

210

Po IN

TISSUES OF TUNA FISH FROM SEVERAL OCEANS

Region Species Tissue

210

Po concentration

(Bq/kg, FW)

Ref.

NE Atlantic Ocean

Azores Islands

Several

Muscle

Liver

Bone

5 (3–8)

288 (278–297)

20 (5–51)

[8.53]

Pacific Ocean,

S. California

Several

Muscle

Liver

Gonad

11–48

144–777

85–278

[8.34]

Japan

Thunnus thynnus

Muscle 24.4 ± 1.6 [8.85]

Marshall Islands Bonito Muscle 36.9 (21.5–53.3) [8.86]

160

CHAPTER 8

concentrations in the first trophic levels of the food chain (i.e. in phytoplankton

and zooplankton in their environment). The

210

Po analyses of plankton from the

open ocean, especially from oligotrophic waters in the Pacific Ocean, actually

show that

210

Po concentrations in zooplankton are very much affected by plankton

density (see Ref. [8.101] for a model to describe the relationship). Low biomass

prompts higher concentration levels in phytoplankton, which is transferred

to zooplankton, and so forth. Polonium-210 concentrations in zooplankton in

oligotrophic regions of the Pacific Ocean are indeed much higher than

210

concentrations in zooplankton of the North Atlantic Ocean. This is propagated in

the food chain and reflected by

210

Po concentrations in muscle tissue of tuna fish

of both oceans

[8.5, 8.87].

8.5.4. Mesopelagic and bathypelagic zones

Zooplankton from the deep waters of the north-eastern Atlantic Ocean

display very different concentrations according to the type of organism.

Carvalho

[8.28] reports that deep-sea medusa (Atolla sp. and Crossota sp.), salpes

(Salpa spp.) and Pyrosoma sp. — all gelatinous organisms — contain the lowest

210

Po concentrations measured in marine organisms at 0.7–35 Bq/kg (with Po:Pb

ratios generally <

5). Their

210

Po concentrations are similar to those determined

in epipelagic jellyfish (i.e. Hydromedusae Sarsia spp. and Obelia spp.).

Research on mesopelagic decapod shrimp shows the existence of

specialized diets, with the consumption of

210

Po rich organic particles in the

oceanic water column, leading in some species to very high concentrations of

210

Po. High

210

Po concentrations are reported, especially in the hepatopancreas

of some shrimp species

[8.106] and in the liver and gut of some mesopelagic fish

feeding on particles and zooplankton

[8.54, 8.94]. This accumulation leads to

unusually high internal radiation doses from the alpha emission of

210

Po, reaching

values of 3–5

Sv/a, reported in the 1980s, in the hepatopancreas of shrimp and

liver and other internal organs of small mesopelagic fish (Argyropelecus and

Myctophum)

[8.54, 8.57].

Carvalho

[8.28] analyses more than 50 fish species living in the mid-water

column and sampled near Madeira Island, at the Great Meteor East Seamountin of

the Madeira Abyssal Basin, and in the Porcupine Abyssal Basin. Generally, most

mesopelagic and bathypelagic fish species are small (0.5–60

g adult weight),

but there are large differences in their

210

Po concentrations, suggesting feeding

regimes based on different prey and different trophic levels (see Fig.

8.6).

Large predators in mesopelagic and bathypelagic domains include

teleosts, such as Aphanopus carbo and Mora mora, and black sharks, such as

Scynodesmus sp., all with low

210

Po concentrations, around 1 Bq/kg in muscle

161

MARINE SYSTEMS

tissue, and comparable to concentrations in other fish in other ocean layers

occupying similar trophic levels

[8.28].

8.5.5. Abyssal zone

The amount of data available on abyssal species is much less than for

other oceanic regions. Nevertheless, the available data show that

210

Pb and

210

concentrations determined in ascidians, polychaetes, Ophiuridae, cephalopods,

crustaceans and fish are similar to concentrations reported for organisms at other

oceanic depths (see Table

8.5).

Carvalho

[8.28] finds that:

“Also

210

Po concentrations measured in the giant abyssal mysid

Gnatophausia ingens and Eryonidae crabs, 18

± 2 and 22 ± 2 Bq kg

−1

respectively, were close to the

210

Po concentration measured in the coastal

shrimp Leander serratus, 25

± 0.8 Bq kg

−1

”.

In Carvalho

[8.28], a small bivalve mollusc from the abyssal plain

(Sillicula fragilis) displays

210

Po concentrations in soft tissue comparable

to the north-eastern Atlantic clam (Tapes decussatus) and the North Sea clam

(Mya arenaria

) (see also Ref. [8.84]). Unusually high

210

Po concentrations,

1.6

× 10

Bq/kg (DW) were reported in the 1980s for infaunal xenophytophores,

large protozoans that live in sediment and are closely related to foraminifera, from

a hadal trench in the Pacific Ocean, which might be exposed to radiation doses

comparable to the dose in the hepatopancreas of mid-water shrimp

[8.57, 8.107].

Carvalho

[8.28] concludes that:

“One singularity of abyssal fauna shall be highlighted. In 3 out of the

15 abyssal species analyzed the Po:Pb ratios were around the unity. In

the abyssal fish Nematonurus (Coryphaenoides) armatus, a rattail fish

characteristic of the abyssal zone, the

210

Po:

210

Pb ratio in muscle was 0.5,

which was uncommon amongst all fish analyzed

[8.84]. These low Po:Pb

ratio values are probably linked to the long fasting periods of abyssal

fauna

[Ref. [8.108]].”

162

CHAPTER 8

TABLE 8.5. ACTIVITY CONCENTRATIONS (AND RANGES) OF

210

Po AND

210

Pb (FW) IN BENTHIC FAUNA OF THE PORCUPINE ABYSSAL PLAIN,

THE NORTH-EASTERN ATLANTIC OCEAN AND HYDROTHERMAL

VENTS IN THE MID-ATLANTIC RIDGE

Species and tissue N DW:FW

210

(Bq/kg ± SD)

210

(Bq/kg ± SD)

Po:Pb

Ascidian

Chitonanthus abyssorum

4 0.15 38 ± 2 38 ± 2 1

Polychaetes 3

Gut 0.22 27 ± 2 11 ± 0.5 2.5

Remainder 0.18 52 ± 2 13 ± 0.5 4

Asteridae

Astropecten sp., wt = 0.6 g

3 0.38 50 ± 3 7.6 ± 0.4 6.4

Astropecten sp., wt = 13 g

1 0.23 171 ± 13 25 ± 2 6.7

Bivalve mollusc (soft tissue)

Silicula fragilis

1 0.16 124 ± 8 123 ± 4 1

Bathymodiolus azoricus

—

0.19

123.9 ± 21.4

—

Isopods 2

Whole body 0.23 27 ± 1

1.91 ± 0.09

Gut 0.41 94 ± 3 11.4 ± 0.3 8.2

Exoskeleton 0.68 194 ± 5 8.4

± 0.2 23

Amphipods

Eurythenes grillus

3 0.20 32 ± 10 28 ± 11 1.1

Eurythenes grillus

26 0.28 104.1 (62.4–285.9) 26.5 (11.8–67.4) 3.9

Cephalopods

Vampiroteuthis infernalis

1 0.07 39 ± 3 6.9 ± 0.3 5.6

Fish

Nematonurus (C.) armatus

Muscle 9 0.17 0.31 ± 0.28 0.60 ± 0.55 0.5

Liver 9 —

3.58 ± 2.29 2.50 ± 1.64 1.4

Skin 4 —

1.94 ± 0.73 0.68 ± 0.33 2.8

Bone 4 —

3.25 ± 1.96 0.85 ± 0.24 3.8

Source: See Ref. [8.83] unless otherwise indicated.

Note: SD — standard deviation.

Source is Ref. [8.73].

—: data not available.

163

MARINE SYSTEMS

8.6. POLONIUM AND LEAD CONCENTRATION RATIOS IN

MARINE ORGANISMS

Concentration ratios (CR) are defined as the ratio of activity concentration

in the organism (Bq/kg, FW) to the activity concentration in filtered sea water

(Bq/L). Determinations of CRs are frequently performed when comparing

bioaccumulation of radionuclides in different locations and seas and are used

in modelling. As this ratio uses the concentration of radionuclide dissolved

in sea water, it allows for an easy identification of organisms that concentrate

more of the radionuclide, which might be of interest for further study and use in

environmental radiological monitoring programmes as biomonitors. Organisms

that concentrate a particular radionuclide might also be the main transfer route of

that radionuclide to humans.

However, computation of CRs does not imply that sea water is the source

of radionuclide to the organism. Indeed, it is food and not sea water that is the

source of internally accumulated

210

Po. Nevertheless, the information on CRs

for biota groups is useful to model the transfer of radionuclides in food chains

and to predict potential radiation exposures in specified ecosystems and food

webs. This is the main purpose of the data compilation and recommended

CRs in IAEA publications such as the Handbook of Parameter Values for the

Prediction of Radionuclide Transfer to Wildlife

[8.109]. Furthermore, CRs of

210

Po concentrations in sea water are not different from those of similar species in

coastal areas (see Fig.

8.7).

8.7. TRANSFER OF

210

Pb AND

210

Po IN MARINE FOOD CHAINS

In the past, the interpretation that

210

Po concentration in tissues of marine

biota could be a reflection of

210

Po content in their diet was gaining credence, but

the relative importance of water and food as

210

Po sources remained unknown.

The use of radioactive tracers clarified that diet is the almost exclusive source of

210

Po accumulated in the internal organs of marine organisms [8.46, 8.47]. Since

then, several studies have adopted this point of view and interpreted

210

Po data in

marine biota in the light of food chains and trophic levels (see Refs [8.28, 8.52,

8.110, 8.111]).

A consistent view of

210

Po transfer in marine food chains in the pelagic

environment is depicted in Fig.

8.7. Whole body

210

Po CRs are depicted, relating

210

Po concentrations in biota to

210

Po concentrations in sea water. Furthermore,

210

Po transfer factors (TFs) relating

210

Po concentrations in predator tissues to

210

Po concentrations in the diet of each species are indicated.

164

CHAPTER 8

Although these are simplified trophic chains and crude representations

of real food webs in the oceans, they are based on the known diet of species

and predator–prey relationships

[8.28]. The advantage is that it clearly shows

that in the same oceanic ecosystem there are organisms often in the same

zoological group, such as fish, that display very different concentration factors

and thus highlight that water cannot be the main source of

210

Po to organisms.

Furthermore, organisms with similar

210

Po CRs can be found in different regions

and ecosystems without any relationship between them.

Carvalho [8.28] reports that

210

Po TFs from prey to predator are often

around 0.1, with interesting deviations such as 0.7 in the trophic link from

phytoplankton to copepods, 0.3 in the trophic link from small pelagic fish

and euphausiids to tuna, and 0.7 in the transfer of

210

Po from zooplankton to

the lantern fish (Myctophidae). Carvalho

[8.28] proposes that these

210

Po TFs

are similar to energy transfer in marine food chains, known as eco-trophic

coefficients, and that this strongly suggests an association between

210

Po transfer

and energy transfer in marine food chains. Furthermore [8.28]:

Source: Figure 1 of Ref. [8.84].

Note: Concentration ratios are equivalent to concentration factors (CFs).

FIG. 8.7. Transfer factors (TFs) and concentration ratios for

210

Po for the demersal food chain

at the abyssal north-eastern Atlantic Ocean (reproduced with permission courtesy of Oxford

Journals).

165

MARINE SYSTEMS

“Based on

210

Po concentration data in marine biota combined with

210

behaviour in biological systems, such as

210

Po binding to amino acids

and proteins

[Ref. [8.48]], and experimental research on

210

Po and

210

assimilation in marine species

[Refs [8.46, 8.47, 8.49]], we interpret

210

transfer in marine food chains as a tracer of protein transfer from prey to

predator tissues. In every ecosystem of the ocean, the accumulation of

210

Po in the marine biota thus reflects the trophic position of the species

in the food web rather than the geographic location, water depth, or other

environmental parameter.

“The behaviour of

210

Pb in the food chains seems slightly different.

210

concentrations do not increase from phytoplankton to copepods as much as

210

Po, and the

210

Pb transfer to planktivorous fish, such as sardines, is less

efficient than

210

Po transfer.”

At higher trophic levels, such as in marine mammals,

210

Po levels are high,

while

210

Pb levels are as low as in lower trophic levels. The Po:Pb ratio in marine

food webs seem to increase with trophic level, from around 10 in phytoplankton

and zooplankton, 3–10 in herbivorous fish and 50–100 in piscivores (carnivorous

fish), rising to 200 in the muscle tissue of marine mammals

[8.96].

In summary, the concentrations of dissolved

210

Pb and

210

Po in the oceans

vary within a relatively narrow range. Differences between regions and ocean

layers do exist, and there are geochemical and geographical reasons for this, such

210

Pb and

210

Po concentrations in surface water of the oceans and land mass

distribution. However, these differences are relatively small and are not sufficient

to explain the wide range of

210

Po concentrations in biota, especially when

species coexist in the same water body or region. Carvalho [8.28] finds that:

“Po-210 and

210

Pb concentrations were determined in marine species from

all the ecological zones in the ocean, from the shoreline to abyssal depths. In

the tissues of all species the

210

Po concentration values spread across 5 orders

of magnitude, from 0.5

Bq kg

−1

in jellyfish to (2.8 ±0.2) × 10

Bq kg

−1

in the gut wall of the common sardine [Sardina pilchardus] and even to

(3.33

± 0.17) × 10

Bq kg

−1

in the gut walls of the pelagic shrimp Plesionika

edwardsi.”

Furthermore,

210

Po concentrations seem to be similar in the tissues of a

coastal bivalve mollusc and in a deep-sea mollusc, and in the muscle tissue of

tuna fish in the Pacific Ocean and the Atlantic Ocean, provided that those species

occupy a similar position in the food chain. Conversely,

210

Po concentrations in

marine biota might be very different in species living in the same water mass, and

166

CHAPTER 8

therefore exposed to the same

210

Po concentrations in water, as reviewed in this

appendix for many species in marine ecosystems.

The concentration of dissolved

210

Po in suspended particulate matter,

including living microorganisms, such as bacteria and phytoplanton cells,

probably starts with surface adsorption on ultra fine particles and accumulation

of polonium in bacteria and phytoplankton as an element similar to sulphur in

the building blocks of proteins — the amino acids. According to Carvalho [8.28],

the enhancement of

210

Po concentrations is very pronounced in biota feeding

on bacteria and phytoplankton at the base of marine food chains, such as

observed in small zooplankton organisms (e.g. copepods and mysids), but also

in large planktivorous organisms, such as mussels, sardines and blue whales,

and it is transferred to carnivores, such as marlins, dolphins and sperm whales.

Carvalho [8.28] concludes that:

“Actually, the

210

Po activity concentration levels recorded in marine

organisms and

210

Po:

210

Pb ratios may depend upon the number of trophic

levels in the food chain. The less trophic links exist in the food chain, the

higher will be the

210

Po concentration in the top predator tissues.”

Furthermore, in oligotrophic regions of the oceans, more

210

Po is

concentrated per unit of mass of phytoplankton and zooplankton, and this

210

is transferred to upper trophic levels in the food chain, while in coastal seas with

high plankton biomass, the reverse is observed (i.e. plankton display lower

210

concentrations) [8.101]. In marine organisms [8.28]:

“

210

Po:

210

Pb ratios mostly ranged from near 1 up to 100, and occasionally

higher, showing the enhancement of

210

Po concentrations comparatively to

210

Pb in internal organs, especially in those related to digestive functions

and the gonad.

.......

“Regarding the

210

Po transfer in marine food chains,

210

Po concentrations

in trophic levels and in characteristic species highlighted the similarities of

210

Po transfer factors with energy transfer rates in marine food chains.”

This has been experimentally verified in a few species using a double tracer

technique

[8.46, 8.47]. Carvalho [8.28] suggests that:

167

MARINE SYSTEMS

“

210

Po follows protein transfer in food chains and may thus give a surrogate

measure of the fraction of prey tissues that is incorporated in predator’s

tissues in marine food webs.”

With regard to

210

Po transfer from the marine environment to human

populations, the consumption of seafood from coastal seas is the main pathway

and deserves much more attention in radiation exposure assessment. Humans,

as consumers of seafood, are also elements of sea-borne food chains and,

according to their dietary habits, will have different intakes of

210

Po and thus

different internal radiation doses. This was pointed out in early studies on

seafood consumers and documented in dietary studies for several countries [8.70,

8.86, 8.88]. This exposure of humans to

210

Po is not constant and, at present

and due to overfishing, humans are increasingly fishing down the marine food

chains, which implies that current trends of wild seafood consumption are likely

to increase

210

Po ingestion and the collective radiation dose in seafood prone

populations

[8.28].

REFERENCES TO CHAPTER 8

[8.1] TUREKIAN, K.K., NOZAKI, Y., BENNINGER, L.K., Geochemistry of atmospheric

radon and radon products, Annu. Rev. Earth Planet. Sci.

5 (1977) 227–255.

[8.2] CARVALHO, F.P., Origins and concentrations of

222

Rn,

210

Pb,

210

Bi and

210

Po in the

surface air at Lisbon, Portugal, at the Atlantic edge of the European continental

landmass, Atmos. Environ.

29 (1995) 1809–1819.

[8.3] HEINRICH, P., JAMELOT, A., Atmospheric transport simulation of

210

Pb and

Be by

the LMDz general circulation model and sensitivity to convection and scavenging

parameterization, Atmos. Res.

101 (2011) 54–66.

[8.4] CARVALHO, F.P., Distribution, cycling and mean residence time of

226

Ra,

210

Pb and

210

Po in the Tagus estuary, Sci. Total Environ. 196 (1997) 151–161.

[8.5] “Summary”, Naturally Occurring Radioactive Material (NORM V) (Proc. Int. Symp.

Seville, 2007), IAEA, Vienna (2008)

1–26.

[8.6] BOLIVAR, J.P., GARCIA-TENORIO, R., VACA, F., Radioecological study of an

estuarine system located in the south of Spain, Water Res.

34 (2000) 2941–2950.

[8.7] GERMAIN, P., LECLERC, G., SIMON, S., Distribution of

210

Po in Mytilus edulis

and Fucus vesiculosus along the Channel coast of France: Influence of industrial

releases in the Seine River and Estuary, Radiat. Prot. Dosim.

45 (1992) 257–260.

[8.8] CAMPLIN, W.C., BAXTER, A.J., ROUND, G.D., The radiological impact of

discharges of natural radionuclides from a phosphate plant in the United Kingdom,

Environ. Int.

22 (1996) 259–270.

[8.9] ROLLO, S.F.N., CAMPLIN, W.C., ALLINGTON, D.J., YOUNG, A.K., Natural

radionuclides in the UK marine environment, Radiat. Prot. Dosim.

45 (1992) 203–209.

168

CHAPTER 8

[8.10] EL SAMAD, O., AOUN, M., NSOULI, B., KHALAF, G., HAMZE, M., Investigation

of the radiological impact on the coastal environment surrounding a fertilizer plant,

J. Environ. Radioact. 133 (2014) 69–74.

[8.11] OECD NUCLEAR ENERGY AGENCY, Co-ordinated Research and Environmental

Surveillance Programme Related to Sea Disposal of Radioactive Waste, CRESP Final

Report

1981–1955, OECD, Paris (1996).

[8.12] INTERNATIONAL ATOMIC ENERGY AGENCY, Inventory of Radioactive Waste

Disposals at Sea, IAEA-TECDOC-1105, IAEA, Vienna

(1999).

[8.13] GERMAIN, P., LECLERC, G., SIMON, S., Transfer of polonium-210 into Mytilus

edulis (L.) and Fucus vesiculosus (L.) from the baie de Seine (Channel coast of

France), Sci. Total Environ.

164 (1995) 109–123.

[8.14] RYAN, T.P., DOWDALL, A.M., McGARRY, A.T., POLLARD, D., CUNNINGHAM,

J.D.,

210

Po in Mytilus edulis in the Irish marine environment, J. Environ. Radioact. 43

(1999)

325–342.

[8.15] CARVALHO, F.P., OLIVEIRA, J.M., SOARES, A.M.M, Sediment accumulation and

bioturbation rates in the deep Northeast Atlantic determined by radiometric

techniques, ICES J. Mar. Sci.

68 (2011) 427–435.

[8.16] MASQUÉ, P., SANCHEZ-CABEZA, J.A., BRUACH, J.M., PALACIOS, E.,

CANALS, M., Balance and residence times of

210

Pb and

210

Po in surface waters of

the northwestern Mediterranean Sea, Cont. Shelf Res.

22 (2002) 2127–2146.

[8.17] TATEDA, Y., CARVALHO, F.P., FOWLER, S.W., MIQUEL, J.-C., Fractionation of

210

Po and

210

Pb in coastal waters of the NW Mediterranean continental margin, Cont.

Shelf Res.

23 (2003) 295–316.

[8.18] BACON, M.P., BELASTOCK, R.A., TECOTZKY, M., TUREKIAN, K.K.,

SPENCER, D.W., Lead-210 and polonium-210 in ocean water profiles of the

continental shelf and slope south of New England, Cont. Shelf Res.

8 (1988) 841–853.

[8.19] SPENCER, D.W., BACON, M.P., BREWER, P.G., The Distribution of

210

Pb and

210

Po in the North Sea, Center for Marine Research (1980).

[8.20] CHUNG, Y.,

210

Pb in the western Indian Ocean: Distribution, disequilibrium, and

partitioning between dissolved and particulate phases, Earth Planet. Sci. Lett.

(1987)

28–40.

[8.21] CHUNG, Y., FINKEL, R.,

210

Po in the western Indian Ocean: Distributions,

disequilibria and partitioning between the dissolved and particulate phases, Earth

Planet. Sci. Lett.

88 (1988) 232–240.

[8.22] CHERRY, R.D., HEYRAUD, M., “Lead-210 and polonium-210 in the world’s

oceans”, Inventories of Selected Radionuclides in the Oceans, IAEA-TECDOC-481,

IAEA, Vienna (1988)

139–158.

[8.23] CARVALHO, F.P., “Uranium series radionuclides in the water column of the

northeast Atlantic Ocean”, Isotopes in Hydrology, Marine Ecosystems and Climate

Change Studies (Proc. Int. Symp. Monaco, 2011), Vol.

2, IAEA, Vienna

(2013)

427–441.

[8.24] FOWLER, S.W., KNAUER, G.A., Role of large particles in the transport of elements

and organic compounds through the oceanic water column, Prog. Oceanogr.

(1986)

147–194.

169

MARINE SYSTEMS

[8.25] STEWART, G.M., MORAN, S.B., LOMAS, M.W., Seasonal POC fluxes at BATS

estimated from

210

Po deficits, Deep Sea Res. Part I 57 (2010) 113–124.

[8.26] VERDENY, E., et al., POC export from ocean surface waters by means of

234

Th/

238

and

210

Po/

210

Pb disequilibria: A review of the use of two radiotracer pairs, Deep Sea

Res. Part

II 56 (2009) 1502–1518.

[8.27] LE MOIGNE, F.A.C., et al., Export of organic carbon and biominerals derived from

234

Th and

210

Po at the Porcupine Abyssal Plain, Deep Sea Res. Part I 72

(2013)

88–101.

[8.28] CARVALHO, F.P., Polonium (

210

Po) and lead (

210

Pb) in marine organisms and their

transfer in marine food chains, J. Environ. Radioact. 102 (2011) 462–472.

[8.29] CHERRY, R.D., SHANNON, L.V., The alpha radioactivity of marine organisms, At.

Energy Rev. 12 (1974) 3–45.

[8.30] PARFENOV, Yu.D., Polonium-210 in the environment and in the human organism,

At. Energy Rev. 12

(1974) 75–143.

[8.31] SHANNON, L.V., CHERRY, R.D., Polonium-210 in marine plankton, Nature 216

(1967) 352–353.

[8.32] HOLTZMAN, R.B., “Concentrations of the naturally occurring radionuclides

226

Ra,

210

Pb and

210

Po in aquatic fauna”, paper presented at 2nd Nat. Symp. Radioecology,

Ann Arbor, MI (1967).

[8.33] SCHELL, W.R., JOKELA, T., EAGLE, R., “Natural

210

Pb and

210

Po in a marine

environment”, Radioactive Contamination of the Marine Environment (Proc. Symp.

Seattle, 1972), IAEA, Vienna (1973) 701–724.

[8.34] FOLSOM, T.R., BEASLEY, T.M., “Contributions from the alpha emitter,

polonium-210, to the natural radiation environment of the marine organisms”, ibid.,

625–632.

[8.35] HOFFMAN, F.L., HODGE, V.F., FOLSOM, T.R.,

210

Po radioactivity in organs of

selected tunas and other marine fish, J. Radiat. Res.

15 (1974) 103–106.

[8.36] BEASLEY, T.M., HEYRAUD, M., HIGGO, J.J.W., CHERRY, R.D., FOWLER,

S.W.,

210

Po and

210

Pb in zooplankton fecal pellets, Mar. Biol. 44 (1978) 325–328.

[8.37] CHERRY, R.D., FOWLER, S.W., BEASLEY, T.M., HEYRAUD, M., Polonium-210:

Its vertical oceanic transport by zooplankton metabolic activity, Mar. Chem.

(1975)

105–110.

[8.38] JEFFREE, R.A., SZYMCZAK, R., Enhancing effect of marine oligotrophy on

environmental concentrations of particle-reactive trace elements, Environ. Sci.

Technol.

34 (2000) 1966–1969.

[8.39] HEYRAUD, M., FOWLER, S.W., BEASLEY, T.M., CHERRY, R.D., Polonium-210

in euphausiids: A detailed study, Mar. Biol.

34 (1976) 127–136.

[8.40] HEYRAUD, M., CHERRY, R.D., Polonium-210 and lead-210 in marine food chains,

Mar. Biol.

52 (1979) 227–236.

[8.41] CHERRY, R.D., HEYRAUD, M., HIGGO, J.J.W., Polonium-210: Its relative

enrichment in the hepatopancreas of marine invertebrates, Mar. Ecol. Prog. Ser.

(1983)

229–236.

170

CHAPTER 8

[8.42] HIGGO, J.J.W., CHERRY, R.D., HEYRAUD, M., FOWLER, S.W., BEASLEY,

T.M., “Vertical oceanic transport of alpha-radioactive nuclides by zooplankton fecal

pellets”, Natural Radiation Environment

III (Proc. Symp. Houston, 1978), Vol. 1

(GESELL, T.F., LOWDER, W.M., Eds), United States Department of Energy, Oak

Ridge, TN (1980) 502–513.

[8.43] HEYRAUD, M., DOMANSKI, P., CHERRY, R.D., FASHAM, M.J.R., Natural

tracers in dietary studies: Data for

210

Po and

210

Pb in decapod shrimp and other

pelagic organisms in the Northeast Atlantic Ocean, Mar. Biol.

97 (1988) 507–519.

[8.44] PENTREATH, R.J., Radionuclides in marine fish, Oceanogr. Mar. Biol. Annu.

Rev.

15 (1977) 365–460.

[8.45] RYAN, T.P., Transuranic biokinetic parameters for marine invertebrates: A review,

Environ. Int.

28 (2002) 83–96.

[8.46] CARVALHO, F.P., FOWLER, S.W., A double-tracer technique to determine the

relative importance of water and food as sources of polonium-210 to marine prawns

and fish, Mar. Ecol. Prog. Ser.

103 (1994) 251–264.

[8.47] CARVALHO, F.P., FOWLER, S.W., An experimental study on the bioaccumulation

and turnover of polonium-210 and lead-210 in marine shrimp, Mar. Ecol. Prog.

Ser.

102 (1993) 125–133.

[8.48] DURAND, J.P., et al.,

210

Po binding to metallothioneins and ferritin in the liver of

teleost marine fish, Mar. Ecol. Prog. Ser.

177 (1999) 189–196.

[8.49] WILDGUST, M.A., McDONALD, P., WHITE, K.N., Assimilation of

210

Po by the

mussel Mytilus edulis from the alga Isochrysis galbana, Mar. Biol.

136 (2000) 49–53.

[8.50] REINFELDER, J.R., FISHER, N.S., The assimilation of elements ingested by marine

copepods, Science

251 (1991) 794–796.

[8.51] FISHER, N.S., REINFELDER, J., “The trophic transfer of metals in marine systems”,

Metal Speciation and Bioavailabliity in Aquatic Systems (TESSIER, A., TURNER,

D.R., Eds), John Wiley and Sons, Chichester (1995)

363–406.

[8.52] STEWART, G.M., et al., Contrasting transfer of polonium-210 and lead-210 across

three trophic levels in marine plankton, Mar. Ecol. Prog. Ser.

290 (2005) 27–33.

[8.53] CARVALHO, F.P., Contribution a l’étude du cycle du polonium-210 et du plomb-210

dans l’environnement, PhD thesis, Univ. Nice Sophia Antipolis, Nice (1990).

[8.54] CARVALHO, F.P.,

210

Po in marine organisms: A wide range of natural radiation dose

domains, Radiat. Prot. Dosim.

24 (1988) 113–117.

[8.55] SMITH, D.J., TOWLER, H., Polonium-210 in cartilaginous fishes (Chondrichthyes)

from south-eastern Australian waters, Aust. J. Mar. Freshwater Res.

(1993)

727–733.

[8.56] CHERRY, R.D., HEYRAUD, M., RINDFUSS, R., Polonium-210 in teleost fish and

in marine mammals: Interfamily differences and a possible association between

polonium-210 and red muscle content, J. Environ. Radioact.

24 (1994) 273–291.

[8.57] CHERRY, R.D., HEYRAUD, M., Evidence of high natural radiation doses in certain

mid-water oceanic organisms, Science

218 (1982) 54–56.

[8.58] GODOY, J.M., et al.,

210

Po concentration in Perna perna mussels: Looking for

radiation effects, J. Environ. Radioact.

99 (2008) 631–640.

171

MARINE SYSTEMS

[8.59] PÉRÈS, J.M., Précis d’océanographie biologique, Presses Universitaires de France,

Paris (1976).

[8.60] NEWELL, R.C., Biology of Intertidal Animals, Marine Ecological Surveys,

Faversham

(1979).

[8.61] KAISER, M.J., et al., Marine Ecology: Processes, Systems, and Impacts, Oxford

University Press

(2005).

[8.62] WONG, K.M., HODGE, V.F., FOLSOM, T.R., Plutonium and polonium inside giant

brown algae, Nature

237 (1972) 460–462.

[8.63] FOLSOM, T.R., HODGE, V.F., Experiments suggesting some first steps in the

dispersal and disposal of plutonium and other alpha emitters in the ocean, Mar. Sci.

Commun.

1 (1975) 213–247.

[8.64] CARVALHO, F.P., FOWLER, S.W., Americium adsorption on the surface of

macrophytic algae, J. Environ. Radioact.

2 (1985) 311–317.

[8.65] FISHER, N.S., BURNS, K.A., CHERRY, R.D., HEYRAUD, M., Accumulation and

cellular distribution of

241

Am,

210

Po and

210

Pb in two marine algae, Mar. Ecol. Prog.

Ser.

11 (1983) 233–237.

[8.66] XUE, H.-B., STUMM, W., SIGG, L., The binding of heavy metals to algal surfaces,

Water Res.

22 (1998) 917–926.

[8.67] DAVIS, T.A., VOLESKY, B., MUCCI, A., A review of the biochemistry of heavy

metal biosorption by brown algae, Water Res.

37 (2003) 4311–4330.

[8.68] WILSON, R.C., WATTS, S.J., VIVES i BATLLE, J., McDONALD, P., Laboratory

and field studies of polonium and plutonium in marine plankton, J. Environ.

Radioact.

100 (2009) 665–669.

[8.69] SURIYANARAYANAN, S., BRAHMANANDHAN, G.M., SAMIVEL, K.,

RAVIKUMAR, S., SHAHUL HAMEED, P., Assessment of

210

Po and

210

Pb in marine

biota of the Mallipattinam ecosystem of Tamil Nadu, India, J. Environ. Radioact.

101

(2010)

1007–1010.

[8.70] CARVALHO, F.P.,

210

Po and

210

Pb intake by the Portuguese population: The

contribution of seafood in the dietary intake of

210

Po and

210

Pb, Health Phys. 69

(1995)

469–480.

[8.71] FEROZ KHAN, M., UMARAJESWARI, S., GODWIN WESLEY, S., Biomonitoring

210

Po and

210

Pb in marine brachyuran crabs collected along the coast of Kudankulam,

Gulf of Mannar (GOM), India, J. Environ. Monit.

13 (2011) 553–562.

[8.72] HEYRAUD, M., et al., Polonium-210 and lead-210 in edible molluscs from near the

Cape of Good Hope: Sources of variability in polonium-210 concentrations,

J. Environ. Radioact. 24 (1994) 253–272.

[8.73] CHARMASSON, S., FAOUDER, A., LOYEN, J., COSSON, R.P., SARRADIN,

P.M.,

210

Po and

210

Pb in the tissues of the deep-sea hydrothermal vent mussel

Bathymodiolus azoricus from the Menez Gwen field (Mid-Atlantic Ridge), Sci. Total

Environ.

409 (2011) 771–777.

[8.74] BUSTAMANTE, P., GERMAIN, P., LECLERC, G., MIRAMAND, P., Concentration

and distribution of

210

Po in the tissues of the scallop Chlamys varia and the mussel

Mytilus edulis from the coasts of Charente-Maritime (France), Mar. Pollut. Bull.

(2002)

997–1002.

172

CHAPTER 8

[8.75] DESIDERI, D., MELI, M.A., ROSELLI, C., A biomonitoring study:

210

Po and heavy

metals in marine organisms from the Adriatic Sea (Italy), J. Radioanal. Nucl.

Chem.

285 (2010) 373–382.

[8.76] UĞUR, A., YENER, G., BAŞSARI, A., Trace metals and

210

Po (

210

Pb) concentrations

in mussels (Mytilus galloprovincialis) consumed at western Anatolia, Appl. Radiat.

Isot.

57 (2002) 565–571.

[8.77] ROŽMARIĆ, M., et al.,

210

Po and

210

Pb activity concentrations in Mytilus

galloprovincialis from Croatian Adriatic coast with the related dose assessment to the

coastal population, Chemosphere

87 (2012) 1295–1300.

[8.78] GOUVEA, R.C., SANTOS, P.L., DUTRA, I.R., Lead-210 and polonium-210

concentrations in some species of marine molluscs, Sci. Total Environ.

112

(1992)

263–267.

[8.79] McDONALD, P., BAXTER, M.S., SCOTT, E.M., Technological enhancement of

natural radionuclides in the marine environment, J. Environ. Radioact.

(1996)

67–90.

[8.80] CARVALHO, F.P., OLIVEIRA, J.M., ALBERTO, G., Factors affecting

210

Po and

210

Pb activity concentrations in mussels and implications for environmental

bio-monitoring programmes, J. Environ. Radioact.

102 (2011) 128–137.

[8.81] INTERNATIONAL ATOMIC ENERGY AGENCY, Sources of Radioactivity in the

Marine Environment and their Relative Contributions to Overall Dose Assessment

from Marine Radioactivity, IAEA-TECDOC-838, IAEA, Vienna (1995).

[8.82] THEBAULT, H., RODRIGUEZ Y BAENA, A.M., Mediterranean Mussel Watch: A

regional program for detecting radionuclides, trace- and emerging-contaminants,

Rapp. Comm. Int. Mer Médit.

38 (2007) 41.

[8.83] DAHLGAARD, H., Polonium-210 in mussels and fish from the Baltic–North Sea

estuary, J. Environ. Radioact.

32 (1996) 91–96.

[8.84] CARVALHO, F.P., OLIVEIRA, J.M., MALTA, M., Radionuclides in deep-sea fish

and other organisms from the North Atlantic Ocean, ICES J. Mar. Sci.

(2011)

333–340.

[8.85] CARVALHO, F.P., OLIVEIRA, J.M., ALBERTO, G., VIVES i BATLLE, J.,

Allometric relationships of

210

Po and

210

Pb in mussels and their application to

environmental monitoring, Mar. Pollut. Bull.

60 (2010) 1734–1742.

[8.86] YAMAMOTO, M., et al., Polonium-210 and lead-210 in marine organisms: Intake

levels for Japanese, J. Radioanal. Nucl. Chem.

178 (1994) 81–90.

[8.87] NOSHKIN, V.E., ROBISON, W.L., WONG, K.M., Concentration of

210

Po and

210

in the diet at the Marshall Islands, Sci. Total Environ.

155 (1994) 87–104.

[8.88] POLLARD, D., RYAN, T.P., DOWDALL, A., The dose to Irish seafood consumers

from

210

Po, Radiat. Prot. Dosim. 75 (1998) 139–142.

[8.89] KANNAN, V., IYENGAR, M.A.R., RAMESH, R., Dose estimates to the public from

210

Po ingestion via dietary sources at Kalpakkam (India), Appl. Radiat. Isot. 54

(2001)

663–674.

[8.90] MISHRA, S., BHALKE, S., PANDIT, G.G., PURANIK, V.D., Estimation of

210

and its risk to human beings due to consumption of marine species at Mumbai, India,

Chemosphere

76 (2009) 402–406.

173

MARINE SYSTEMS

[8.91] SKWARZEC, B., Accumulation of

210

Po in selected species of Baltic fish, J. Environ.

Radioact. 8 (1988) 111–118.

[8.92] CHERRY, R.D., HEYRAUD, M., JAMES, A.G., Diet prediction in common clupeoid

fish using polonium-210 data, J. Environ. Radioact. 10 (1989) 47–65.

[8.93] LAZORENKO, G.E., POLIKARPOV, G.G., BOLTACHEV, A.R., Natural

radioelement polonium in primary ecological groups of Black Sea fishes, Russ. J.

Mar. Biol. 28 (2002) 52–56.

[8.94] HOFFMAN, F.L., HODGE, V.F., FOLSOM, T.R., Polonium radioactivity in certain

mid-water fish of the Eastern Temporal Pacific, Health Phys. 26 (1974) 65–70.

[8.95] MALTA, M., CARVALHO, F.P., Radionuclides in marine mammals off the

Portuguese coast, J. Environ. Radioact.

102 (2011) 473–478.

[8.96] GODOY, J.M., SICILIANO, S., DE CARVALHO, Z.L., DE MOURA, J.F., GODOY,

M.L.D.P.,

210

Polonium content of small cetaceans from Southeastern Brazil, J.

Environ. Radioact. 106 (2012) 35–39.

[8.97] HASSONA, R.K., SAM, A.K., OSMAN, O.I., SIRELKHATIM, D.A., LAROSA, J.,

Assessment of committed effective dose due to consumption of Red Sea coral reef

fishes collected from the local market (Sudan), Sci. Total Environ. 393

(2008) 214–218.

[8.98] HORN, M.H., MESSER, K.S., Fish guts as chemical reactors: A model of the

alimentary canals of marine herbivorous fishes, Mar. Biol. 113 (1992) 527–535.

[8.99] SKWARZEC, B., FABISIAK, J., Bioaccumulation of polonium

210

Po in marine

birds, J. Environ. Radioact. 93 (2007) 119–126.

[8.100] GWYNN, J.P., GABRIELSEN, G.W., JÆGER, I., LIND, B., “Bioaccumulation of

210

Po and

210

Pb in Arctic seabirds”, Radioecology and Environmental Radioactivity,

(Poster Proc. Int. Conf. Bergen, 2008), Part I (STRAND, P., BROWN, J., JOLLE, T.,

Eds), Bergen (2008) 101–104.

[8.101] JEFFREE, R.A., CARVALHO, F., FOWLER, S.W., FARBER-LORDA, J.,

Mechanism for enhanced uptake of radionuclides by zooplankton in French

Polynesian oligotrophic waters, Environ. Sci. Technol. 31 (1997) 2584–2588.

[8.102] SKWARZEC, B., BOJANOWSKI, R.,

210

Po content in sea water and its accumulation

in southern Baltic plankton, Mar. Biol. 97 (1988) 301–307.

[8.103] STEWART, G.M., FISHER, N.S., Experimental studies on the accumulation of

polonium-210 by marine phytoplankton, Limnol. Oceanogr. 48 (2003) 1193–1201.

[8.104] COWLES, T.J., OLSON, R.J., CHISHOLM, S.W., Food selection by copepods:

Discrimination on the basis of food quality, Mar. Biol. 100 (1988) 41–49.

[8.105] CHERRY, M.I., CHERRY, R.D., HEYRAUD, M., Polonium-210 and lead-210 in

Antarctic marine biota and sea water, Mar. Biol. 96 (1987) 441–449.

[8.106] CHERRY, R.D., HEYRAUD, M., Polonium-210 content of marine shrimp: Variation

with biological and environmental factors, Mar. Biol. 65 (1981) 165–175.

[8.107] SWINBANKS, D.D., SHIRAYAMA, Y., High levels of natural radionuclides in a

deep-sea infaunal xenophyophore, Nature 320 (1986) 354–358.

[8.108] MARSHALL, N.B., Developments in Deep-Sea Biology, Blandford Press,

Dorset (1979).

174

CHAPTER 8

[8.109] INTERNATIONAL ATOMIC ENERGY AGENCY, Handbook of Parameter Values

for the Prediction of Radionuclide Transfer to Wildlife, Technical Reports Series

No. 479, IAEA, Vienna (2014).

[8.110] ALAM, L., MOHAMED, C.A.R., A mini review on bioaccumulation of

210

Po by

marine organisms, Int. Food Res. J. 18 (2011) 1–10.

[8.111] FOWLER, S.W.,

210

Po in the marine environment with emphasis on its behaviour

within the biosphere, J. Environ. Radioact. 102 (2011) 448–461.

175

Chapter 9

POLONIUM IN HUMANS

9.1. INTRODUCTION

Polonium-210 is one of the most radiotoxic natural radioactive isotopes

known to humans due to its relatively long half-life (138 days) compared to

many radon progeny in the uranium series, its high specific activity (166 TBq/g)

and its emission of high linear energy transfer alpha radiation (5.304 MeV) [9.1].

On account of its high specific activity, it is also one of the rarest elements in

nature. In addition to occurring naturally, it can also be produced by neutron

activation of

209

Bi. The fractional intestinal uptake was given as 10–90% by

Hill [9.2] in 1965 and as 50% by the International Commission on Radiological

Protection (ICRP) in 1993 [9.3]. The daily intake is in the range of 37–370 mBq

for a European diet [9.2], but higher daily intake occurs in seafood rich countries,

such as Japan (480–690 mBq) [9.4]. The biological half-life is in the range of

26–100 days (see Table 9.1). There is a prompt excretion of about 3% within a

period of 9 days [9.6].

With a half-life of 22.3 years,

210

Pb is the parent nuclide of

210

Po and builds

up from the decay of

222

Rn in the uranium series. The inert noble gas

222

is generated in all soils and rocks during the decay of

226

Ra and is subject to

exhalation from the ground to the atmosphere. The daughter radionuclides of

222

Ra are then deposited by rain, wind and gravitational settling onto terrestrial

vegetation. Hence, humans are exposed to radioactive polonium as a consequence

of natural processes, and its estimated contribution to the annual background

effective dose is 120 μSv/a [9.1].

Polonium accumulates readily in biota, mainly owing to its affinity

for protein

[9.2], and hence easily passes through the food chain to humans.

For example,

210

Po has been found to bind to unidentified proteins with high

molecular weight (70

kilodaltons), which could be haemocyanin or metal binding

enzymes

[9.15]. Protein rich foods, such as crustaceans and other shellfish,

contain the highest activity concentrations of

210

Po. Elevated body burdens of

210

Po are also observed in humans that consume large amounts of reindeer meat,

for example subarctic native populations

[9.16, 9.17].

Most polonium enters the body through the consumption of food and then

reaches the gastrointestinal tract, from which much of it is excreted with faeces.

Polonium and lead absorbed into plasma are distributed throughout the soft tissue

of the body and accumulate in the liver and kidneys

[9.6]. The accumulation

176

CHAPTER 9

210

Pb in the skeleton provides another source of

210

Po over time. Since

210

delivers the main absorbed dose of the Pb–Bi–Po decay chain, due to its alpha

emission (E

= 5.3 MeV), the behaviour of this isotope is of great interest.

Human biomonitoring is conducted mainly by urine and faecal sampling.

Blood sampling, however, has the advantage of being an integrated indicator

of polonium deposits across the metabolically active parts of the body;

hence, changes in the polonium body burden should be reflected in the blood

concentration. There is an absence, however, of suitable analogues (i.e. elements

with similar biochemical behaviour to polonium) to help to understand the

metabolic pathways in the human body which influence

210

Po biokinetics.

Polonium is a chalcogen element and, hence, it has chemical similarities to the

elements of the oxygen group in the periodic table (Group

VIA). According to

Waska et al.

[9.15],

210

Po has similar pathway characteristics in biological matter

as another chalcogen, selenium.

Most of the data for polonium activity concentrations in human tissues

and other matrices are from the 1960s and 1970s. The analytical procedures at

TABLE 9.1. A SURVEY OF AVAILABLE DATA FOR FRACTIONAL UPTAKE

AND BIOLOGICAL HALF-LIFE OF

210

Po IN HUMANS

Matrix N

Gastrointestinal

uptake fraction

Biological half-life

(days)

Ref.

Crab meat 7 0.60–0.94 —

[9.5]

Caribou meat 14 0.31–0.70 100 [9.6]

Normal food 15–63

—

0.35

0.1–0.3

0.05–0.45

0.1

—

30–60

73 (Inuit), 114 (Caucasians)

[9.7]

[9.8]

[9.9]

[9.10]

[9.11]

Po in water 5 0.38–0.78 31 [9.12]

Puncture wound

(accident)

1 —

26 [9.13]

Shellfish 7

—

0.44–0.52

0.50

28–51

[9.14]

[9.3]

—: data not available.

Depending on the organ.

177

POLONIUM IN HUMANS

that time did not undergo any inter-calibration or quality control [9.18]. Yield

determinants were often not used for the analytical procedure and a recovery of

100% was assumed, meaning that the true activity concentrations of

210

Po in the

samples were often underestimated. Furthermore, the results might be affected

by in

vitro buildup of

210

Po from

210

Pb. Samples need to be analysed shortly after

collection, otherwise

210

would also have to be determined to correlate the

result to the day of collection.

9.2. INGESTION AND INHALATION

The average daily intake of naturally occurring

210

Po for people following a

typical European diet is estimated to be 37–370 mBq [9.2, 9.19]. Populations that

consume marine food, such as crustaceans and shellfish, can have a much higher

intake, and this can be exacerbated for people living in areas of low oceanic

primary productivity, such as on the islands of the central Pacific Ocean [9.20]

(see also Chapter 8). Certain critical groups, such as radium dial painters, exhibit

exceptionally high levels [9.2, 9.21]. It is also clear that populations that consume

caribou and reindeer have higher concentrations in their body [9.11, 9.16].

A health risk evaluation for ingestion of

210

Po has been performed by Scott [9.10]

and an internal dose assessment using a biokinetic model is reported by Li et

al. [9.9].

The main sources of

210

Po (and its progenitor

210

Pb) in humans are food,

water and inhalation, with additional in

vivo contributions from

222

Rn dissolved

in the body as well as from

210

Pb and

226

Ra in bones. A 1966 study shows

indications of higher levels of

210

Po in the organs of smokers [9.22]. The intake

via inhalation is much smaller than that from food [9.7], even at uranium milling

facilities [9.23].

9.3. DISTRIBUTION IN TISSUES

The highest activity concentrations of

210

Po in humans are found in the

liver, kidneys and bone. The content in the skeleton and muscle tissue comprises

about 70% of the total body burden. For bone, a significant proportion is due to

buildup from

210

Pb [9.24]. Activity concentrations generally increase with age by

accumulation of

226

Ra and

210

Pb. For other animals, the concentrations are also

highest in the liver and kidneys, but the fraction of body content is different due

to a large portion being contained in the fur [9.25]. Table 9.2 shows the results

of a survey of

210

Po activity concentrations and the

210

Po:

210

Pb activity ratio in

different human organs. Almost all of the data are from the 1960s and 1970s,

178

CHAPTER 9

TABLE 9.2. POLONIUM IN HUMAN ORGANS

Matrix

Activity

concentration

(mBq/kg, FW)

Provenance Ref.

Lung 326 Former USSR

[9.7]; bb = 1.7%;

210

Po:

210

Pb = 1.4

318 USA, smokers [9.2]

118 USA, non-smokers [9.2];

210

Po:

210

Pb = 1.7

200 United Kingdom [9.16]

189 USA [9.17];

210

Po:

210

Pb = 0.75

Lung

parenchyma

290 USA, smokers [9.22]

93 USA, non-smokers [9.22]

Lung bronchi 470 USA, smokers [9.22]

370 USA, non-smokers [9.22]

Lung nodes 4 000 USA, smokers [9.22]

1 400 USA, non-smokers [9.22]

Liver 970 Former USSR

[9.7]; bb = 8.7%;

210

Po:

210

Pb = 2.2

400 USA [9.24]; bb = 1.7%

625 United Kingdom [9.16]

7 000 Radium dial painters [9.2];

210

Po:

210

Pb = 1.0

740 USA, smokers [9.2]

550 USA, non-smokers [9.2];

210

Po:

210

Pb = 2.1

Liver 529 USA [9.17];

210

Po:

210

Pb = 2.1

503 Children, AK, USA [9.17];

210

Po:

210

Pb = 2.1

530 USA, smokers [9.22]

220 USA, non-smokers [9.22]

750 Former USSR [9.26]

Bone 1 070 United Kingdom [9.16]

55 000 Radium dial painters [9.2];

210

Po:

210

Pb = 1.0

8 700

USA [9.24]; bb = 63%

1 070 Former USSR [9.7]; bb = 50.2%

1 700 —

[9.26]

179

POLONIUM IN HUMANS

TABLE 9.2. POLONIUM IN HUMAN ORGANS (cont.)

Matrix

Activity

concentration

(mBq/kg, FW)

Provenance Ref.

Bone (femur) 1 006 Children, AK, USA [9.17];

210

Po:

210

Pb = 0.82

Blood 7 900 Radium dial painters [9.21];

210

Po:

210

Pb = 0.7

292

Lapland, Finland [9.27]

Helsinki, Finland [9.27]

370

Sami, male, Sweden [9.28];

210

Po:

210

Pb = 3.3

180

Sami, female, Sweden [9.28];

210

Po:

210

Pb = 2.6

Muscle 111 Former USSR [9.7]; bb = 17.6%;

210

Po:

210

Pb = 0.8

222 USA [9.24]; bb = 17%

22 USA, smokers [9.22]

37 USA, non-smokers [9.22]

Hair

2 000–11 000

Sweden [9.29];

210

Po:

210

Pb = 2.2

900 000 Radium dial painters [9.2];

210

Po:

210

Pb = 1.0

3 300 Former USSR [9.7];

210

Po:

210

Pb = 2.2

9 000–22 000

India, non-smokers [9.30]

23 000

India, smokers [9.30]

800–14 000

USA, non-smokers [9.31]

1 200–12 000

USA, smokers [9.31]

5 300 Former USSR [9.26]

Kidney 640 United Kingdom [9.16]

970 Former USSR

[9.7]; bb = 1.2%;

210

Po:

210

Pb = 2.8

9 300 Radium dial painters [9.2];

210

Po:

210

Pb = 9

758 USA, smokers [9.2]

555 USA, non-smokers [9.2];

210

Po:

210

Pb = 3.2

384 USA [9.17];

210

Po:

210

Pb = 2.6

277 Children, AK, USA [9.17];

210

Po:

210

Pb = 2.6

330 USA, smokers [9.22]

210 USA, non-smokers [9.22]

180

CHAPTER 9

TABLE 9.2. POLONIUM IN HUMAN ORGANS (cont.)

Matrix

Activity

concentration

(mBq/kg, FW)

Provenance Ref.

Placenta 122 United Kingdom [9.16]

740–4 000 Reindeer and caribou

consumers, Canada

[9.16];

210

Po:

210

Pb = 6.3

400 N. Canada [9.11]

Gonads 144 USA, smokers [9.2]

104 USA, non-smokers [9.2]

255 USA [9.17];

210

Po:

210

Pb = 1.1

Testis 144 United Kingdom [9.16]

Spleen 117 USA [9.17];

210

Po:

210

Pb = 1.8

22 Children, AK, USA [9.17];

210

Po:

210

Pb = 0.6

220 Former USSR [9.7]; bb = 0.8%;

210

Po:

210

Pb = 0.2

Pancreas 107 USA [9.17];

210

Po:

210

Pb = 1.8

Thyroid 189 USA [9.17]

Heart 19 USA [9.17]

70 USA, smokers [9.22]

37 USA, non-smokers [9.22]

Eye, choroid

and iris

1 000–10 000 United Kingdom [9.32]

Bone marrow 24 United Kingdom [9.33]

Note: bb — body burden.

mBq/kg, ash.

—: data not available.

mBq/L.

mBq/kg (DW).

when access to autopsy samples was permitted. However, estimation of body

content can be achieved by analysing urine and faeces. Hair also seems to be a

good indicator of polonium contamination [9.12].

181

POLONIUM IN HUMANS

Polonium has an affinity for proteins. For example, gelatin is reported to

contain 1500 mBq/kg of

210

Po and insulin 23 000–70 000 mBq/kg [9.2]. In the

past, people with diabetes may have been exposed to elevated levels of

210

Po in

natural insulin. Insulin is now made artificially; it is no longer extracted from

the pancreas and does not contain any

210

Po. However, another critical group

might be people undergoing dialysis. If polonium became associated with protein

molecules in the body too large to pass through the membrane during dialysis,

the radionuclide might build up in the body. This has not yet been investigated.

9.4. RETENTION AND BIOLOGICAL HALF-LIVES

The kinetics of polonium are rather complex (see Fig. 9.1) and are

influenced by its wide variety of potential oxidation states (−2, +2, +4 and +6;

see Section 3.2). Polonium can be released from one organ and recirculated.

Source: Figure 5 of Ref. [9.34].

FIG. 9.1. Compartments of the systemic model, paths of movement of polonium between

compartments, and connections between the systemic model and user-supplied models of the

respiratory tract, gastrointestinal tract and wounds (reproduced with permission courtesy of

Elsevier).

182

CHAPTER 9

The in vivo concentration of

210

Pb causes buildup of

210

Po, especially in bone

where the biological half-life of

210

Pb is long. In addition,

226

Ra plays a role as

a precursor of

210

Po. Several studies report values between 10 and 60 days for

the effective biological half-life of

210

Po in humans (see Fig. 9.2), and fractional

uptake following ingestion ranges from a few percent to 90%. The dose factor is

thereby quite uncertain. All those factors probably depend, to a large degree, on

the composition of the foodstuff with which polonium is ingested.

The highest activity concentrations in humans are found in the liver and, to

a lesser extent, in the kidneys. It has also been found that hair is a good indicator

of polonium contamination and that hair is an important excretion route in

addition to faecal and urinary routes

[9.11]. Due to its complex biokinetics and its

array of exposure and excretion pathways, the biological and effective half-lives

of polonium are highly variable (see Fig.

9.2).

Leggett and Eckerman [9.34] find that:

“Although there is a relatively large database on the biological behavior

of polonium in man and laboratory animals, some important aspects of the

biokinetics of polonium in man have not been characterized with much

certainty. A central problem is that a substantial portion of the data on

urinary excretion of

210

Po by human subjects may not be reliable. Although

relatively detailed data are available from a controlled study on human

subjects, these data are limited not only by potential errors in the urinary

excretion data but also by the fact that the state of health of the subjects

could have affected the biokinetics of polonium.”

Source: Figure 2 of Ref. [9.34].

FIG. 9.2. Distribution of effective half-lives derived from urinary excretion records for a large

number of workers using polonium (reproduced with permission courtesy of Elsevier).

183

POLONIUM IN HUMANS

Hence, the results might be poorly representative of the general population.

Conclusions about data on laboratory animals are complicated by the variation in

biokinetics of polonium due to differences between “species, route of exposure,

and the chemical form of polonium taken into the body” [9.34].

The human gastrointestinal uptake fraction f

has been established by a

number of different studies, with a wide range of results (for fractional uptake and

biological half-life for

210

Po, see Table 9.1). Thomas et al. [9.6] show that a total

ingestion of 20

Bq with caribou meat resulted in a maximum of 3.2 Bq/d in faecal

excretion over 4–6

days after intake and a maximum of 0.32 Bq/d of urinary

excretion from 5 to 10

days after ingestion. From these results, a gastrointestinal

uptake fraction of 0.56

± 0.04 was established. In the 1950s, Silberstein et al. [9.8]

published gastrointestinal uptake fractions as low as 0.1–0.3. The ICRP

[9.3] has

increased its reported gastrointestinal uptake fraction from 10% to 50%, which,

according to other studies, still seems too small. Hunt and Allington

[9.5] report

gastrointestinal uptake fractions in the range of 0.6–0.94, with a mean of 0.76,

after analysing a study in 1991 in which seven volunteers ingested crab meat

with a total

210

Po activity intake of up to 44.2 Bq. A range of 10–25% for daily

faecal excretion and a maximum daily urine excretion of 0.2–0.3% is reported.

Additional data from the 1970s are provided by Landinskaya et

al. [9.7], who

analysed tissue from humans and estimated their intake from normal food, and in

Ref. [9.11], which reports on the retention of

210

Po in Inuit and Caucasians.

In a 2012 study on radioactive polonium biokinetics by Henricsson [9.12],

volunteers ingested

209

Po in a PoCl

solution. The excretion via faeces and urine

of four of the subjects is shown in Fig.

9.3. It is well known that there are some

critical groups, such as those who eat reindeer or caribou and those who consume

large quantities of marine food (particularly molluscs and crustaceans).

In conclusion, there have been limited data generated on the kinetics and

organ distribution of polonium in humans since the 1970s. Only about eight

studies have been performed and, of these, only three experimental studies were

conducted after the 1970s. Considering that

210

Po is the largest contributor of

radiological dose to humans (from alpha particle emission), this is surprising.

There have been some critical population groups identified, such as those

who consume large amounts of seafood or consume caribou and reindeer.

However, the large variations in fractional uptake and biological residence time

results in a large uncertainty in the estimated dose factor for oral intake of

210

Po.

The fractional intake probably depends on the food composition with which

polonium is consumed, which will vary between regions and lifestyles

[9.16]. It

would be of interest to conduct a survey of the main food for ingestion of

210

Po in

different countries, perhaps by analysing hair, which could be a good indicator of

internal contamination

[9.11, 9.30].

184

CHAPTER 9

REFERENCES TO CHAPTER 9

[9.1] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Sources and Effects of Ionizing Radiation UNSCEAR 2000 Report to

the General Assembly, with Scientific Annexes, Vol. I: Sources, United Nations,

New York (2000) Annex B.

[9.2] HILL, C.R.,

210

Po in man, letter to K. Liden, Lund University, personal communication,

1965.

[9.3] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Age-dependent Doses to Members of the Public from Intake of Radionuclides —

Part

2: Ingestion Dose Coefficients, Publication 67, Pergamon Press, Oxford and

New

York (1993).

[9.4] YAMAMOTO, M., et al., Polonium-210 and lead-210 in marine organisms: Intake

levels for Japanese, J. Radioanal. Nucl. Chem.

178 (1994) 81–90.

[9.5] HUNT, G.J., ALLINGTON, D.J., Absorption of environmental polonium-210 by the

human gut, J. Radiol. Prot.

13 (1993) 119–126.

[9.6] THOMAS, P.A., FISENNE, I., CHORNEY, D., BAWEJA, A.S., TRACY, B.L., Human

absorption and retention of polonium-210 from caribou meat, Radiat. Prot. Dosim.

(2001)

241–250.

[9.7] LADINSKAYA, L.A., PARFENOV, Yu.D., POPOV, D.K., FEDOROVA, A.V.,

210

and

210

Po content in air, water, foodstuffs, and the human body, Arch. Environ.

Health

27 (1973) 254–258.

0.0%

0.5%

1.0%

1.5%

2.0%

2.5%

3.0%

3.5%

4.0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

100%

0 200 400 600 800 1000 1200 1400 1600

Accumulatedfr actionofintake

Hoursfromintake

Source: Data from Ref. [9.12].

FIG. 9.3. Accumulated excretion of

209

Po via faeces.

185

POLONIUM IN HUMANS

[9.8] SILBERSTEIN, H.E., VALENTINE, W.N., MINTO, W.L., LAWRENCE, J.S., FINK,

R.M., “Studies of polonium metabolism in human subjects”, Biological Studies with

Polonium, Radium and Plutonium (FINK, R.M., Ed.), McGraw-Hill, New

York (1950).

[9.9] LI, W.B., GERSTMANN, U., GIUSSANI, A., OEH, U., PARETZKE, H.G., Internal

dose assessment of

210

Po using biokinetic modeling and urinary excretion measurement,

Radiat. Environ. Biophys.

47 (2008) 101–110.

[9.10] SCOTT, B.R., Health risk evaluations for ingestion exposure of humans to

polonium-210, Dose Response

5 (2007) 94–122.

[9.11] ARGONNE NATIONAL LABORATORY, Renal Excretion Rates of

137

Cs,

210

Pb,

210

and Stable Lead by Northern Canadians, ANL-7860, Part II, Argonne Natl Lab.,

IL (1971).

[9.12] HENRICSSON, C.F., RANEBO, Y., HANSSON, M., RÄÄF, C.L., HOLM, E., A

biokinetic study of

209

Po in man, Sci. Total Environ. 437 (2012) 384–389.

[9.13] WRAIGHT, J.C., STRONG, R., “A case study on the retention and excretion of

internally deposited polonium”, Radiation Protection: Theory and Practice (Proc. 4th

Int. Symp. Malvern, 1989), Society for Radiological Protection, Berkeley

(GOLDFINCH, E.P., Ed.) (1989)

227–230.

[9.14] HUNT, G.J., RUMNEY, H.S., The human alimentary tract transfer and body retention

of environmental polonium-210, J. Radiol. Prot.

27 (2007) 405–426.

[9.15] WASKA, H., KIM, S., KIM, G., KANG, M.R., KIM, G.B., Distribution patterns of

chalcogens (S, Se, Te, and

210

Po) in various tissues of a squid, Todarodes pacificus, Sci.

Total Environ.

392 (2008) 218–224.

[9.16] HILL, C.R., Polonium-210 content of human tissues in relation to dietary habit,

Science 152 (1966)

1261–1262.

[9.17] BLANCHARD, R.L., “Relationship between

210

Po and

210

Pb in man and his

environment”, Radioecological Concentration Processes (Proc. Int. Symp. Stockholm,

1966) (ÅBERG, B., HUNGATE, F.P., Eds), Pergamon Press, Oxford (1967)

275–280.

[9.18] HENRICSSON, F., RANEBO, Y., HOLM, E., ROOS, P., Aspects on the analysis of

210

Po, J. Environ. Radioact. 102 (2011) 415–419.

[9.19] BEOWULF, G., HERMANN, M., Intake and excretion of natural radionuclides

210

and

210

Po by humans, Strahlenther. 131 (1966) 281–296.

[9.20] JEFFREE, R.A., SZYMCZAK, R., Enhancing effect of marine oligotrophy on

environmental concentrations of particle-reactive trace elements, Environ. Sci.

Technol.

34 (2000) 1966–1969.

[9.21] HOLTZMAN, R.B., “

210

Pb and

210

Po metabolism in radium-dial painters”, paper

presented at the Meeting on Biology and Ecology of Polonium and Radiolead,

Surrey,

1970.

[9.22] FERRI, E.S., BARATTA, E.J., Polonium 210 in tobacco, cigarette smoke, and selected

human organs, Public Health Rep.

81 (1966) 121–127.

[9.23] NUCLEAR REGULATORY COMMISSION, Final Generic Environmental Impact

Statement on Uranium Milling, Project M-25, NUREG-0706, Office of Nuclear

Material Safety and Safeguards, NRC, Washington, DC (1980).

186

CHAPTER 9

[9.24] HOLTZMAN, R.B., Measurement of the natural content of RaD (

210

Pb) and RaF

(

210

Po) in human bone estimates of whole-body burdens, Health Phys. 9

(1963)

385–400.

[9.25] SCHRECKHISE, R.G., WALTERS, R.L., Internal distribution and milk secretion of

polonium-210 after oral administration to a lactating goat, J. Dairy Sci.

(1969)

1867–1869.

[9.26] ANTONOVA, V.A., Po-210 in the hair, bone tissue and liver of man, Gig. Sanit. 10

(1971) 52–55 (in

Russian).

[9.27] KAURANEN, P., MIETTINEN, J.K.,

210

Po and

210

Pb in the arctic food chain and the

natural radiation exposure of Lapps, Health Phys.

16 (1969) 287–295.

[9.28] PERSSON, B.R., “Lead-210, polonium-210 and stable lead in the food-chain lichen,

reindeer and man”, The Natural Radiation Environment

II (Proc. 2nd Int. Symp.

Houston, 1972) (ADAMS, J.A.S., LOWDER, W.M., GESELL, T.F., Eds), United

States Department of Commerce, Springfield, VA (1972)

347–367.

[9.29] HOLM, E., et al., Hair and Feathers as Indicator of Internal Contamination of

210

and

210

Pb, NKS-217, Nordic Nuclear Safety Research, Roskilde (2010).

[9.30] RATHI, C.R., ROSS, E.M., WESLEY, S.G., Polonium-210 activity in human hair

samples and factors affecting its accumulation, Iran. J. Radiat. Res.

9 (2011) 41–47.

[9.31] LYKKEN, G.I., ALKHATIB, H.A., “Analysis of hair for polonium-210 α-particle

emissions”, Communicating the Radon Issue (Proc. Int. Radon Conf. Colorado, 1993)

(1993)

347–367.

[9.32] HUNT, V.R., “Concentrations of

210

Po,

226

Ra, and

228

Th in the choroid of the eye,

particularly in cattle”, Radioecological Concentration Processes (Proc. Int. Symp.

Stockholm, 1966) (ÅBERG, B., HUNGATE, F.P., Eds), Pergamon Press, Oxford

(1967)

303–311.

[9.33] SALMON, P.L., CLAYTON, R.F., HENSHAW, D.L., Measurement of

210

Po in

biopsied human red bone marrow, Radiat. Prot. Dosim.

78 (1998) 127–131.

[9.34] LEGGETT, R.W., ECKERMAN, K.F., A systemic biokinetic model for polonium, Sci.

Total Environ.

275 (2001) 109–125.

187

Chapter 10

RADIOLOGICAL DOSE ASSESSMENT FOR

210

TO HUMANS AND OTHER BIOTA

10.1. MODELS AND DATA FOR ESTIMATING INTERNAL EXPOSURES

TO HUMANS

Dosimetric models to estimate internal exposures in humans, subsequent to

incorporation of radionuclides via ingestion and inhalation, are developed by the

International Commission on Radiological Protection (ICRP). The most recent

models for

210

Po are described in ICRP Publications 67 [10.1] and 72 [10.2] for

ingestion, and in ICRP Publications 71 [10.3] and 72 [10.2] for inhalation. The end

points of those models are dose coefficients for ingestion and inhalation, which

are calculated for six age groups (3 month old children; 1, 5, 10 and 15 year old

children; and adults). The dose coefficients are based on the equivalent dose rates

received until the age of 70 years following an acute intake at the age of intake:

for a one year old child, the dose contributions are integrated over a period of

69 years; for adults, the dose is integrated over a period of 50 years.

The key processes in the biokinetic models are absorption of the

radionuclides into the blood and the distribution and retention in the organism.

Hence, the time integrated concentrations in the organs and tissues and the

number of disintegrations during the integration period are calculated from these

quantities. For a specific organ, the integrated absorbed dose is calculated from

the following:

(a) The number of decays during the integration period;

(b) The energy per decay;

organ;

(d) The energy deposited in a specific organ from radionuclides decaying in

other organs and tissues.

The model parameters applied in the biokinetic models are derived from

all available and appropriate experiments and observations on the uptake and

distribution of radionuclides, many of which determined from animal studies.

Following ingestion, the transfer of

210

Po in the ICRP model from the gut

to the blood (resorption) is age dependent. The highest resorption is applied for

infants (3

months) with 100%, and 50% for all other age groups [10.1].

188

CHAPTER 10

Distribution and retention are described by complex model simulation of

the fluxes of

210

Po between different organs and tissues. The most important

process for long term retention is the exchange of

210

Po between the kidneys,

liver and skeleton.

To calculate dose per unit intake values, it is assumed that for

210

entering the systemic circulation, fractions of 0.05, 0.1, 0.1, 0.3 and 0.45 are

deposited in the spleen, kidneys, red bone marrow, liver and the rest of the body,

respectively [10.1], where it is retained with a half-life of 50 days. For

210

that has entered the excretory system, a urinary:faecal excretion ratio of 1:2 is

assumed.

The transfer to the blood from the lungs depends on the age and the

chemical form of

210

Po. In general, the ICRP takes into account three absorption

classes: slow (S), medium (M) and fast (F) retention. In Tables 10.1–10.4, the

dose coefficients for

210

Po are summarized for ingestion and inhalation. These are

given as effective dose per unit intake (Sv/Bq), and all of the organs considered

in Refs [10.1, 10.3] are listed. For ingestion of

210

Po, the organs with the highest

dose coefficients are the kidneys, liver, bone surface, red bone marrow and

spleen. For inhalation, the dose coefficients for the lung and the extrathoracic

airways are also important. There are large differences between the dose

coefficients for the different absorption classes, in particular when comparing

organ doses (see Tables 10.2–10.4). To estimate the effective dose from

210

Po,

the highest dose coefficients are calculated for the absorption class S, which are

about a factor of seven higher than for inhalation class F. This is because the less

soluble materials remain in the lungs longer and thus provide more exposure to

those tissues. The ranking of the organ doses also varies with the inhalation class.

10.2. DOSE CONVERSION FACTORS FOR WILDLIFE

10.2.1. General assumptions

Plants and animals can be exposed to ionizing radiation from radionuclides

in the environment by both external and internal exposure. The United Nations

Scientific Committee on the Effects of Atomic Radiation (UNSCEAR)

[10.4]

reports that:

“101. Radionuclides distributed in the environment lead to external

exposure of an organism living in or close to a medium that contains

radionuclides. The external exposure of biota is the result of complex and

non-linear interactions of various factors:

189

DOSE ASSESSMENT

TABLE 10.1. DOSE COEFFICIENTS FOR INGESTION OF

210

Organ/tissue Dose coefficient (Sv/Bq)

Age at intake

(years)

<1 1–2 2–7 7–12 12–17 >17

value

1 0.5 0.5 0.5 0.5 0.5

Adrenals 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Bladder 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Bone surface 8.0 × 10

−5

2.3 × 10

−5

8.9 × 10

−6

4.8 × 10

−6

2.8 × 10

−6

1.6 × 10

−6

Brain 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Breast 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Gastrointestinal

tract

Oesophagus 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Stomach 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Small intestine 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.9 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Upper large

intestine

5.8 × 10

−6

2.1 × 10

−6

1.0 × 10

−6

6.1 × 10

−7

3.6 × 10

−7

2.9 × 10

−7

Lower large

intestine

6.0 × 10

−6

2.3 × 10

−6

1.1 × 10

−6

6.7 × 10

−7

3.9 × 10

−7

3.2 × 10

−7

Colon 5.9 × 10

−6

2.1 × 10

−6

1.1 × 10

−6

6.4 × 10

−7

3.7 × 10

−7

3.0 × 10

−7

Kidney 1.8 × 10

−4

6.2 × 10

−5

3.4 × 10

−5

2.3 × 10

−5

1.6 × 10

−5

1.3 × 10

−5

Liver 1.1 × 10

−4

4.0 × 10

−5

2.0 × 10

−5

1.3 × 10

−5

8.5 × 10

−6

6.6 × 10

−6

Muscle 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Ovaries 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Pancreas 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Red bone marrow 8.1 × 10

−5

2.6 × 10

−5

1.2 × 10

−5

6.4 × 10

−6

3.8 × 10

−6

2.6 × 10

−6

Respiratory tract

Extrathoracic

airways

5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Lungs 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Skin 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Spleen 2.2 × 10

−4

7.6 × 10

−5

4.1 × 10

−5

2.5 × 10

−5

1.6 × 10

−5

1.1 × 10

−5

Testes 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Thymus 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Thyroid 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Uterus 5.7 × 10

−6

2.0 × 10

−6

9.8 × 10

−7

5.8 × 10

−7

3.4 × 10

−7

2.8 × 10

−7

Remainder 9.3 × 10

−5

3.9 × 10

−5

2.1 × 10

−5

1.3 × 10

−5

8.2 × 10

−6

6.6 × 10

−6

Effective dose 2.6 × 10

−5

8.8 × 10

−6

4.4 × 10

−6

2.6 × 10

−6

1.6 × 10

−6

1.2 × 10

−6

Source: See Ref. [10.1].

190

CHAPTER 10

TABLE 10.2. DOSE COEFFICIENTS FOR INHALATION OF

210

Po,

INHALATION CLASS F

Organ/tissue Dose coefficient (Sv/Bq)

Age at intake

(years)

<1 1–2 2–7 7–12 12–17 >17

value

0.2 0.1 0.1 0.1 0.1 0.1

Adrenals 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Bladder 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Bone surface 2.3 × 10

−5

1.3 × 10

−5

4.5 × 10

−6

2.4 × 10

−6

1.4 × 10

−6

8.0 × 10

−7

Brain 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Breast 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Gastrointestinal

tract

Oesophagus 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Stomach 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Small intestine 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Upper large

intestine

1.7 × 10

−6

1.1 × 10

−6

5.0 × 10

−7

3.0 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Lower large

intestine

1.8 × 10

−6

1.2 × 10

−6

5.3 × 10

−7

3.2 × 10

−7

1.8 × 10

−7

1.5 × 10

−7

Colon 1.7 × 10

−6

1.1 × 10

−6

5.2 × 10

−7

3.1 × 10

−7

1.7 × 10

−7

1.6 × 10

−7

Kidney 5.1 × 10

−5

3.4 × 10

−5

1.7 × 10

−5

1.1 × 10

−5

7.7 × 10

−6

6.4 × 10

−6

Liver 3.1 × 10

−5

2.2 × 10

−5

1.0 × 10

−5

6.6 × 10

−6

4.1 × 10

−6

3.3 × 10

−6

Muscle 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Ovaries 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Pancreas 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Red bone marrow 2.3 × 10

−5

1.4 × 10

−5

6.1 × 10

−6

3.2 × 10

−6

1.8 × 10

−6

1.3 × 10

−6

Respiratory Tract

Extrathoracic

airways

1.6 × 10

−6

1.1 × 10

−6

5.0 × 10

−7

3.0 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Lungs 1.7 × 10

−6

1.1 × 10

−6

5.2 × 10

−7

3.1 × 10

−7

1.9 × 10

−7

1.6 × 10

−7

Skin 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Spleen 6.3 × 10

−5

4.2 × 10

−5

2.0 × 10

−5

1.3 × 10

−5

7.8 × 10

−6

5.5 × 10

−6

Testes 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Thymus 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Thyroid 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Uterus 1.6 × 10

−6

1.1 × 10

−6

4.9 × 10

−7

2.9 × 10

−7

1.7 × 10

−7

1.4 × 10

−7

Remainder 3.3 × 10

−5

2.2 × 10

−5

1.0 × 10

−5

6.5 × 10

−6

4.0 × 10

−6

3.3 × 10

−6

Effective dose 7.4 × 10

−6

4.8 × 10

−6

2.2 × 10

−6

1.3 × 10

−6

7.7 × 10

−7

6.1 × 10

−7

Source: See Ref. [10.3].

191

DOSE ASSESSMENT

TABLE 10.3. DOSE COEFFICIENTS FOR INHALATION OF

210

Po,

INHALATION CLASS M

Organ/tissue Dose coefficient (Sv/Bq)

Age at intake

(years)

<1 1–2 2–7 7–12 12–17 >17

value

0.2 0.1 0.1 0.1 0.1 0.1

Adrenals 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Bladder 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Bone surface 8.9 × 10

−6

4.1 × 10

−6

1.5 × 10

−6

7.9 × 10

−7

4.6 × 10

−7

2.8 × 10

−7

Brain 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Breast 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Gastrointestinal

tract

Oesophagus 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Stomach 6.5 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Small intestine 6.5 × 10

−7

3.5 × 10

−7

1.7 × 10

−7

9.8 × 10

−8

5.8 × 10

−8

5.0 × 10

−8

Upper large

intestine

7.1 × 10

−7

3.9 × 10

−7

1.8 × 10

−7

1.1 × 10

−7

6.3 × 10

−8

5.4 × 10

−8

Lower large

intestine

8.2 × 10

−7

4.7 × 10

−7

2.2 × 10

−7

1.3 × 10

−7

7.3 × 10

−8

6.2 × 10

−8

Colon 7.6 × 10

−7

4.3 × 10

−7

2.0 × 10

−7

1.2 × 10

−7

6.7 × 10

−8

5.7 × 10

−8

Kidney 2.0 × 10

−5

1.1 × 10

−5

5.7 × 10

−6

3.8 × 10

−6

2.7 × 10

−6

2.2 × 10

−6

Liver 1.2 × 10

−5

7.1 × 10

−6

3.4 × 10

−6

2.2 × 10

−6

1.4 × 10

−6

1.2 × 10

−6

Muscle 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Ovaries 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Pancreas 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Red bone marrow 9.1 × 10

−6

4.6 × 10

−6

2.1 × 10

−6

1.1 × 10

−6

6.3 × 10

−7

4.6 × 10

−7

Respiratory tract

Extrathoracic

airways

3.2 × 10

−5

2.4 × 10

−5

9.9 × 10

−6

6.6 × 10

−6

3.6 × 10

−6

3.5 × 10

−6

Lungs 1.1 × 10

−4

8.1 × 10

−5

5.1 × 10

−5

3.5 × 10

−5

3.1 × 10

−5

2.6 × 10

−5

Skin 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Spleen 2.5 × 10

−5

1.4 × 10

−5

6.8 × 10

−6

4.2 × 10

−6

2.7 × 10

−6

1.9 × 10

−6

Testes 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Thymus 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Thyroid 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Uterus 6.4 × 10

−7

3.5 × 10

−7

1.6 × 10

−7

9.7 × 10

−8

5.7 × 10

−8

4.9 × 10

−8

Remainder 1.0 × 10

−6

5.6 × 10

−7

2.7 × 10

−7

1.6 × 10

−7

1.0 × 10

−7

8.4 × 10

−8

Effective dose 1.5 × 10

−5

1.1 × 10

−5

6.7 × 10

−6

4.6 × 10

−6

4.0 × 10

−6

3.3 × 10

−6

Source: See Ref. [10.3].

192

CHAPTER 10

TABLE 10.4. DOSE COEFFICIENTS FOR INHALATION OF

210

Po,

INHALATION CLASS S

Organ/tissue Dose coefficient (Sv/Bq)

Age at intake

(years)

<1 1–2 2–7 7–12 12–17 >17

value 0.02 0.01 0.01 0.01 0.01 0.01

Adrenals 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Bladder 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.1 × 10

−9

2.9 × 10

−9

2.4 × 10

−9

Bone surface 6.1 × 10

−7

2.2 × 10

−7

7.7 × 10

−8

4.1 × 10

−8

2.3 × 10

−8

1.4 × 10

−8

Brain 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Breast 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Gastrointestinal

tract

Oesophagus 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Stomach 4.9 × 10

−8

2.1 × 10

−8

9.6 × 10

−9

5.7 × 10

−9

3.2 × 10

−9

2.7 × 10

−9

Small intestine 5.5 × 10

−8

2.6 × 10

−8

1.2 × 10

−8

7.0 × 10

−9

3.8 × 10

−9

3.2 × 10

−9

Upper large

intestine

1.1 × 10

−7

6.5 × 10

−8

2.8 × 10

−8

1.7 × 10

−8

8.6 × 10

−9

7.2 × 10

−9

Lower large

intestine

2.4 × 10

−7

1.5 × 10

−7

6.4 × 10

−8

3.9 × 10

−8

2.0 × 10

−8

1.6 × 10

−8

Colon 1.7 × 10

−7

1.0 × 10

−7

4.4 × 10

−8

2.6 × 10

−8

1.3 × 10

−8

1.1 × 10

−8

Kidney 1.4 × 10

−6

5.9 × 10

−7

2.9 × 10

−7

2.0 × 10

−7

1.3 × 10

−7

1.1 × 10

−7

Liver 8.4 × 10

−7

3.8 × 10

−7

1.7 × 10

−7

1.2 × 10

−7

7.0 × 10

−8

5.7 × 10

−8

Muscle 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Ovaries 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Pancreas 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Red bone marrow 6.2 × 10

−7

2.4 × 10

−7

1.1 × 10

−7

5.6 × 10

−8

3.1 × 10

−8

2.3 × 10

−8

Respiratory tract

Extrathoracic

airways

5.7 × 10

−5

4.4 × 10

−5

1.9 × 10

−5

1.3 × 10

−5

7.1 × 10

−6

6.9 × 10

−6

Lungs 1.5 × 10

−4

1.1 × 10

−4

7.2 × 10

−5

4.9 × 10

−5

4.3 × 10

−5

3.5 × 10

−5

Skin 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Spleen 1.7 × 10

−6

7.2 × 10

−7

3.5 × 10

−7

2.2 × 10

−7

1.3 × 10

−7

9.6 × 10

−8

Testes 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Thymus 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Thyroid 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Uterus 4.4 × 10

−8

1.9 × 10

−8

8.5 × 10

−9

5.0 × 10

−9

2.8 × 10

−9

2.4 × 10

−9

Remainder 1.0 × 10

−7

4.8 × 10

−8

2.2 × 10

−8

1.4 × 10

−8

8.3 × 10

−9

7.5 × 10

−9

Effective dose 1.8 × 10

−5

1.4 × 10

−5

8.6 × 10

−6

5.9 × 10

−6

5.1 × 10

−6

4.3 × 10

−6

Source: See Ref. [10.3].

193

DOSE ASSESSMENT

— The geometrical relation between the source of the radiation and the target;

— The activity levels of the radionuclides in the environment;

— The materials in the environment and their shielding properties;

— The radionuclide-specific decay properties characterized by the radiation

type, the energies emitted and their emission probabilities; and

— The habitat and size of the organism.”

For estimating exposures to wildlife, models have been developed for

deriving dose conversion coefficients (DCCs) for a set of plants and animals.

Since it is impossible to consider all species of flora and fauna explicitly in dose

assessment, DCCs are provided for a set of reference organisms that were selected

within the FASSET and ERICA project to represent typical members of common

ecosystems (see Refs

[10.5–10.7]). The reference organisms considered here

are summarized in Table

10.5. For calculating DCCs for internal and external

exposures, the following simplifying assumptions were made:

(a) The radionuclides are uniformly distributed within the body and the

supporting media.

(b) The shape of all organisms is approximated by spheres and ellipsoids.

10.2.2. Dose conversion factors for flora and fauna

UNSCEAR

[10.4] reports that (see also Refs [10.8–10.11]):

“103. The exposure due to radionuclides incorporated into the organism

is determined by the activity concentrations in the organism, the size of

the organism, and the type and the energy of the emitted radiation. A key

quantity for estimating internal doses is the absorbed fraction of energy

ϕ(E), which is defined as the fraction of energy emitted by a radiation

source that is absorbed within the target tissue, organ or organism. In the

simplest case, the organism is assumed to be in an infinite homogeneous

medium and to have a uniform activity concentration throughout its body.

The densities of the medium and the organism’s body are assumed to be

identical. Under these conditions, both internal (D

int

) and external (D

ext

)

dose conversion coefficients (DCCS; the DCC is defined as either the

absorbed dose or the absorbed dose rate, according to the circumstances,

per unit activity concentration of the relevant radionuclide in the organism

or medium) for monoenergetic radiation can be expressed as a function of

the absorbed fraction…:

int

= E·ϕ(E) and D

ext

= E·(1−ϕ(E)) (4)”

194

CHAPTER 10

TABLE 10.5. REFERENCE ORGANISMS DEFINED IN FASSET/ERICA,

AND REFERENCE ANIMALS AND PLANTS AS DEFINED BY THE

INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION

ERICA reference organisms (examples)

ICRP reference

organisms

Habitat

Mass

(kg)

Terrestrial environment

Detritivorous invertebrates (woodlouse) n.a.

On and in soil 1.7 × 10

−4

Insects Bee In air 5.9 × 10

−4

Lichen and bryophytes (bryophyte) n.a.

On soil 1.1 × 10

−4

Gastropods (snail) n.a.

On soil 1.4 × 10

−3

Grasses and herbs Wild grass On soil 2.6 × 10

−3

Soil invertebrates (earthworm) Earthworm In soil 5.2 × 10

−3

Amphibians (frog) Frog On soil 0.031

Bird eggs Duck egg On soil 0.050

Burrowing mammals (rat) Rat In soil 0.31

Reptiles (snake) n.a.

On soil 0.74

Wading birds (duck) Duck On soil 1.3

Large mammals (deer) Deer On soil 250

Trees Pine tree On soil 470

Shrubs n.a.

On soil —

Marine environment

Phytoplankton n.a.

In water 6.5 × 10

−11

Zooplankton n.a.

In water 6.1 × 10

−5

Sea anemones/true corals n.a.

In water 1.8 × 10

−3

Macroalgae Brown seaweed In water 6.5 × 10

−3

Benthic molluscs n.a.

In water 1.6 × 10

−2

Polychaete worms n.a.

In water 1.7 × 10

−2

Vascular plants n.a.

In water 2.6 × 10

−2

195

DOSE ASSESSMENT

TABLE 10.5. REFERENCE ORGANISMS DEFINED IN FASSET/ERICA,

AND REFERENCE ANIMALS AND PLANTS AS DEFINED BY THE

INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION (cont.)

ERICA reference organisms (examples)

ICRP reference

organisms

Habitat

Mass

(kg)

Pelagic fish n.a.

In water 0.57

Crustaceans Crab In water 0.75

Benthic fish Flat fish In water 1.3

Reptiles (marine turtle) n.a.

In water 140

Mammals n.a.

In water 180

Sea anemones/true corals (colony) n.a.

In water 200

Wading birds Duck In water 1.3

Freshwater environment

Phytoplankton n.a.

In water 2.1 × 10

−12

Zooplankton n.a.

In water 2.4 × 10

−6

Crustaceans n.a.

In water 1.6 × 10

−5

Insect larvae n.a.

In water 1.8 × 10

−5

Vascular plants n.a.

In water 1.1 × 10

−3

Gastropods n.a.

In water 3.5 × 10

−3

Bivalve molluscs n.a.

In water 7.1 × 10

−2

Pelagic fish Salmonid/trout In water 1.3

Benthic fish n.a.

In water 1.5

Mammals n.a.

In water 3.9

Birds Duck In water 1.3

Amphibians Frog In water 3.1 × 10

−2

Source: See Ref. [10.7].

n.a.: not applicable.

—: data not available.

196

CHAPTER 10

Ulanovsky et al. [10.11] find that:

“The second equation is an approximation that — in a strict sense — only

holds if the organism and the surrounding medium are of the same density

and elemental composition.”

UNSCEAR

[10.4] continues:

“104. Absorbed fractions for photon and electron sources uniformly

distributed in soft-tissue spherical bodies immersed in an infinite

water medium have been systematically calculated by Monte Carlo

simulation

[Ref. [10.10]]. The calculations covered a particle energy range

of 10

keV to 5 MeV, a range for the mass of the body from 10

−6

to 10

kg….

…….

“109. The approach was also applied to the calculation of the absorbed

fractions for non-aquatic animals and their internal exposures. With the

use of the absorbed fractions for spheres and the suggested re-scaling and

interpolation techniques, a set of internal DCCs has been calculated for all

reference animals and plants [Ref. [10.10]].”

In Ulanovsky and Pröhl [10.12], DCCs for external exposure are calculated

separately for aquatic and terrestrial reference animals and plants. For aquatic

organisms, which are immersed by water, external exposure is calculated

according to equations above.

For terrestrial reference animals and plants, the estimation of external

exposures is more complex, since soil, air and organic matter differ considerably

in composition and density [10.11, 10.13]. Therefore, the derivation of DCCs is

based on radiation transport simulated for monoenergetic photons by means of

Monte Carlo techniques [10.11, 10.13]. Owing to the complexity of the processes

and the variability of life forms, reference geometries, as defined by energy,

contaminated media, and organism size and shape, are considered in detail by

Taranenko et al. [10.13]. External exposures under conditions for which specific

calculations have not been performed, can be estimated with sufficient accuracy

by interpolation. The DCCs for

210

Po are summarized in Table 10.6 for terrestrial

organisms, and in Tables 10.7 and 10.8 for aquatic organisms.

The values for internal exposure are given in units of dose rate per unit

activity in the organisms (μGy/h per Bq/kg). The values are given for absorbed

dose (i.e. radiation weighting factors to account for the effectiveness of the

197

DOSE ASSESSMENT

TABLE 10.6. DOSE CONVERSION COEFFICIENTS FOR TERRESTRIAL

REFERENCE ORGANISMS FOR

210

Po FOR INTERNAL AND EXTERNAL

EXPOSURE

Reference organism

(example)

Internal

exposure

External

exposure

DCC

(μGy/h per

Bq/kg)

(%)

In soil/

infinite

(μGy/h per

Bq/kg )

On soil/

planar

(μGy/h per

Bq/m

)

On soil/

volume

(μGy/h per

Bq/kg)

Woodlouse 3.1 × 10

−3

100 0 0 4.6 × 10

−9

2.9 × 10

−11

1.7 × 10

−9

Bees 3.1 × 10

−3

100 0 0 —

2.9 × 10

−11

1.7 × 10

−9

Terrestrial lichens 3.1 × 10

−3

100 0 0 —

1.5 × 10

−11

9.1 × 10

−10

Terrestrial

gastropods (snail)

3.1 × 10

−3

100 0 0 4.6 × 10

−9

2.9 × 10

−11

1.7 × 10

−9

Terrestrial grasses

(wild grass)

3.1 × 10

−3

100 0 0 —

6.5 × 10

−11

1.7 × 10

−9

Soil invertebrates

(earthworm)

3.1 × 10

−3

100 0 0 4.5 × 10

−9

—

Terrestrial

amphibians (frog)

3.1 × 10

−3

100 0 0 4.5 × 10

−9

2.9 × 10

−11

1.7 × 10

−9

Bird eggs (duck

egg)

3.1 × 10

−3

100 0 0 —

2.8 × 10

−11

1.7 × 10

−9

Terrestrial

burrowing

mammals (rat)

3.1 × 10

−3

100 0 0 4.3 × 10

−9

2.8 × 10

−11

1.7 × 10

−9

Terrestrial reptiles

(snake)

3.1 × 10

−3

100 0 0 4.1 × 10

−9

2.7 × 10

−11

1.6 × 10

−9

Large terrestrial

mammals (deer)

3.1 × 10

−3

100 0 0 —

1.4 × 10

−11

8.6 × 10

−10

Trees (pine tree) 3.1 × 10

−3

100 0 0 —

2.0 × 10

−11

1.4 × 10

−9

Terrestrial shrubs —

—

3.1 × 10

−11

1.6 × 10

−9

Terrestrial wading

birds (duck)

3.1 × 10

−3

100 0 0 —

2.6 × 10

−11

1.6 × 10

−9

2.3 × 10

−11

1.5 × 10

−9

Source: See Ref. [10.13].

—: data not available.

198

CHAPTER 10

different radiation types in causing effects to biota are not applied, since the

discussion on this topic is ongoing [10.14]).

To estimate the possible impact of different radiation qualities, the

contributions of the different types of radiation to the DCCs for internal exposure

are given separately in the tables. These contributions are quantified through the

factors f

, f

and f

, respectively, where f

specifies densely ionizing radiation

(alpha particles and spontaneous fission fragments), f

represents the fraction

of low energy beta radiation (E

< 10 keV), and f

gives the contributions of

higher energy beta radiation (E

> 10 keV) and gamma radiation of all energies.

TABLE 10.7. DOSE CONVERSION COEFFICIENTS FOR MARINE

REFERENCE ORGANISMS FOR

210

Po FOR INTERNAL AND EXTERNAL

EXPOSURE

Reference organism

(example)

Internal

exposure

External

exposure

DCC

(

μGy/h per

Bq/kg)

(%)

DCC

(

μGy/h

per Bq/L)

Marine phytoplankton

2.0

× 10

−7

100 0 0

3.1

× 10

−3

Zooplankton

3.1

× 10

−3

100 0 0

4.9

× 10

−9

Sea anemones/true corals

3.1

× 10

−3

100 0 0

4.8

× 10

−9

Macroalgae (brown seaweed)

3.1

× 10

−3

100 0 0

4.9

× 10

−9

Benthic molluscs

3.1

× 10

−3

100 0 0

4.7

× 10

−9

Vascular plants

3.1

× 10

−3

100 0 0

4.7

× 10

−9

Marine worms (polychaete worm)

3.1

× 10

−3

100 0 0

4.8

× 10

−9

Marine pelagic fish

3.1

× 10

−3

100 0 0

4.4 × 10

−9

Marine crustaceans (crab)

3.1

× 10

−3

100 0 0

4.3

× 10

−9

Marine benthic fish (flatfish)

3.1

× 10

−3

100 0 0

4.5

× 10

−9

Marine reptiles

3.1

× 10

−3

100 0 0

2.3

× 10

−9

Marine mammals (dolphin)

3.1

× 10

−3

100 0 0

2.2

× 10

−9

Source: See Ref. [10.13].

199

DOSE ASSESSMENT

Consequently, any user-defined radiation weighting factors w

can be applied to

these fractions to calculate radiation weighted DCCs:

int int rr

D D wf=×

(10.1)

where D

int

represents the DCC as provided in Tables 10.6–10.8.

The discussion on the derivation of radiation weighting to be applied in

dosimetry for flora and fauna is still ongoing. Radiation weighting factors

for alpha radiation are reviewed in Ref.

[10.14]. For screening models being

applied to obtain conservative results, radiation weighting factors of 10 and 20

for deterministic and stochastic end points, respectively, were considered to be

TABLE 10.8. DOSE CONVERSION COEFFICIENTS FOR FRESHWATER

TERRESTRIAL REFERENCE ORGANISMS FOR

210

Po FOR INTERNAL

AND EXTERNAL EXPOSURE

Reference organism

(example)

Internal

exposure

External

exposure

DCC

(

μGy/h per

Bq/kg)

(%)

DCC

(

μGy/h

per Bq/L)

Freshwater phytoplankton

6.3

× 10

−9

100 0 0

3.1

× 10

−3

Freshwater zooplankton

3.1

× 10

−3

100 0 0

4.9

× 10

−9

Freshwater crustaceans

3.1

× 10

−3

100 0 0

4.9

× 10

−9

Freshwater insect larvae

3.1

× 10

−3

100 0 0

4.9

× 10

−9

Freshwater plants (vascular plant)

3.1

× 10

−3

100 0 0

4.9

× 10

−9

Freshwater gastropods

3.1

× 10

−3

100 0 0

4.8

× 10

−9

Freshwater molluscs (bivalve mollusc)

3.1

× 10

−3

100 0 0

4.6

× 10

−9

Freshwater pelagic sh (trout)

3.1

× 10

−3

100 0 0

4.3 × 10

−9

Freshwater benthic sh

3.1

× 10

−3

100 0 0

4.3

× 10

−9

Freshwater mammals (muskrat)

3.1

× 10

−3

100 0 0

3.9

× 10

−9

Amphibians (frog)

3.1

× 10

−3

100 0 0

4.7

× 10

−9

Wading birds (duck)

3.1

× 10

−3

100 0 0

4.2

× 10

−9

Source: See Ref. [10.13].

200

CHAPTER 10

appropriate to obtain conservative results. To evaluate radiological impacts on

flora and fauna, deterministic end points are relevant [10.7].

Values for external exposures for aquatic organisms are given for a total

immersion in water in units of μGy/h per Bq/L. For terrestrial organisms, various

different source geometries are considered, depending on the habitat of the

organism considered

[10.13]:

(a) For on or above soil organisms, two geometries are considered:

— A planar source, the source being at a depth of 0.3 cm to account for

surface roughness (μGy/h per Bq/m

);

— A 10 cm thick volume source (soil density: 1.6 g/cm

)(μGy/h per

Bq/kg).

(b) For in soil organisms, the values are given for organisms living in the

middle of a volume source with a thickness of 50

cm (soil density:

1.6 g/cm

) (μGy/h per Bq/kg).

The decay properties are reflected in the contributions of alpha, beta and

gamma radiation as quantified in Tables 10.6–10.8 by f

, f

and f

, respectively.

Only alpha radiation contributes to the internal exposure. Due to the short

range of alpha particles, external exposure due to

210

Po in the environment is

negligible [10.13]. Only very small organisms, with dimensions in the order of

the range of alpha particles (ca. 100 μm), can be exposed externally [10.13]. The

short range of alpha particles means that all energy emitted is absorbed locally.

Therefore, the DCCs for internal exposure are the same for all organisms, with

the exception of very small organisms, with dimensions that are in the order of

the range of alpha particles [10.13].

10.2.3. Non-homogeneous distribution

The DCCs for internal absorbed dose rate are calculated assuming a

homogeneous distribution of radionuclides within the body

[10.15]. However,

this is often not the case: for example, iodine accumulates in the thyroid,

polonium in the liver and transuranic elements in the kidney. The influence of

inhomogeneous distributions on the average whole body dose is investigated by

Gómez-Ros et al. [10.15]. In the case of

210

Po, for which only alpha radiation is

relevant for internal exposure, the exposure due to

210

Po in tissues and organs is

directly proportional to the

210

Po activity in the organ.

When a radionuclide accumulates in a specific organ, the dose will be

higher than the whole body average. The dose in the specific organ is enhanced

by a factor that can approximately be estimated from the ratio of whole body

mass to organ mass.

201

DOSE ASSESSMENT

10.2.4. Application of dose conversion coefficients

The DCCs can be used to estimate external and internal exposures for

organisms living in different habitats. The total dose is calculated as the sum of

external and internal exposure. The internal exposure of a terrestrial or aquatic

organism (D

int

) is estimated from the activity concentration in soil or water

, C

), concentration ratios (CR

soil-organism

, CR

water-organism

) and the DCC for

internal exposure according to:

Terrestrial organisms: D

int

= C

·CR

soil-organism

·DCC

int

(10.2)

Aquatic organisms: D

int

= C

·CR

water-organism

·DCC

int

(10.3)

Dose rates can also be estimated directly from measured radionuclide

concentrations in biota. External exposures are directly estimated from the

activity in relevant media and the appropriate DCC. Some organisms move

between different habitats owing to their lifestyle or in different stages of

development. This effect can be accounted for through a simple superposition of

the DCCs for isolated habitats.

Ulanovsky and Pröhl [10.12] provide a simple example: an aquatic

benthic organism is considered that lives on the interface of water and sediment.

Assuming that sediment density and composition are close to those of water, and

that the radionuclide concentrations in water C

and sediment C

are known, the

external dose to such an organism can be calculated according to:

ext

= (0.5C

+ 0.5C

) DCC

ext

(10.4)

where D

ext

is the DCC for the radionuclide in water. The factor of 0.5 is to account

for the habitat at the interface of water and sediment; the organism is exposed to

50% of the radionuclides in sediments and water, respectively.

A more complex exposure has to be considered, for example, for a wading

bird, such as a duck, that lives in three different habitats: on soil, above soil and

on the water surface

[10.12]. For assessing the total external exposure of a duck,

the time spent in these three habitats has to be known.

202

CHAPTER 10

REFERENCES TO CHAPTER 10

[10.1] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Age-dependent Doses to Members of the Public from Intake of Radionuclides —

Part

2: Ingestion Dose Coefficients, Publication 67, Pergamon Press, Oxford and

New York

(1992).

[10.2] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Age-dependent Doses to Members of the Public from Intake of Radionuclides —

Part

5: Compilation of Ingestion and Inhalation Coefficients, Publication 72,

Pergamon Press, Oxford and New York

(1995).

[10.3] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Age-dependent Doses to Members of the Public from Intake of Radionuclides —

Part

4: Inhalation Dose Coefficients, Publication 71, Pergamon Press, Oxford and

New York

(1995).

[10.4] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Sources and Effects of Ionizing Radiation, UNSCEAR 2008 Report to

the General Assembly, with Scientific Annexes, Vol. II: Effects, United Nations,

New York (2008) Annex E.

[10.5] LARSSON, C.M., An overview of the ERICA Integrated Approach to the assessment

and management of environmental risks from ionising contaminants, J. Environ.

Radioact.

99 (2008) 1364–1370.

[10.6] BROWN, J.E., et al., The ERICA Tool, J. Environ. Radioact. 99 (2008) 1371–1383.

[10.7] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Environmental Protection: The Concept and Use of Reference Animals and Plants,

Publication

108, Amsterdam, Elsevier (2008).

[10.8] NATIONAL COUNCIL ON RADIATION PROTECTION AND MEASUREMENTS,

Effects of Ionizing Radiation on Aquatic Organisms, NCRP Report No.

109, NCRP,

Bethesda, MD

(1991).

[10.9] VIVES i BATLLE, J., JONES, S.R., GÓMEZ-ROS, J.M., A method for calculation

of dose per unit concentration values for aquatic biota, J. Radiol. Prot.

(2004)

A13–A34.

[10.10] ULANOVSKY, A., PRÖHL, G., A practical method for assessment of dose

conversion coefficients for aquatic biota, Radiat. Environ. Biophys.

(2006)

203–214.

[10.11] ULANOVSKY, A., PRÖHL, G., GÓMEZ-ROS, J.M., Methods for calculating dose

conversion coefficients for terrestrial and aquatic biota, J. Environ. Radioact. 99

(2008) 1440–1448.

[10.12] ULANOVSKY, A., PRÖHL, G., Tables of dose conversion coefficients for estimating

internal and external radiation exposures to terrestrial and aquatic biota, Radiat.

Environ. Biophys. 47 (2008) 195–203.

[10.13] TARANENKO, V., PRÖHL, G., GÓMEZ-ROS, J.M., Absorbed dose rate conversion

coefficients for reference terrestrial biota for external photon and internal exposures,

J. Radiol. Prot. 24 (2004) A35–A62.

203

DOSE ASSESSMENT

[10.14] CHAMBERS, D.B., OSBORNE, R.V., GARVA, A.L., Choosing an alpha radiation

weighting factor for doses to non-human biota, J. Environ. Radioact.

87 (2006) 1–14.

[10.15] GÓMEZ-ROS, J.M., PRÖHL, G., ULANOVSKY, A., LIS, M., Uncertainties of

internal dose assessment for animals and plants due to non-homogeneously

distributed radionuclides, J. Environ. Radioact.

99 (2008) 1449–1455.

205

Appendix I

GEOCHRONOLOGICAL APPLICATIONS

OF POLONIUM

I.1. INTRODUCTION

The systematics of the decay of

210

Pb to

210

Po (through the decay of

210

Bi)

can be used to determine ages and residence times over several half-lives of

210

Po, in the order of approximately 500 days. This can be applied to situations in

which

210

Po is strongly fractionated from

210

Pb so that substantial disequilibrium

is generated.

I.2. DATING VOLCANIC ROCKS

There have only been a few examples of dating using

210

Po. The method

was first developed for, and mostly applied to, volcanic rocks, especially

submarine basalts where volcanic events are not so readily observed. In a rock

that forms incorporating

210

Pb and

210

Po in disequilibrium, the

210

Po will evolve

according to the following equation [I.1]:

( ) ( ) ( )

( )

234 234

Po Po

210 210 210

Po Po e Pb 1 e

ll--

= +-

(I.1)

For time periods substantially longer than the half-life of

210

Bi, the

ingrowth of this intermediate nuclide can be ignored. Samples of recent volcanics

are analysed for (

210

Po) soon after sample collection, and subsequent analyses

can be used to determine

210

Pb. An age can then be determined once an estimate

of (

210

Po)

has been made. This method was first adapted by Rubin et al. [I.1] to

determine the age of mid-ocean ridge basalts from a volcanically active region.

The method relies on disequilibrium within the basalt at the time of solidification

and no subsequent gains or losses of

210

Pb or

210

Po. Since

210

Po can be completely

volatilized at temperatures over 400°C, it might be expected that the samples

completely lose

210

Po during eruption. However, the retention of some

210

cannot be discounted due to incomplete degassing as basalts are quickly cooled,

especially under increased pressures at large ocean depths.

Rubin et al. [I.1] measure the (

210

Po) in each sample over two years

after collection (see Fig. I.1). The simple observation of

210

Po ingrowth

demonstrates that the samples erupted within approximately two years. Through

206

APPENDIX I

repeated analyses of each sample, best-fit ingrowth curves can be determined

which defined the maximum value of (

210

Po) (equal to (

210

Pb)), and when

(

210

Po) = 0. This approach puts a greater constraint on

210

Po ingrowth than

a simple determination of (

210

Pb) once secular equilibrium is achieved. While

the time of eruption is constrained between when (

210

Po) = 0 and the time of

collection, the actual eruption time could be between these if the basalt did not

completely lose all

210

Po and (

210

Po)

> 0. The lowest measured value of (

210

Po)

is approximately 0.25 and represents the highest possible initial value. Age

windows, corresponding to (

210

Po)

= 0–0.25, of up to several months are then

obtained [I.1]. Rubin et al. [I.2] find that

210

Po–

210

Pb dates of fresh basalts at a

ridge section are consistent with eruption during recent recorded seismic events,

and Johnson et al. [I.3] find the age of a seamount sample to be approximately

five months. Basalt from Samoa is assumed to have completely lost

210

Po during

subaerial eruption, and an age of around 260 days was determined [I.4].

REFERENCES TO APPENDIX I

[I.1] RUBIN, K.H., MACDOUGALL, J.D., PERFIT, M.R.,

210

Po–

210

Pb dating of recent

volcanic eruptions on the sea floor, Nature 368 (1994) 841–844.

[I.2] RUBIN, K.H., SMITH, M.C., PERFIT, M.R., CHRISTIE, D.M., SACKS, L.F.,

Geochronology and geochemistry of lavas from the 1996 North Gorda Ridge eruption,

Deep Sea Res. Part II 45 (1998) 2571–2597.

[I.3] JOHNSON, K.T.M., et al., Boomerang Seamount: The active expression of the

Amsterdam–St. Paul hotspot, Southeast Indian Ridge, Earth Planet. Sci. Lett. 183

(2000) 245–259.

Source: Figure 1 of Ref. [I.1].

FIG.

I.1. Polonium-210 activities per gram in two samples of mid-ocean ridge basalts from

the East Pacific Rise (reproduced with permission courtesy of Nature).

207

GEOCHRONOLOGICAL APPLICATIONS

[I.4] SIMS, K.W.W., et al.,

238

U–

230

Th–

226

Ra–

210

Pb–

210

Po,

232

Th–

228

Ra, and

235

U–

231

constraints on the ages and petrogenesis of Vailulu’u and Malumalu Lavas, Samoa,

Geochem. Geophys. Geosyst. 9 (2008).

209

Appendix II

RADIOCAESIUM,

210

Po AND

210

Pb IN WOLV ES

II.1. INTRODUCTION

Since the Chernobyl accident,

137

Cs (T

1/2

= 30.07 years) activity deposition

in Sweden has varied in the range of 1000–100 000 Bq/m

, and more than half

of this activity has now decayed. The isotope

210

Po (T

1/2

= 138 days) originates

from

222

Rn exhaled from the ground and its intermediate

210

Pb. The more even

terrestrial distribution of

210

Po depends on the uranium content of the ground,

the topography of the surrounding land mass and the amount of wet deposition.

Persson [II.1] estimated the annual deposition of

210

Pb in central Sweden in

1970 to be around 63

Bq/m

. The integrated activity of

210

Pb, supposing that the

same amount is decaying as is introduced by precipitation, gives 2000

Bq/m

The

210

Po:

210

Pb activity ratio in precipitation is around 0.1–0.2. The integrated

deposition is then slightly lower for

210

Po than for

210

Pb, since it takes about two

years for

210

Po to come into radioactive equilibrium with

210

Pb.

Polonium-210 and

137

Cs can accumulate in the food chain, and especially in

reindeer, which can contain enhanced concentrations in soft tissue after feeding

on lichen during the winter. A study by Gjelsvik et al. [II.2] published in 2014

measured activity levels of

137

Cs,

210

Po, and the

210

Po:

210

Pb ratio in different

tissues of 28 wolves in Sweden to determine the aggregated transfer factor (TF)

and to calculate doses to the animals.

II.2. MATERIALS AND METHODS

Gjelsvik et al. [II.2] examined samples of muscle, blood, liver and kidney

28 wolves killed in Sweden in 2009 and 2010 (see Table II.1). The samples for

analysis were prepared by drying and grinding, and

137

Cs was measured with

high purity germanium gamma spectrometry. The

210

Po activity was determined

with alpha spectrometry (ion implanted silicon detectors) after wet ashing and

spontaneous deposition onto silver discs. The radiochemical yield determinant

was

209

Po; and

210

Pb was analysed with dry incineration of a larger amount where

Po evaporates. After six months, the buildup

210

Po from

210

Pb was analysed.

The ingrowth rate gives the contribution of

210

Pb, and the

137

Cs and

210

concentrations were decay corrected to date of death.

210

APPENDIX II

TABLE II.1. DATA FOR SAMPLED WOLVES

Wolf County Community Sex Weight (kg) Age (years)

1 Jämtland Härjedalen F 32.1 1

2 Jämtland Härjedalen M 34.9 1

3 Dalarna Malung M 40 0

4 Dalarna Malung M 41 1

5 Dalarna Malung F 28 0

6 Dalarna Malung F 29 0

7 Örebro Degerfors M 48 1

8 Norrbotten Gällivare M 47 1

9 Jämtland Härjedalen F 32.6 1

10 Gävleborg Ockelbo M 46.5 1

11 Gävleborg Ockelbo F 34 1

12 Dalarna Malung-Sälen M 47 6

13 Dalarna Ludvika M 28 0

14 Dalarna Ludvika F 36 4

15 Dalarna Vansbro M 38 2

16 Dalarna Ludvika M —

17 Dalarna Ludvika F 36 1

18 Gävleborg Ljusdal M 49 6

19 Västra Götaland Mellerud M 49.5 3

20 Värmland Munkfors F 40 1

21 Värmland Sunne M 32

22 Värmland Eda M 28 0

23 Örebro Örebro F 37.5 1

24 Värmland Arvika F 38 3

25 Örebro Lindesberg F 29 0

26 Värmland Torsby M 48 1

27 Värmland Sunne M 40 5

28 Västra Götaland Dals-Ed F 38 5

Source: Table 1 of Ref. [II.2].

—:data not available.

211

RADIOCAESIUM,

210

Po AND

210

Pb IN WOLVES

II.3. RESULTS AND DISCUSSION

The activity concentrations of

137

Cs and

210

Po in different tissues from the

wolves are Table

II.2. The information comprises summaries from Refs [II.2–II.4].

After uptake, polonium mainly accumulates in the liver and kidneys,

while caesium is more homogenously distributed in the body (but is lower in

bone). Food is the major intake pathway for both caesium and polonium. For

210

Po, there is also in vivo buildup from

210

Pb. Reindeer have enhanced levels

of both

137

Cs and

210

Po, especially after winter due to the consumption of lichen.

However, during the summer, 74% of the wolf diet consists of moose. The

aggregated TF for

210

Po shows large variations (0.000 5–0.27) depending on the

organ (see Table II.2). Gjelsvik et al. [II.2] report that:

“The bio-kinetics of polonium in the body is rather complicated. The

fractional uptake by man has been reported from 0.1 to 0.8. The biological

residence time has been reported to be between 20 and 80

days [Ref. [II.5]].

Thomas et al. [Ref. [II.3]] have studied the transfer of

210

Po and

210

through the lichen–caribou–wolf food chain of northern Canada where

caribou was the main food item for the wolves. Their results are in good

agreement with our data. Thomas et al.

[Ref. [II.3]] reported

210

Po/

210

activity ratios of 30–56 for liver and kidney. Fuller

[Ref. [II.6]] estimated

that the average daily intake was 2.7 kg per day when eating caribou and

5.4

kg per day when eating moose.”

The fresh weight activity concentrations are: 7.5–13

Bq/kg in muscle;

145–369

Bq/kg in the liver; 142–343 Bq/kg in the kidneys; and 20–29 Bq/kg

in bone. The lower values are from a region with less

210

Po in lichen and is thus

reflected in the food chain. Gjelsvik et al. [II.2] continue that:

“Howard et al. [Ref. [II.7]] estimated the transfer coefficient F

(d kg

−1

)

210

Po from caribou to wolves as the ratio of activity concentration in

the muscle of wolf (Bq kg

−1

w.w.) divided by the daily amount intake of

caribou meat (kg w.w. d

−1

) with the activity concentration of

210

Po in the

quantity Bq kg

−1

w.w. which gave a value of 0.089 d kg

−1

w.w.”

Using the same coefficient for the Swedish wolves, the food

would contain 8–170

Bq/kg (DW). The caribou from Canada contained

30–50

Bq/kg (DW) in muscle, 500–1300 Bq/kg (DW) in the kidneys and

550–1400

Bq/kg (DW) in the liver, depending on the region [II.3]. In the muscle

of moose from Finland, sampled in March, the concentrations were 25

Bq/kg

and in reindeer, 45

Bq/kg [II.4]. Persson [II.8] reported in 1972 concentrations

212

APPENDIX II

TABLE II.2. SUMMARY OF DATA ON

137

Cs,

210

Po AND

210

Pb IN WOLVES

Matrix

137

(Bq/kg)

210

(Bq/kg)

210

Po:

210

Pb Remarks Ref.

Blood

—

17 19 Canada, Snowdrift, 1993 [II.3]

4–959 2.2–54 2.3–54 Sweden, 2010–2011 [II.2]

Bone

—

4.2–25 0.29–0.7 Canada, Baker Lake, 1993 [II.3]

—

20 0.6 Canada, Snowdrift, 1993 [II.3]

—

4.8 0.26 E. Finland, 1968 [II.4]

—

11–67 0.29–0.7 N. Finland, 1968 [II.4]

Liver

—

1 400 58 Canada, Baker Lake, 1993 [II.3]

—

551 30 Canada, Snowdrift, 1993 [II.3]

—

6.8 22 E. Finland, 1968 [II.4]

—

49–310 26–115 N. Finland, 1968 [II.4]

36–4 050 20–523 9.4–56 Sweden, 2010–2011 [II.2]

Kidney

—

1 300 58 Canada, Baker Lake, 1993 [II.3]

—

539 28 Canada, Snowdrift, 1993 [II.3]

31–3 453 24–942 1.8–7.9 Sweden, 2010–2011 [II.2]

Muscle

—

49 —

Canada, Baker Lake, 1993 [II.3]

—

29 —

Canada, Snowdrift, 1993 [II.3]

350 1.2 57 E. Finland, 1968 [II.4]

270–7 000 0.3–1.0 30–80 N. Finland, 1968 [II.4]

70–8 410 1–43 2–77 Sweden, 2010–2011 [II.2]

Aggregated transfer factor (Po·kg

−1

·m

−1

) Ref.

Liver 0.012–0.23

[II.2]

Kidney 0.001–0.27

Muscle

0.000 5–0.021

0.01–0.26

Source: Adapted from Ref. [II.2].

Bq/L.

—: data not available.

Bq/kg, fresh weight.

Bq/kg,

dry weight.

213

RADIOCAESIUM,

210

Po AND

210

Pb IN WOLVES

in the muscle of Swedish reindeer as 16–48 Bq/kg, 28–220 Bq/kg in the kidneys

and 88–148

Bq/kg in the liver. Kauranen et al. [II.4] observed in 1971 lower

concentrations in wolves, lynx and wolverine in southern Finland compared to

northern Finland

[II.2].

The gastrointestinal uptake of

210

Po is higher than

210

Pb. Consequently, the

210

Po:

210

Pb activity ratio increases with level in the food chain. In lichen, the ratio

is slightly less than 1, and in reindeer about 5 in the liver and 20 in muscle

[II.8].

Thomas et al.

[II.3] find

210

Po:

210

Pb activity ratios of 28–58 for the liver and

kidneys of wolves, which is in fair agreement with the the data in Ref. [II.2].

In the south of Sweden, moose is the major food item for wolves. The

activity concentration is generally higher in reindeer than in moose but in the

north the concentrations do not differ much. However, the data do not support the

tendency for higher concentrations in northern Sweden in contrast to the results

in Ref.

[II.4] for Finland.

Rather high concentrations of

137

Cs are found in areas with high deposition

from the Chernobyl accident, such as in Gävleborg county. However, wolves

move long distances to find their prey.

Gjelsvik et al. [II.2] report that the activity ratio of

210

Po:

137

Cs in liver is

in the range of 0.011–4.0, in muscle 0.002–0.05 and in kidneys 0.012–2.1. This

differential bio-distribution shows the much larger accumulation of

210

Po in

the liver and kidneys relative to

137

Cs. The large spread in the data is due to the

variability in

137

Cs deposition in different areas and also reflects the range over

which wolves travel to obtain prey.

The dose conversion factors in humans for

137

Cs and

210

Po are

1.2

× 10

−8

Sv/Bq and 7 × 10

−7

Sv/Bq, respectively. This is for a person with

a body weight of 70 kg and using a quality factor of 20 for

210

Po. The dose is

not applicable to animals, given the lack of specific tissue weighting factors for

non-human biota, and only the absorbed dose in grays can be calculated. The

absorbed dose depends on the biological residence time, organ distribution and,

for

137

Cs, the body weight as well (a substantial fraction of the gamma energy

from 662 keV photons escapes from the body depending on body mass). Gjelsvik

et al. [II.2] report that:

“The biological half-time of caesium in the human body is about 70 days

but is supposed to be shorter for wolves both for caesium and polonium.

Under an equilibrium situation doses can be calculated even if the biological

half-time is not known but this is not possible under a dynamic situation.

We suppose a stable condition and 35

days half-life. The maximal dose

from

137

Cs to a wolf would be 3000 μGy per year…. The maximal dose

from

210

Po to liver would be 2600 μGy per year.”

214

APPENDIX II

For estimating the aggregated TFs for

137

Cs, charts of known deposition in

Sweden from the Chernobyl accident should be used. Estimates of the aggregated

TF can be as high as 0.1 but with large variations.

II.4. SUMMARY

There are significantly high activity concentrations of caesium and

polonium in wolves, and the concentrations are higher in areas where reindeer

are a significant part of the diet. Wolves are believed to prefer organs such as

the liver and kidneys, which contain higher quantities of polonium. However,

the proportion of different organs eaten is less important for

137

Cs, since it is

more homogenously distributed in the body. A high Po:Cs ratio might indicate

that moose is a major food source. Particularly high activity concentrations of

137

Cs are found in tissues from wolves from areas with high fallout from the

Chernobyl accident. Further studies of wolves from the summer season could

yield important information on food habits.

REFERENCES TO APPENDIX II

[II.1] PERSSON, R.B.R.,

Fe,

Sr,

134

Cs,

137

Cs, and

210

Po in the Biosphere: Radiological

Health Aspects of the Environmental Contamination from Radioactive Materials in

Northern Sweden, PhD Thesis, Lund Univ.

(1970).

[II.2] GJELSVIK, R., HOLM, E., KÅLÅS, J.A., PERSSON, B., ÅSBRINK, J.,

Polonium-210 and Caesium-137 in lynx (Lynx lynx), wolverine (Gulo gulo) and wolves

(Canis lupus), J. Environ. Radioact.

138 (2014) 402–409.

[II.3] THOMAS, P.A., SHEARD, J.W., SWANSON, S., Transfer of

210

Po and

210

Pb through

the lichen–caribou–wolf food chain for northern Canada, Health Phys.

(1994)

666–677.

[II.4] KAURANEN, P., MIETTINEN, J.K., PULLIANINEN, E., Polonium-210 and

lead-210 in some terrestrial animals in Finland, Ann. Zool. Fenn. 8 (1971) 318–323.

[II.5] HENRICSSON, C.F., RANEBO, Y., HANSSON, M., RÄÄF, C.L., HOLM, E., A

biokinetic study of

209

Po in man, Sci. Total Environ. 437 (2012) 384–389.

[II.6] FULLER, T.K., Population dynamics of wolves in north-central Minnesota, Wildl.

Monogr. 105 (1989) 1–41.

[II.7] HOWARD, B.J., BERESFORD, N.A., HOVE, K., Transfer of radiocesium to

ruminants in natural and semi-natural ecosystems and appropriate countermeasures,

Health Phys. 61 (1991) 715–725.

[II.8] PERSSON, B.R., “Lead-210, polonium-210 and stable lead in the food-chain lichen,

reindeer and man”, The Natural Radiation Environment

II (Proc. 2nd Int. Symp.

Houston, 1972) (ADAMS, J.A.S., LOWDER, W.M., GESELL, T.F., Eds), United

States Department of Commerce, Springfield, VA (1972)

347–367.

215

Appendix III

210

Po IN MARINE MAMMALS

III.1. INTRODUCTION

Marine mammals encompass 125

species in 20 families that merge in

different mammalian groups [III.1]:

(a) Cetaceans (whales, dolphins and porpoises);

(b) Sirenians (manatees and dugongs);

(d) Marine and sea otters;

(e) Polar bears.

These marine mammalian groups are more closely related to one or another

terrestrial mammalian group (e.g. horses, bears or elephants) than they are to

each other. However, all marine mammals have one thing in common: they derive

virtually all of their food from the marine environment

[III.1–III.3]. Jefferson et

al. [III.1] report that:

“Marine mammals are not randomly distributed in the world’s oceans.

It has long been known, for example, that certain species are found

exclusively or primarily in waters of a particular depth, temperature range,

or oceanographic regime, and not in areas lacking one or all of these

characteristics. For most species, however, little is known of the particular

factors that cause them to be found in one area and not in another that

appears, qualitatively at least, the same.

“One major factor affecting productivity, and thus indirectly influencing the

distribution of marine mammals, is the pattern of major ocean currents.

…….

“Wherever oceanic conditions promote high nutrient content, it is likely that

some species of marine mammal will be present to exploit that richness.

Thus, the presence of marine mammals and other high order predators and

consumers in an area is related primarily to prey, and secondarily to the

water conditions supporting that productivity.”

216

APPENDIX III

Mammals do not substantially absorb chemical elements, including radionuclides,

directly from sea water, as do fish through the gill epithelium. The mammalian

diet is exclusively based on common marine species, and the large majority of

radionuclide intake takes place through ingestion and inhalation.

III.2. MARINE MAMMAL FOOD CHAINS

In general, marine mammals are at the top of marine food chains, for

example: the blue whale, the largest animal, feeds on zooplankton (krill); the

sperm whale feeds on large size pelagic cephalopods and deep-sea sharks; the

dolphins of coastal seas (Delphinus delphis and other Delphinidae) feed on

common coastal fish species such as Clupeidae and Carangidae fish; sea otters

and seals feed on crustaceans, bivalves and gastropods; and the killer whale

feeds on sea otters and penguins. In their environments, marine mammals occupy

different trophic levels (planktivorous, piscivores and carnivores) and often

occupy the top position in their food webs, but marine mammals may, in turn,

have humans, sharks and bears, as well as larger cetaceans and pinnipeds as

predators

[III.3, III.4].

Owing to their position in marine food webs and following international

conventions on enhancing the protection and conservation of marine

mammals, studies have been conducted on several species to investigate

contamination by pesticide residues, polychlorinated biphenyls, heavy metals

and other toxicants, as well as on reproductive success and health impairment

(e.g. Refs [III.5, III.6]). Several reports have also covered the accumulation of

radionuclides in marine mammals. The radiosensitivity of marine mammals is

often comparable to that of terrestrial mammals. Thus knowledge on food chain

transfer and radionuclide absorption into internal organs can be used as a proxy

for radionuclide accumulation and radiation effects in other mammals, including

humans [III.7–III.9].

III.3. BIOACCUMULATION AND BIODISTRIBUTION

Although there have been limited studies on radioactivity in marine

mammals, current knowledge includes data on

210

Po activity concentrations

for some species as well as computed whole body radiation doses from from

210

Po and several other natural and artificial radionuclides [III.10–III.13]. Most

210

Po analyses on marine mammals have been performed on samples from beach

stranded animals or on samples from by-catch netted in tuna fishing gear, which

explains the scattered nature of data, both in species and geography.

217

210

Po IN MARINE MAMMALS

In 1974, Cherry and Shannon [III.14] published a review of the few results

for

210

Po reported. Since then, more data have been published ,but they do not yet

encompass most marine mammal families and, furthermore, many reports focus

on determination of artificial radionuclides. This is the case, for example, for

reports on:

137

Cs and

239+240

Pu in harbour porpoises and seals with regard to local

radioactive waste discharges from Sellafield, the United Kingdom; temporal

trends of global radioactive fallout from nuclear weapon tests in seals from Lake

Baikal, the Russian Federation; and

137

Cs geographical distribution in oceans and

specific coastal areas (such as Antarctic seas) determined in dolphins, porpoises

and seals (see Refs [III.10, III.12, III.15–III.20]).

Less attention has been given to measuring naturally occurring radionuclides

in marine mammals, with the exception of

K, which has often been used to

account for the natural radiation background and for comparison with internal

radiation dose received from artificial radionuclides [III.10, III.13, III.18].

However, current concerns regarding radiation protection of non-human biota and

the definition of reference species for environmental radiological risk assessment

has tilted radioecologists’ research interests to increasing knowledge about the

actual radiation dose received by biota [III.7–III.9, III.21–III.23]. Furthermore,

it has been recognized that, in marine organisms at least, the radiation dose

from internally deposited alpha emitters, such as

210

Po, is much larger than that

from internally deposited gamma emitters, such as

137

Cs and other artificial

radionuclides [III.24, III.25].

The results on

210

Po in internal tissues of marine mammals show large

inter-individual variation. In the same species, and for individuals collected in

the same season and same area,

210

Po concentrations in muscle tissues vary over

relatively wide ranges. For example, in the common dolphin of the north-eastern

Atlantic Ocean,

210

Po concentrations (FW) are in the range of 9–87 Bq/kg,

averaging 56 Bq/kg [III.13]. A wide activity concentration (FW) range of

48–102 Bq/kg is also found in the South Atlantic Franciscana dolphin, averaging

66.1 Bq/kg [III.26]. Of all marine mammals analysed so far,

210

Po concentrations

in muscle fall in the range of 1–102 Bq/kg (FW) (see Table III.1).

Polonium-210 concentrations (FW) in other organs are higher than in

muscle, averaging 123

Bq/kg in the liver and 110 Bq/kg in the kidneys of the

north-eastern Atlantic common dolphin

[III.13], and are even higher in the

liver of the South Atlantic Franciscana dolphin, averaging 315

Bq/kg [III.26]

(see Table III.1).

Analysis of the data indicates that in each marine mammal species, the

highest

210

Po concentrations are consistently in the liver and kidneys, followed

by muscle. Lower concentrations are measured in mammary glands, gonads,

bone and fat tissue (see Table

III.1). In each species, the liver to muscle

210

concentration ratio is always much higher than unity and among individual

218

APPENDIX III

TABLE III.1. ACTIVITY CONCENTRATIONS OF

210

Po IN MARINE

MAMMALS

Region Species

210

(Bq/kg, FW)

Ref.

NE Atlantic

Common dolphin

Muscle

Liver

Kidney

Gonad

Fat tissue

Bone

Mammary gland

56 ± 32

123 ± 42

110 ± 49

11 ± 2

11 ± 11

4.63 ± 0.12

29.5 ± 0.9

[III.13]

Striped dolphin

Muscle

Mammary gland

42.0 ± 0.8

8.2 ± 0.2

[III.13]

Pilot whale

Muscle

Fat tissue

13.7 ± 0.3

0.70 ± 0.06

[III.13]

North Atlantic

Sperm whale

Muscle 5.0 ± 0.2

[III.25]

South Atlantic

Franciscana dolphin

Muscle

Liver

66.1 ± 18.9

315 ± 168

[III.26]

Guiana dolphin

Muscle

Liver

24.6 ± 17.2

243 ± 213

[III.26]

Clymene dolphin

Muscle 15.7 ± 1.1

[III.26]

Atlantic spotted dolphin

Muscle

Liver

37.1 ± 36.4

125.7 ± 6.6

[III.26]

Bottlenose dolphin

Muscle

Liver

5.8 ± 6.3

180.5 ± 244.0

[III.26]

219

210

Po IN MARINE MAMMALS

TABLE III.1. ACTIVITY CONCENTRATIONS OF

210

Po IN MARINE

MAMMALS (cont.)

Region Species

210

(Bq/kg, FW)

Ref.

South Atlantic

Common dolphin

Muscle

Liver

9.34 ± 0.7

54.3 ± 3.2

[III.26]

Common dolphin

Muscle

Liver

Kidney

Gonad

Fat tissue

Bone

86 ± 2

122 ± 0.4

180 ± 71

3.6 ± 0.5

6.4 ± 0.6

26 ± 0.5

[III.27]

Whale

Muscle

Liver

Kidney

Gonad

Bone

1.1 ± 0.03

18 ± 0.3

43.1 ± 0.9

1.8 ± 0.1

2.5 ± 0.2

[III.27]

Pacific

Common dolphin

Muscle 4.5

[III.28]

Sperm whale

Muscle 23.3 ± 0.4

[III.29]

Atlantic

Fin whale

Muscle

Kidney

Fat tissue

1.9 ± 0.1

9.07 ± 0.30

1.3 ± 0.1

[III.29]

Arctic

Ringed seal

Muscle

Liver

Kidney

20.9

98.7

82.7

[III.11]

Bering Sea

Pacific walrus

Muscle

Liver

Kidney

28.7 ± 17

189 ± 157

174 ± 85

[III.30]

220

APPENDIX III

dolphins, this ratio varies from 2 to 39 [III.26]. High

210

Po concentrations in the

liver are thought to reflect

210

Po absorption from recently ingested food (within

a few hours), as in terrestrial mammals. From human and animal data, it can

be inferred that one week after intake the liver retains, on average, 30% of the

210

Po absorbed into the blood, while 8–10% is retained in the kidneys, 9–10%

TABLE III.1. ACTIVITY CONCENTRATIONS OF

210

Po IN MARINE

MAMMALS (cont.)

Region Species

210

(Bq/kg, FW)

Ref.

Bering Sea

Bearded seal

Muscle

Liver

Kidney

27.1 ± 2.3

207 ± 30.4

128 ± 20.5

[III.30]

Swedish coast

Grey and harbour seals

Muscle

Liver

Kidney

Gonad

2.5–20

6–63

5–65

0.5–1.8

[III.19]

Greenland, Bylot

Sound

Seal

Muscle

Liver

8.1 ± 0.8

43 ± 4

[III.31]

Antarctica

Leopard seal

Muscle

Liver

Kidney

1.6 ± 0.1

11.5 ± 0.3

15.8 ± 0.4

[III.15]

Crabeater seal

Muscle

Liver

Kidney

1.3 ± 0.2

22.5 ± 3.3

11.4 ± 2.4

[III.15]

Weddell seal

Muscle

Liver

Kidney

5.0 ± 2.5

48.8 ± 22.5

33.0 ± 11.7

[III.15]

Note: The arithmetic mean and standard deviation are recalculated from the original data.

Converted to fresh weight using dry weight to fresh weight conversion factors determined

in dolphins, as follows: muscle: 0.31; liver: 0.25; kidney: 0.22; gonad: 0.22; fat tissue: 0.64;

bone: 0.47; mammary gland: 0.24 [III.13].

221

210

Po IN MARINE MAMMALS

in blood and 5–6% in the skeleton [III.32]. Polonium-210 in the muscle and in

systemic circulation is mostly derived from ingestion as well. Bone is the long

term site for deposition of

210

Pb, which decays into

210

Po. Concentrations of

210

and

210

Po in dolphin bone were measured in near radioactive equilibrium [III.13].

It is interesting to note that radionuclide distribution modelling in the ringed seal

concluded that probably not more than 1% of the

210

Po in the ringed seal soft

tissue would be from ingrown

210

Po derived from decay of

210

Pb deposit in the

bone

[III.11].

Polonium-210 activity concentration determined in the dorsal muscle of

dolphins is always sampled in the same muscle and body zone, and is therefore

comparable. It shows an increase with dolphin body length and thus with

age

[III.13], which suggests a steady increase of

210

Po body burden, probably due

to the higher ingestion rate by larger individuals and increasing contribution of

210

Po from

210

Pb in the bone as the animal ages, such as observed in humans (see

Chapter

9). In dolphins and whales, other authors have reported an allometric

variation in

210

Po concentration, with lower

210

Po in larger animals [III.27].

However, body size is not the only parameter which influences

210

concentration in marine organisms. A relationship was also identified between

high

210

Po concentrations in muscle and muscle tissue myoglobin content,

ranging from low in white muscle to high in red muscle

[III.27]. This is likely

due to a strong association between

210

Po and metal binding proteins (such as

found with ferritin in fish), contributing to enhancing

210

Po accumulation in red

muscle tissue compared to white muscle tissue

[III.33].

Among individuals of the same species, such as observed in the common

dolphin, the variable

210

Po concentrations in muscle and other internal organs

might be dependent on diet composition and

210

Po assimilation efficiency from

recently ingested food. This receives strong support from previous studies

in other marine species, such as fish and shrimp

[III.34, III.35], as well as in

humans

[III.36].

The diet of marine mammals is not constant every day and changes over

the lifespan of the species and to vary from one feeding ground to another. In

spite of this,

210

Po ingestion is still the main source of

210

Po accumulated in

internal organs. Therefore, the differences in average

210

Po concentrations

between species might reflect the different diets of marine mammals. With

this assumption in mind,

210

Po concentrations (FW) in muscle tissue are:

10–100

Bq/kg in dolphins; 2.5–20 Bq/kg in seals off the Swedish coast;

Bq/kg in sperm whales; and as low as 1.1 Bq/kg in South Atlantic whales

(see Table III.1). All of these mammals have different food chains which could

assist data interpretation. However, data interpretation is difficult because

samples have origins in different oceanic regions and ecosystems, and were

obtained at different times. This is important because whales, for example, starve

222

APPENDIX III

during migration. Furthermore, in most cases the exact muscle sampled was not

recorded or described, and

210

Po results for each species have different statistical

weight, in many cases being represented by only a single sample.

III.4. REGIONAL DIFFERENCES

Consistent datasets for the same ecosystem are even more scant. For

the Antarctic Ocean food chain, however, Roos et al.

[III.15] report

210

concentrations in muscle tissue for single individuals of the Weddell seal, the

crabeater seal and the leopard seal of 16, 4.1 and 5.0

Bq/kg (FW), respectively. In

the Antarctic food web, the crabeater seal specializes in krill and other crustaceans;

the leopard seal eats krill, fish and, on occasion, crabeater seals; and the Weddell

seal feeds on fish, squid, krill and, infrequently, penguins

[III.3]. While krill

and other crustaceans generally display relatively higher

210

Po concentrations,

penguins were reported to display relatively low

210

Po concentration in their

flesh, 1.5–2.2

Bq/kg (FW) [III.15]. Therefore, the different

210

Po levels in these

three mammal species seem to reasonably mirror their feeding habits.

Conversely,

210

Po concentrations could be expected to be similar in the

same or related species from different oceanic regions when they occupy the same

ecological niche and trophic level. This was the case for dolphins in the tropical

South Atlantic and temperate north-eastern Atlantic Ocean

[III.13, III.26].

Globally,

210

Po transfer in marine food chains seems to be associated with

protein transfer and

210

Po concentration in consumer tissues mostly reflects

210

levels in the prey and thus the trophic level occupied by marine species [III.37].

Although

210

Po concentrations in mammals are also expected to reflect

210

Po food

chain transfer and their trophic level, the low number of marine mammal species

and specimens analysed so far is not sufficient to draw clear conclusions.

While polonium concentrations in marine mammal tissues are highly

variable and depend on absorption from food,

K concentrations determined

in tissues of porpoises, bottlenose dolphins and seals do not show wide

inter-individual variation within the same species

[III.10, III.13, III.20]. This

is consistent with the homeostatic control of

K concentration in the internal

organs of mammals.

The generally higher

210

Po concentrations in marine mammal tissue

compared to terrestrial mammals, such as muscle tissue from cows, lambs

and rabbits (see Chapter

6), is a direct consequence of the

210

Po concentration

in the diet. While the diet of marine mammals is exclusively based on marine

prey that consistently have

210

Po:

210

Pb activity ratios well above unity, the

diet of terrestrial mammals (which can be herbivorous, omnivorous or strictly

carnivorous) is generally lower in

210

Po than in

210

Pb, at least for herbivorous

223

210

Po IN MARINE MAMMALS

and omnivorous species. Similarly, some human populations (e.g. in Japan

and Portugal) have a diet rich in marine food (

210

Po:

210

Pb > 1 in the diet), with

average

210

Po intake substantially higher than for populations with a continental

diet (

210

Po:

210

Pb < 1) [III.38, III.39]. These divergent radionuclide intakes are

likely to be reflected in

210

Po concentrations in human tissues and whole body

burdens; however, this has not yet been evaluated (see Chapter

9).

Radiation doses due to natural (

210

Pb,

210

Po and

K) and artificial

(

137

Cs) radionuclides accumulated in internal organs have been computed for

some marine mammals. Due to its alpha radiation emission,

210

Po is the main

contributor to the absorbed radiation dose and is estimated to account for more

than 97% of the dose in dolphins [III.13, III.26,

III.40].

In 1982,

210

Po activity concentrations in dolphin muscle tissue mostly

fell in the range of 10–100

Bq/kg (FW), and are two to three orders of

magnitude higher than the average activity concentration in tissues of the

human body, 0.2

Bq/kg (FW) [III.41]. From internally accumulated

210

Po, the

whole body radiation dose for dolphins is around 1.5

μGy/h and 1.9 μGy/h

for Delphinus delphis and Pontoporia blainvillei, respectively. Such natural

radiation doses received by these marine mammals (~15

mGy/a) are unusually

high compared to humans who receive an average dose from internal

210

Po of

about 6.2

× 10

−4

μGy/h [III.26, III.37]. Nevertheless, such dose rates in marine

mammals are still below the threshold radiation dose rate of 100 μGy/h, at which

biological effects have been observed in other mammals [III.9, III.42].

Further study of

210

Po in marine mammals would be of great value, in

particular to understand how they cope with relatively high radiation doses and

to identify detoxification and repair mechanisms for genetic damage caused by

radiation in internal tissues.

REFERENCES TO APPENDIX III

[III.1] JEFFERSON, T.A., LEATHERWOOD, S., WEBBER, M.A., Marine Mammals of

the World, Food and Agriculture Organization of the United Nations, Rome

(1993).

[III.2] CULIK, B.M., Review of Small Cetaceans: Distribution, Behaviour, Migration and

Threats, Marine Mammal Action Plan/Regional Seas Reports and Studies No.

177,

United Nations Environment Programme/Secretariat of the Convention on the

Conservation of Migratory Species of Wild Animals, Bonn

(2004).

[III.3] NATIONAL OCEANIC AND ATOMOSPHERIC ADMINISTRATION, Marine

Mammals,

www.nmfs.noaa.gov/pr/species/mammals

[III.4] PERRIN, W.F., WÜRSIG, B., THEWISSEN, J.G.M. (Eds), Encyclopaedia of Marine

Mammals, Academic Press, San Diego, CA

(2009).

224

APPENDIX III

[III.5] DUIGNAN, P.J., Disease Investigations in Stranded Marine Mammals, 1999–2002,

DOC Science Internal Series

104, New Zealand Department of Conservation,

Wellington (2003).

[III.6] BOSSART, G.D., Marine mammals as sentinel species for oceans and human health,

Oceanogr. 19 (2006) 134–137.

[III.7] PRÖHL, G. (Ed.), Dosimetric Models and Data for Assessing Radiation Exposures to

Biota, FASSET Deliverable 3, Norwegian Radiation Protection Authority,

Østerås (2003).

[III.8] BROWN, J.E., STRAND, P., HOSSEINI, A., BØRRETZEN, P. (Eds), Handbook for

Assessment of the Exposure of Biota to Ionising Radiation from Radionuclides in the

Environment, FASSET Deliverable 5 (2003).

[III.9] BROWN, J.E., BØRRETZEN, P., HOSSEINI, A., Biological transfer of radionuclides

in marine environments: Identifying and filling knowledge gaps for environmental

impact assessments, Radioprot. 40 Suppl. 1 (2005) S533–S539.

[III.10] CALMET, D., WOODHEAD, D., ANDRÉ, J.M.,

210

Pb,

137

Cs and

K in three species

of porpoises caught in the Eastern Tropical Pacific Ocean, J. Environ. Radioact.

(1992)

153–169.

[III.11] GWYNN, J.P., BROWN, J.E., KOVACS, K.M., LYDERSEN, C., The derivation of

radionuclide transfer parameters for and dose-rates to adult ringed seal (Phoca

hispida) in an Arctic environment, J. Environ. Radioact.

90 (2006) 197–209.

[III.12] UDAKA, M., et al., Radionuclide (

137

Cs and

K) concentrations in the muscle of

Baikal seal (Pusa sibirica) from Lake Baikal, Mar. Pollut. Bull.

58 (2009) 290–294.

[III.13] MALTA, M., CARVALHO, F.P., Radionuclides in marine mammals off the

Portuguese coast, J. Environ. Radioact.

102 (2011) 473–478.

[III.14] CHERRY, R.D., SHANNON, L.V., The alpha radioactivity of marine organisms,

Atom. Energ. Rev.

12 (1974) 3–45.

[III.15] ROOS, P., HOLM, E., PERSSON, R.B.R., AARKROG, A., NIELSEN, S.P., “Flux of

210

Pb,

137

Cs,

239+240

Pu and

241

Am in the Antarctic Peninsula area”, Det Sjette Nordiske

Radioøkologi Seminar (6th Nordic Sem. Radioecology, Torshavn, Faroe Islands,

1992), Forskningscenter Risø, Roskilde (1992)

1–12.

[III.16] BERROW, S.D., et al., Radionuclides (

137

Cs and

K) in harbour porpoises Phocoena

phocoena from British and Irish coastal waters, Mar. Pollut. Bull.

36 (1998) 569–576.

[III.17] WATSON, W.S., et al., Radionuclides in seals and porpoises in the coastal waters

around the UK, Sci. Total Environ.

234 (1999) 1–13.

[III.18] YOSHITOME, R., et al., Global distribution of radionuclides (

137

Cs and

K) in

marine mammals, Environ. Sci. Technol.

37 (2003) 4597–4602.

[III.19] HOLM, E., et al., “Radioactivity in seals from the Swedish coast”, Scientific Trends

in Radiological Protection of the Environment: ECORAD 2004 (BRÉCHIGNAC, F.,

HOWARD, B.J., Eds), Institut de radioprotection et de sûreté nucléaire, Fontenay-

aux-Roses (2005) 85–95.

[III.20] NIELSEN, S.P., JOENSEN, H.P., Recent trends of environmental radioactivity in

Greenland and the Faroe Islands, Radioprot. 44 (2009) 843–848.

225

210

Po IN MARINE MAMMALS

[III.21] PENTREATH, R.J., WOODHEAD, D.S., A system for protecting the environment

from ionizing radiation: Selecting reference fauna and flora, and possible dose

models and environmental geometrics that could be applied to them, Sci. Total

Environ.

277 (2001) 33–43.

[III.22] BRÉCHIGNAC, F., Protection of the environment: How to position radioprotection

in an ecological risk assessment perspective, Sci. Total Environ.

307 (2003) 35–54.

[III.23] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Environmental Protection: The Concept and Use of Reference Animals and Plants,

Publication

108, Elsevier (2008).

[III.24] AARKROG, A., et al., A comparison of doses from

137

Cs and

210

Po in marine food: A

major international study, J. Environ. Radioact.

34 (1997) 69–90.

[III.25] CARVALHO, F.P., OLIVEIRA, J.M., “Radioactivity in marine organisms from

Northeast Atlantic Ocean”, The Natural Radiation Environment (NRE VIII) (Proc.

8th Int. Symp. Rio de Janeiro, 2007) (PASCHOA, A.S., STEINHAUSER, F., Eds),

American Institute of Physics, Melville, NY (2008)

387–392.

[III.26] GODOY, J.M., SICILIANO, S., DE CARVALHO, Z.L., DE MOURA, J.F., GODOY,

M.L.D.P.,

210

Polonium content of small cetaceans from Southeastern Brazil, J.

Environ. Radioact.

106 (2012) 35–39.

[III.27] CHERRY, R.D., HEYRAUD, M., RINDFUSS, R., Polonium-210 in teleost fish and

in marine mammals: Interfamily differences and a possible association between

polonium-210 and red muscle content, J. Environ. Radioact.

24 (1994) 273–291.

[III.28] FOLSOM, T.R., WONG, K.M., HODGE, V.F., “Some extreme accumulations of

natural polonium radioactivity observed in marine organisms”, The Natural Radiation

Environment (Proc. 2nd Int. Symp. Houston, 1972) (ADAMS, J.A.S., LOWDER,

W.M., GESELL, T.F., Eds), United States Department of Commerce, Springfield, VA

(1972) 863–882.

[III.29] HOLTZMAN, R.B., Natural levels of lead-210, polonium-210 and radium-226 in

humans and biota of the Arctic, Nature

210 (1966) 1094–1097.

[III.30] HAMILTON, T., SEAGARS, D., JOKELA, T., LAYTON, D.,

137

Cs and

210

Po in

Pacific walrus and bearded seal from St.

Lawrence Island, Alaska, Mar. Pollut.

Bull.

56 (2008) 1158–1167.

[III.31] NIELSEN, S.P., ROOS, P., Radionuclides in Seal Flesh and Liver, Summary Sem.

within the NKS B Programme, 2002–2005 (2006).

[III.32] HARRISON, J., LEGGER, R., LLOYD, D., PHIPPS, A., SCOTT, B., Polonium-210

as a poison, J. Radiol. Prot. 27 (2007) 17–40.

[III.33] DURAND, J.P., et al.,

210

Po binding to metallothioneins and ferritin in the liver of

teleost marine fish, Mar. Ecol. Prog. Ser.

177 (1999) 189–196.

[III.34] CARVALHO, F.P., FOWLER, S.W., An experimental study on the bioaccumulation

and turnover of polonium-210 and lead-210 in marine shrimp, Mar. Ecol. Prog.

Ser.

102 (1993) 125–133.

[III.35] CARVALHO, F.P., FOWLER, S.W., A double tracer technique to determine the

relative importance of water and food as sources of polonium-210 to marine prawns

and fish, Mar. Ecol. Prog. Ser.

103 (1994) 251–264.

226

APPENDIX III

[III.36] HUNT, G.J., ALLINGTON, D.J., Absorption of environmental polonium-210 by the

human gut, J. Radiol. Prot.

13 (1993) 119–126.

[III.37] CARVALHO, F.P., Polonium (

210

Po) and lead (

210

Pb) in marine organisms and their

transfer in marine food chains, J. Environ. Radioact.

102 (2011) 462–72.

[III.38] CARVALHO, F.P.,

210

Po and

210

Pb intake by the Portuguese population: The

contribution of seafood in the dietary intake of

210

Po and

210

Pb, Health Phys. 69

(1995)

469–480.

[III.39] YAMAMOTO, M., SAKAGUCHI, A., TOMITA, J., IMANKA, T., SIRAICHI, K.,

Measurements of

210

Po and

210

Pb in total diet samples: Estimate of dietary intakes of

210

Po and

210

Pb for Japanese, J. Radioanal. Nucl. Chem. 279 (2009) 93–103.

[III.40] BROWN, J.E., JONES, S.R., SAXÉN, R., THØRRING, H., VIVES i BATLLE, J.,

Radiation doses to aquatic organisms from natural radionuclides, J. Radiol. Prot.

(2004)

A63–A77.

[III.41] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Ionizing Radiation: Sources and Biological Effects, UNSCEAR 1982

Report to the General Assembly, with Annexes, United Nations, New

York (1982).

[III.42] REAL, A., SUNDELL-BERGMAN, S., KNOWLES, J.F., WOODHEAD, D.S.,

ZINGER, I., Effects of ionising radiation exposure on plants, fish and animals:

Relevant data for environmental radiation protection, J. Radiol. Prot.

(2004)

A123–A138.

227

Appendix IV

POLONIUM RELEASES FROM

VEGETATION AND FOREST FIRES

IV.1. INTRODUCTION

Humans absorb

210

Po much more than other heavy radioactive elements,

such as thorium and uranium [IV.1]. In the human body, the main origin of

210

is in food and water; however, it can also be accumulated from the air through

inhalation and lung absorption [IV.2, IV.3] (see also Chapter 9). Early studies

from the 1970s on

210

Pb and

210

Po activity concentrations in atmospheric aerosols

reported the occurrence of higher than expected

210

Po activities and abnormally

high

210

Po:

210

Pb activity ratios (i.e. >1) [IV.4, IV.5]. Those reports often attributed

these exceptional ratios to volcanic emissions (see Ref. [IV.5]). Enhanced

210

concentrations and

210

Po:

210

Pb activity ratios were also attributed to releases from

industrial sources and vegetation fires, but these were usually reported with scarce

experimental evidence in support of that attribution

[IV.4, IV.6]. Later work

measuring

210

Pb and

210

Po in smoke from savannah fires in Africa showed that

smoke particles can contain high

210

Pb and

210

Po levels and increase the usually

low activity concentration ratios in surface air aerosols in the region

[IV.7, IV.8].

The determination of radionuclides in effluent releases to the atmosphere

(gases and particulate matter) from non-nuclear industries also reveal enhanced

concentrations of

210

Po and occasionally high

210

Po:

210

Pb ratios. These are

reported, for example, for industries such as coal fired power plants, metal

smelters and elemental phosphorus producing plants

[IV.9]. High

210

concentrations and other naturally occurring radioactive material (NORM)

in effluents from these industries are generally associated with relatively high

temperatures used in their core processes.

In 1982, the United Nations Scientific Committee on the Effects of

Atomic Radiation compiled what was known about radioactive emissions from

NORM industries, pointing out that substantial enrichment of natural radioactive

elements with low boiling point could be observed within the wastes generated

in some of those industries

[IV.2]. Later, reports were provided of the existence

of high activity deposits and scales of certain volatile radioisotopes, particularly

210

Po, in filters, ash and on the surfaces of the furnaces and subsequent devices

where the processes take place

[IV.10–IV.13]. Mora et al. [IV.13] attempt to

model the behaviour of

210

Po in these industries was made based on its physical

properties, particularly its low melting point.

228

APPENDIX IV

Morawska [IV.14] investigates burning wood in stoves for house heating

as a source of contaminants to indoor air. The

210

Po concentrations are found to

be enhanced in smoke from domestic wood combustion and to contribute to poor

indoor air quality; although in modern houses the smoke is efficiently extracted

to the atmosphere. Nonetheless, measurable concentrations of

210

Po are reported

in association with soot and organic compounds in the extracted smoke

[IV.15].

Enhanced

210

Po levels and elevated

210

Po:

210

Pb ratios in aerosols in the atmosphere

of Seoul, Republic of Korea, are associated with coal burning

[IV.10].

High concentrations of

210

Po in cigarette smoke and enhanced radiation

doses to the smokers’ lung have been reported. Although tobacco leaves have

moderate

210

Po concentrations,

210

Po volatilization and re-concentration

in smoke particles are associated with the high temperature of tobacco

combustion

[IV.16–IV.19].

Altogether, evidence from investigations of NORM industries, wood

combustion and cigarette smoke case studies has sufficiently underpinned

the case for enhanced

210

Po activity concentrations occurring in particulates

formed during the transformation of common raw materials that undergo high

temperature processes. Despite this knowledge, assessment of

210

Po and other

natural radionuclides released by forest and bush fires has been essentially

under-investigated in atmospheric radioactivity research.

IV.2. VEGETATION AND FOREST FIRES

Wildfires burn many thousands of hectares of forest and bush around the

world every year. Although global statistics of vegetation fires are far from

complete, analysis of satellite data for 2000 revealed that 350

million ha were

affected by vegetation fires

[IV.20]. The plant biomass burned annually can

vary according to the ecosystem affected, but reaches many millions of tonnes.

This combustion releases considerable amounts of carbon, particles and gases

(including toxic organic substances) to the atmosphere

[IV.21–IV.23]. More

than 200 volatile substances have already been identified in forest fire smoke,

including toxic gases, such as carbon monoxide, formaldehyde, acrolein,

hydrocarbon compounds and dioxins

[IV.15, IV.23, IV.24]. Smoke particles also

adversely affect the respiratory tract and their inhalation in areas impacted by

forest fires have been connected to a higher frequency of asthma and other lung

diseases

[IV.23]. Furthermore, smoke rapidly travels around the globe and raises

concerns about air quality in areas far from forest fires

[IV.25].

Compared with what is known about the environmental stressors described

in Section

IV.1, the release of radionuclides by vegetation fires has been less

investigated and attention has been focused almost exclusively on artificial

229

VEGETATION AND FOREST FIRES

radionuclides. One example is the work undertaken to evaluate resuspension

of artificial radioactivity, such as

137

Cs, released by the Chernobyl accident, in

1986, and deposited onto soils and forests in Eastern Europe

[IV.26–IV.31]. This

resuspension was observed recently, following summer forest fires in Belarus

and Ukraine. These forest fires reinitiated atmospheric transport processes to

carry

137

Cs to central, northern and western European countries, although on a

much smaller scale than in 1986

[IV.1, IV.32, IV.33].

However, besides radioactivity of artificial origin, vegetation fires are

also likely to release naturally occurring radionuclides, such as

C and

K, and

members of the uranium, thorium and actinium decay series

[IV.1]. Plants absorb

naturally occurring radionuclides through root and foliar uptake, and accumulate

them throughout their tissues (see Chapter

6). So far, the radiological impact of

smoke from forest or grass fires has scarcely been investigated

[IV.32, IV.34].

Nevertheless, work done on

137

Cs resuspension by forest fires in Eastern Europe

allows for comparison of measurements of naturally occurring radionuclides

in surface air in Finland. The air activity concentrations of resuspended

137

are of little radiological significance and are comparable to, or lower than,

those of naturally occurring radionuclides in aerosol particles

[IV.32]. Artificial

transuranic radionuclides and total alpha radioactivity associated with smoke

from prescribed fires have also been measured at the Savannah River Site and

other woodland areas in the southern United States of America

[IV.34].

With fire, natural radionuclides concentrated in plants are released to

the atmosphere as smoke components. It is known that concentrations of

naturally occurring radionuclides are generally low in vegetation. Atmospheric

radionuclides, such as the long lived

222

Rn progeny,

210

Pb,

210

Bi and

210

Po, deposit

onto the surface of leaves and bark, where the highest

210

Po concentrations in

plants are usually measured. In the tree trunk and branch wood,

210

Po content is

usually lower than in leaves

[IV.35]. The burning of this wood still releases some

210

Po along with other compounds [IV.15]. More significant than combustion

of tree wood are the

210

Po releases by grass and bush vegetation fires to the

atmosphere (see Table

IV.1).

Carvalho et al. [IV.1] hypothesize that the release of radionuclides from

plant biomass can occur through two mechanisms:

“One consists in the volatilization of elements [such as polonium] when the

flame temperature exceeds the volatilization point of those elements. The

other is the enrichment through enhancement of element concentrations in

ash/smoke particles (refractory fraction) due to reduction of plant material

mass with the volatilization of water, organic compounds, and elements of

low volatilization point.”

230

APPENDIX IV

In contrast to the more refractory elements, such as uranium and thorium,

Carvalho et al. [IV.1] report that polonium is volatile at lower temperatures and

can be released to the atmosphere as gaseous compounds.

More research has recently been conducted on

210

Po and other natural

radionuclides released from vegetation by wildfires [IV.24, IV.35]. Carvalho et

al. [IV.1] report that in aerosols sampled in the vicinity of fires, “the radionuclide

activities per unit of air volume was systematically higher with higher particle

load in the air, suggesting that smoke particles collected on filters were the

vehicle for transport of radioactivity in the atmosphere” (see Fig. IV.1). The

210

activity concentrations in smoke particles were enriched by a factor of 100–1000

over those in the vegetation on a mass basis [IV.24]. Furthermore, Carvalho et

al. [IV.1] find (see also Refs [IV.35, IV.36]):

“The proportions of radionuclides in the smoke were modified in comparison

with background aerosols and with unburned plants. For example, the

concentration ratios

210

Po/

210

Pb in vegetation before the fire were often

about 0.1, while in the smoke from vegetation fires this ratio increased up

to 12, revealing an extraordinary enrichment of smoke particles in

210

Po.”

Paatero et al.

[IV.32] report that the

210

Po:

210

Pb activity ratio in air increased

by almost an order of magnitude when smoke particulates were present. Carvalho

et al. [IV.36] determine

210

Po in smoke from vegetation fires was determined

in several particle size classes of smoke aerosols sampled with a cascade

impactor

[IV.36]. This reveals an enrichment of the activity concentration of

TABLE IV.1. RADIONUCLIDE ACTIVITY CONCENTRATIONS IN PLANTS

AND IN SMOKE FROM PLANT COMBUSTION (Bq/kg, DW)

238

234

226

210

232

Leaves —

—

9.4 1.05 —

Trunk —

—

4.3 0.14 —

Smoke of:

Eucalyptus wood

Acacia wood

Bush wildfire

Pinewood wildfire

65 ± 5

11 ± 1

169 ± 7

204 ± 11

88 ± 7

36 ± 4

167 ± 7

217 ± 12

318 ± 65

503 ± 165

2492 ± 639

3547 ± 1678

52.2 ± 4

440 ± 19

369 ± 36

—

125 ± 10

611 ± 28

4422 ± 186

3972 ± 157

35 ± 3

<29

—

117 ± 8

Source: Table 1 of Ref. [IV.1] (see also Refs [IV.24, IV.35]).

—: data not available.

231

VEGETATION AND FOREST FIRES

radionuclides in aerosol particles collected near the vegetation fire compared to

reference aerosol samples collected in other areas in the absence of vegetation

fires

[IV.1]. In the case of

210

Po, this enrichment is observed mostly in the

smallest particles, with aerodynamic diameters less than 0.5

μm. In contrast,

the distribution of refractory elements, such as

232

Th and

238

U, in smoke particle

aerosols across the range of size classes show reasonable homogeneity

[IV.1].

Owing to the much higher number of smaller particles (<0.5

μm) per unit

volume, these are the main contributors for inhaled activity, particularly for

210

(see Fig. IV.2.)

Polonium atoms in the gas phase are positive ions and can easily be

recaptured by electrostatic charges onto the aerosol particles, and smaller

particles (coagulation nuclei) can gradually enter accumulation mode and later

settle as coarse aerosol particles

[IV.14]. Considering the

210

Po radionuclide

distribution in aerosol fractions, inhalation of smoke in the vicinity of a

vegetation fire can lead to inhalation of significant activity. Inhalation of up to

mBq/m

, mostly due to the

210

Po fraction associated with very fine particles,

has been estimated

[IV.35]. Prolonged exposure to smoke from vegetation fires

Source: Figure 1 of Ref. [IV.24].

Note: Correlations are statistically significant at p < 0.01.

FIG.

IV.1. Aerosol load in surface air with smoke from vegetation fires, and

210

Pb and

210

radioactivity in the air (reproduced with permission courtesy of EIT Press).

232

APPENDIX IV

might lead to enhanced radiation doses to the lung, especially from

210

Po, that

could be comparable or even higher than the radiation dose received by heavy

cigarette smokers. This exposure to ionizing radiation is sufficient justification

for recommending respiratory protection, in particular for firefighters and

members of the public in areas heavily affected by smoke from vegetation and

forest fires

[IV.35, IV.36]. Increasing worldwide consumption of coal is also

likely to enhance

210

Po levels in the atmosphere and its impact on public health

should be carefully monitored

[IV.10].

REFERENCES TO APPENDIX IV

[IV.1] CARVALHO, F.P., OLIVEIRA, J.M., MALTA, M., “Exposure to forest fires,

radioactivity and health risks”, Occupational Safety and Hygiene (SHO 2012) (Int.

Symp. Guimarães, 2012) (AREZES, P., Eds.), Portuguese Society of Occupational

Safety and Hygiene, Guimarães (2012) 126–130.

[IV.2] UNITED NATIONS SCIENTIFIC COMMITTEE ON THE EFFECTS OF ATOMIC

RADIATION, Ionizing Radiation: Sources and Biological Effects, UNSCEAR 1982

Report to the General Assembly, with Annexes, United Nations, New York (1982).

Source: Figure 2 of Ref. [IV.36].

FIG.

IV.2. Radionuclide concentrations in aerosol particles in the smoke from a forest fire

sorted by aerodynamic size classes (AS) (reproduced with permission courtesy of Elsevier).

233

VEGETATION AND FOREST FIRES

[IV.3] CARVALHO, F.P.,

210

Po and

210

Pb intake by the Portuguese population: The

contribution of seafood in the dietary intake of

210

Po and

210

Pb, Health Phys. 69

(1995) 469–480.

[IV.4] POET, S.E., MOORE, H.E., MARTELL, E.A., Lead 210, bismuth 210, and polonium

210 in the atmosphere: Accurate ratio measurement and application to aerosol

residence time determination, J. Geophys. Res.

77 (1972) 6515–6527.

[IV.5] LAMBERT, G., BUISSON, A., SANAK, J., ARDOUIN, B., Modification of the

atmospheric polonium 210 to lead 210 ratio by volcanic emissions, J. Geophys.

Res.

84 (1979) 6980–6986.

[IV.6] CARVALHO, F.P., Origins and concentrations of

222

Rn,

210

Pb,

210

Bi and

210

Po in the

surface air at Lisbon, Portugal, at the Atlantic edge of the European continental

landmass, Atmos. Environ.

29 (1995) 1809–1819.

[IV.7] NHO, E.-Y., ARDOUIN, B., LE CLOAREC, M.-F., RAMONET, M., Origins of

210

Po in the atmosphere at Lamto, Ivory Coast: Biomass burning and Saharan dusts,

Atmos. Environ.

30 (1996) 3705–3714.

[IV.8] NHO, E.-Y., LE CLOAREC, M.-F., ARDOUIN, B., RAMONET, M.,

210

Po, an

atmospheric tracer of long-range transport of volcanic plumes, Tellus

B 49

(1997)

429–438.

[IV.9] KOSTER, H.W., et al., Po-210, Pb-210, Ra-226 in Dutch aquatic ecosystems and

polders: Anthropogenic sources, distribution and radiation doses, Radiat. Prot.

Dosim. 45 (1992) 715–719.

[IV.10] YAN, G., CHO, H.-M., LEE, I., KIM, G., Significant emissions of

210

Po by coal

burning into the urban atmosphere of Seoul, Korea, Atmos. Environ.

54 (2012) 80–85.

[IV.11] KHATER, A.E.M., BAKR, W.F., Technologically enhanced

210

Pb and

210

Po in iron

and steel industry, J. Environ. Radioact.

102 (2011) 527–530.

[IV.12] TROTTI, F., et al., “A study concerning NORM in integrated steelworks”, Naturally

Occurring Radioactive Material (NORM

V) (Proc. Int. Symp. Seville, 2007), IAEA,

Vienna (2008)

351–354.

[IV.13] MORA, J.C., ROBLES, B., CORBACHO, J.A., GASCÓ, C., GÁSQUEZ, M.J.,

Modelling the behaviour of

210

Po in high temperature processes, J. Environ.

Radioact.

102 (2011) 520–526.

[IV.14] MORAWSKA, L., “Indoor particles, combustion products and fibres”, The Handbook

of Environmental Chemistry, Vol.

4, Part F: Indoor Air Pollution (PLUSCHKE, P.,

Ed.) Springer, Heidelberg (2004)

117–147.

[IV.15] GONÇALVES, C., et al., Characterisation of PM10 emissions from woodstove

combustion of common woods grown in Portugal, Atmos. Environ.

(2010)

4474–4480.

[IV.16] KILTHAU, G.F., Cancer risk in relation to radioactivity in tobacco, Radiol.

Technol.

67 (1996) 217–222.

[IV.17] PAPASTEFANOU, C., Radioactivity in tobacco leaves, J. Environ. Radioact. 53

(2001)

67–73.

[IV.18] CARVALHO, F.P., OLIVEIRA, J.M., Polonium in cigarette smoke and radiation

exposure of lungs, Czechoslov. J. Phys.

56 (2006) 697–703.

234

APPENDIX IV

[IV.19] DESIDERI, D., MELI, M.A., FEDUZI, L., ROSELLI, C.,

210

Po and

210

Pb inhalation

by cigarette smoking in Italy, Health Phys.

92 (2007) 58–63.

[IV.20] FOOD AND AGRICULTURE ORGANIZATION OF THE UNITED NATIONS,

Fire Management Global Assessment

2006, FAO Forestry Paper No. 151, FAO,

Rome

(2007).

[IV.21] SAARIKOSKI, S., HILLAMO, R., “Wildfires as a source of aerosol particles

transported to the Northern European regions”, The Handbook of Environmental

Chemistry, Vol. 26: Urban Air Quality in Europe (VIANA, M., Ed.), Springer,

Heidelberg (2012)

101–121.

[IV.22] LAZARIDIS, M., et al., Contribution of forest fire emissions to atmospheric pollution

in Greece, Air Qual. Atmos. Health

1 (2008) 143–158.

[IV.23] DE VOS, A.J.M., REISEN, F., COOK, A., DEVINE, B., WEINSTEIN, P.,

Respiratory irritants in Australian bushfire smoke: Air toxics sampling in a smoke

chamber and during prescribed burns, Arch. Environ. Contam. Toxicol.

(2009)

380–388.

[IV.24] CARVALHO, F.P., OLIVEIRA, J.M., MALTA, M., “Vegetation fires and release of

radioactivity into the air”, Environmental Health and Biomedicine (BREBBIA, C.A.,

et al., Eds), WIT Press, Southhampton (2011) 3–9.

[IV.25] DAMOAH, R., et al., Around the world in 17 days: Hemispheric-scale transport of

forest fire smoke from Russia in May

2003, Atmos. Chem. Phys. 4 (2004) 1311–1321.

[IV.26] PALIOURIS, G., TAYLOR, H.W., WEIN, R.W., SVOBODA, J., MIERZYNSKI, B.,

Fire as an agent in redistributing fallout

137

Cs in the Canadian boreal forest, Sci. Total

Environ.

160/161 (1995) 153–166.

[IV.27] AMIRO, B.D., SHEPPARD, S.C., JOHNSTON, F.L., EVENDEN, W.G., HARRIS,

D.R., Burning radionuclide question: What happens to iodine, cesium and chlorine in

biomass fires? Sci. Total Environ.

187 (1996) 93–103.

[IV.28] KASHPAROV, V.A., et al., Forest fires in the territory contaminated as a result of the

Chernobyl accident: Radioactive aerosol resuspension and exposure of fire-fighters,

J. Environ. Radioact.

51 (2000) 281–298.

[IV.29] YOSCHENKO, V.I., et al., Resuspension and redistribution of radionuclides during

grassland and forest fires in the Chernobyl exclusion zone — Part

II: Modeling the

transport process, J. Environ. Radioact.

87 (2006) 260–278.

[IV.30] YOSCHENKO, V.I., et al., Resuspension and redistribution of radionuclides during

grassland and forest fires in the Chernobyl exclusion zone — Part

I: Fire experiments,

J. Environ. Radioact.

86 (2006) 143–163.

[IV.31] MIN HAO, W., BONDARENKO, O.O., ZIBTSEV, S., HUTTON, D., “Vegetation

fires, smoke emissions, and dispersion of radionuclides in the Chernobyl exclusion

zone”, Developments in Environmental Science, Vol.

8 (BYTNEROWICZ, A.,

ARBAUGH, M., RIEBAU, A., ANDERSEN, C., Eds), Elsevier (2009)

265–275.

[IV.32] PAATERO, J., et al., Resuspension of radionuclides into the atmosphere due to forest

fires, J. Radioanal. Nucl. Chem.

282 (2009) 473–476.

[IV.33] JARGIN, S.V., Forest fires in the former Soviet Union: No reasons for radiophobia, J.

Environ. Radioact.

102 (2011) 218–219.

235

VEGETATION AND FOREST FIRES

[IV.34] COMMODORE, A.A., et al., Radioactivity in smoke particulates from prescribed

burns at the Savannah River Site and at selected southeastern United States forests,

Atmos. Environ.

54 (2012) 643–656.

[IV.35] CARVALHO, F.P., OLIVEIRA, J.M., MALTA, M., Forest fires and resuspension of

radionuclides into the atmosphere, Am. J. Environ. Sci.

8 (2012) 1–4.

[IV.36] CARVALHO, F.P., OLIVEIRA, J.M, MALTA, M., Exposure to radionuclides in

smoke from vegetation fires, Sci. Total Environ.

472 (2012) 421–424.

237

Appendix V

POLONIUM IN TOBACCO AND DOSES TO HUMANS

V.1. INTRODUCTION

Tobacco (Nicotina rustica) is in the Solonaceae family, together with bell

pepper (capsicum), tomato and potato, and there are many species in the tobacco

genus, Nicotiana. It was originally grown as a crop in Central America and is

named after the Caribbean island of Tobago. Most tobacco is smoked, but other

forms exist, such as snuff and chewing tobacco. Nicotine, the active substance in

tobacco, is present in lower doses in potato plants, tomato plants and, together

with cocaine, in coca plants. It is named after a French ambassador to Portugal,

Jean Nicot, who introduced snuff, taken through the nose, to the French court

in 1560.

V.2.

210

Pb AND

210

Po IN TOBACCO

Inhalation of tobacco smoke is ranked second to food as a source of

210

and

210

Po in humans (see Ref. [V.1] for a review), and there is a clear correlation

between lung cancer and smoking, and evidence of a synergistic effect when

smoking and elevated radon levels are combined

[V.2–V.4]. There are several

publications which deal with

210

Pb and

Po in tobacco and the transfer to humans

(see Table

V.1). Activity concentrations are in the range of 2.8–37 mBq/g. There

are few data on activity concentrations in pipe tobacco, cigar tobacco and snuff.

A cigarette contains about 1

g of tobacco.

When

210

Pb is also analysed, it is generally in radioactive equilibrium to

210

Po. In addition, while all results are of the same order of magnitude, the activity

concentrations of

210

Pb and

210

Po in tobacco vary depending on the geographical

region in which the tobacco is grown and harvested. If foliar deposition is the

main source for the radionuclides, the

210

Po:

210

Pb activity ratio should be lower

than one at harvest. However, owing to the time elapsed from harvest to purchase,

equilibrium has usually been attained in tobacco products.

238

APPENDIX V

V.3. TRANSFER TO HUMANS AND RADIOLOGICAL DOSES

The effective dose conversion factors for

210

Pb and

210

Po for inhalation

by adults are 1.1

× 10

−6

Sv/Bq and 3.3 × 10

−6

Sv/Bq, respectively [V.12]. The

dose conversion factors by ingestion are slightly lower, at 6.7 × 10

−7

Sv/Bq and

1.2 × 10

−6

Sv/Bq, respectively.

Additional studies from the 1960s include information on the radionuclide

contents in the wrapping paper, filters, cigarette ash, post-smoking filters and

cigarette ends, as well as in the mainstream and sidestream

smoke, and that

sometimes a smoking machine was used for the radionuclide budget in the

different matrices

[V.10, V.11] (see Table V.2).

There are large variations in the data. Apart from the variation of

radionuclide concentrations in tobacco, the authors often have different estimates

of the amount of smoke inhaled and the fraction of

210

Po retained in the lungs.

Different dose conversion factors have been used where recommendations

have changed over time. Sometimes, calculations are based on assumed lung

TABLE V.1. EXAMPLES OF

210

Po ACTIVITY CONCENTRATIONS

IN TOBACCO

Country

Activity per gram

(mBq/g)

Activity per cigarette

(mBq)

Ref.

Various 6.6–24 7.4–23 [V.4]

Egypt 12.6–24 9.7–21.8 [V.5]

Greece (tobacco leaves) 6–18 5–15 [V.6]

Poland 5.7–33 5.7–33.2 [V.7]

Portugal 2.8–37 2.6–27.7 [V.8]

Sweden (snuff) 6–60 —

[V.9]

USA

Cigarettes

Pipe tobacco

Cigar tobacco

13.3–20.7

10.5–22

7.6–13.9

10.5–30

12.6–17.8 [V.10]

—

USA

Pipe tobacco

Cigar tobacco

7–13

12–18

—

[V.11]

—: data not available.

239

POLONIUM IN TOBACCO

concentrations. The quality factor used to estimate dose from alpha radiation

(now 20) is not always given in these studies. As for less common exposure

routes, using tobacco as snuff gives a lower dose than smoking. It would be of

interest to study the impact of

210

Po derived from using water pipes (hookahs).

A fraction of the radionuclides will transfer to the smoker and will also

expose persons in the vicinity. Smokers have higher concentrations of these

radionuclides in their lungs but also in other organs, such as the liver and

kidneys (see Chapter

9). The dose to the lungs is much higher from smoking

TABLE V.2. ACTIVITY CONCENTRATIONS IN DIFFERENT

SMOKING MATRICES AND ANNUAL DOSES (SMOKING

20 CIGARETTES PER DAY) FROM

210

Study

Activity per gram

(mBq/g)

Dose

(μSv/a)

Skwarzek et al. [V.7]

Filter

Paper

Filter after smoking

Ash only

0.1–4.3

1.5

0.3–9.1

0.43–14.7

—

Ferri and Baratta [V.11]

Filter after smoking

Inhaled

End+ash

1.1–2.0

1.7–4.0

4.3–7.1

165–460 (lungs)

Khater [V.5]

Filter

Paper

Filter after smoking

Ash only

0.9–9.7

0.3–8.1

1.5–12.6

2.2–35

193 (effective dose)

Carvalho and Oliveira [V.8]

Inhaled

End+ash

0.97–1.52

0.3–4.9

420 (lungs)

Black and Bretthauer [V.10]

Inhaled

End only

2.6

7.0

—

Watson [V.4] —

5–220 (lung)

Papastefanou [V.6] —

47–135

(effective dose)

—: data not available.

240

APPENDIX V

than from breathing outdoor ambient air, and a correlation exists between the

prevalence of lung cancer in smokers with those living in houses with high radon

concentrations.

REFERENCES TO APPENDIX V

[V.1] FINKELSTEIN, M.M., Clinical measures, smoking, radon exposure, and risk of lung

cancer in uranium miners, Occup. Environ. Med.

53 (1996) 697–702.

[V.2] STEENLAND, K., Age specific interactions between smoking and radon among

United States uranium miners, Occup. Environ. Med.

51 (1994) 192–194.

[V.3] HARLEY, N.H., HARLEY, J.H., Potential lung cancer risk from indoor radon

exposure, CA: Cancer J. Clin.

40 (1990) 265–275.

[V.4] WATSON, A.P., Polonium-210 and Lead-210 in Food and Tobacco Products: A Review

of Parameters and an Estimate of Potential Exposure and Dose, ORNL/TM-8831, Oak

Ridge Natl Lab., TN

(1983).

[V.5] KHATER, A.E.M., Polonium-210 budget in cigarettes, J. Environ. Radioact. 71

(2004)

33–41.

[V.6] PAPASTEFANOU, C., Radioactivity of tobacco leaves and radiation doses induced

from smoking, Int. J. Environ. Res. Public Health

6 (2009) 558–567.

[V.7] SKWARZEK, B., STRUMIŃSKA, D.I., BORYŁO, A., ULATOWSKI, J., Polonium

210

Po in cigarettes produced in Poland, J. Environ. Sci. Health A 36 (2001) 465–474.

[V.8] CARVALHO, F.P., OLIVEIRA, J.M., Polonium in cigarette smoke and radiation

exposure of lungs, Czechoslov. J. Phys.

56 (2006) 697–703.

[V.9] SAMUELSSON, C., Medicinska konsekvenser av polonium i snus, Läkartidningen 86

(1989) 2290–2292.

[V.10] BLACK, S.C., BRETTHAUER, E.W., Polonium-210 in tobacco, Radiol. Health Data

Rep.

9 (1968) 145–152.

[V.11] FERRI, E.S., BARATTA, E.J., Polonium-210 in tobacco, cigarette smoke, and selected

human organs, Public Health Rep.

81 (1966) 121–127.

[V.12] INTERNATIONAL COMMISSION ON RADIOLOGICAL PROTECTION,

Age-dependent Doses to the Members of the Public from Intake of Radionuclides —

Part 5: Compilation of Ingestion and Inhalation Coefficients, Publication 10, Pergamon

Press, Oxford and New York (1995).

241

Appendix VI

ESTIMATION OF AEROSOL RESIDENCE TIMES

IN THE ATMOSPHERE

VI.1. AEROSOL RESIDENCE TIME

The residence time (τ) for a radionuclide in a parcel of air (or any other

environmental compartment) represents the average time an atom of that

radionuclide is present in the air parcel before it is removed

[VI.1]. If N is the total

number of atoms of the radionuclide in the air parcel, I is the rate of introduction

from sources (e.g. by ingrowth from a parent or transfer from another air parcel)

and R is the rate of removal (e.g. by radioactive decay, or wet or dry deposition),

then:

(VI.1)

If N is a constant, then steady state conditions hold, and I = R. In this case,

the residence time is given by:

(VI.2)

In the case of the total amount of

222

Rn in the atmosphere, the only

significant removal process is radioactive decay. Under the assumption that a

steady state condition applies, then:

5.5 days

= ==

(VI.3)

However, for a portion of the whole atmosphere, such as the planetary

boundary layer (PBL), the average residence time for

222

Rn will be less than this

due to mixing, which results in additional net removal from the PBL to the free

troposphere. These residence times are not long in comparison with characteristic

mixing times in the troposphere; as a result,

222

Rn is not well mixed either

horizontally or vertically.

The

222

Rn progeny,

210

Pb,

210

Bi and

210

Po, are attached to aerosols in the

atmosphere, and their behaviour reflects that of those aerosols. In the troposphere,

242

APPENDIX VI

aerosol residence times range from days to weeks, depending on the size of the

aerosol, height above ground level and the prevailing meteorological conditions.

In contrast, residence times in the stratosphere are of the order of one year

[VI.1].

The degrees of disequilibrium between

222

Rn and

210

Pb, and between

210

Pb and its radioactive progeny,

210

Bi and

210

Po, have sometimes been used

to calculate apparent atmospheric aerosol residence times

[VI.2–VI.4]. For this

purpose, concentrations of the pair (

222

Rn,

210

Pb) in air can be used [VI.5]. For

the pairs (

210

Pb,

210

Bi) and (

210

Pb,

210

Po), concentrations in air [VI.6–VI.8], in

precipitation

[VI.9–VI.11] and in dry fallout [VI.11] have been used.

In the case that

210

Pb,

210

Bi and

210

Po are produced solely by decay of

222

in the atmosphere, that their production and removal are in steady state at the

time and place of measurement, and that the removal term follows first order

kinetics, then:

( )

11 2 2 R

l ll

= - +=

(VI.4)

where 1 is a parent radionuclide, 2 is its immediate daughter, λ

and λ

are their radioactive decay constants, and λ

is the first order rate constant for

removal of aerosols by all processes. The mean aerosol residence time is:

(VI.5)

For the individual parent–progeny pairs, the apparent aerosol residence

time is given by

[VI.2, VI.5]:

( )

Pb Rn

Bi Pb Bi

Rn, Pb

Pb, Bi

Pb, Po

b b ac

æö

èø

æö

èø

-+ -

(VI.6)

(VI.7)

(VI.8)

243

AEROSOL RESIDENCE TIMES

where

Pb Po

Bi Po

;

and .

X XX

aA A

æö

=- +

èø

In each case, these should represent the residence time of the aerosols to

which

210

Pb is attached. In practice, the apparent residence times calculated from

these three radionuclide pairs often do not agree (e.g. see Refs

[VI.2, VI.5]).

Table

VI.1 shows a summary of apparent residence times calculated for

tropospheric aerosols from a number of studies

[VI.6, VI.12]. The residence

times are highly variable. In general, those obtained from the (

210

Pb,

210

Po) pair

are higher than for the other two pairs. However, relatively high residence times

are also sometimes obtained for the (

210

Pb,

210

Bi) pair.

Resuspension of dust can be an interfering factor in these calculations, since

210

Pb and its progeny will be at, or close to, secular equilibrium on this dust. This

factor can be greater for the larger aerosol size fractions

[VI.14, VI.23, VI.24].

It can also be greater for dry fallout than for precipitation

[VI.11]. One approach

to this problem is to use the determination of a member of the uranium decay

chain above

210

Pb to apply a correction to the measured

210

Pb,

210

Bi and

210

concentrations to allow calculation of corrected residence times

[VI.6, VI.10].

In the case of the (

210

Pb,

210

Po) pair, another factor is the influence of

significant sources of

210

Po from the Earth’s surface due to the volatility of

polonium, such as volcanic emissions, biomass burning and coal combustion (see

Chapter

5). The influence of such sources is expected to be much lower in the

case of (

210

Pb,

210

Bi). In addition, the mean residence time of

210

Bi (τ

= 7.2 d)

is more comparable to aerosol residence times in the troposphere than that of

210

Po. Consequently, a residence time calculation based on (

210

Pb,

210

Bi) is more

reliable than one based on (

210

Pb,

210

Po).

The apparent residence times obtained represent an average for the activity

in the air or precipitation sample. They can therefore represent an average of

residence times for several air masses which have mixed, such as a mixture

of upper tropospheric and lower tropospheric air with different residence

times

[VI.5, VI.24]. For precipitation samples, they can represent an average of

residence times for in-cloud rainout and below-cloud washout, and mixing of air

masses during the storm

[VI.9, VI.10].

244

APPENDIX VI

The use of the (

210

Pb,

210

Bi) and (

210

Pb,

210

Po) pairs relies on determination

of these radionuclides in air filter or precipitation samples. The

210

Bi and

210

determinations should be carried out as soon as practicable after collection of

the sample in order to minimize uncertainties resulting from decay and ingrowth

corrections

[VI.25]. When determining the sample

210

Po activity, ingrowth from

TABLE VI.1. APPARENT AEROSOL RESIDENCE TIMES OBTAINED FOR

ATMOSPHERIC OR PRECIPITATION SAMPLES

Location

(Rn, Pb)

(d)

(Pb, Bi)

(d)

(Pb, Po)

(d)

Ref.

Portugal

Savacem —

0.5–82 (5.7) 7–757 (33) [VI.6]

USA

Oak Ridge, TN

Boulder, CO

Argonne, IL

Madison, WI

Fayetteville, AR

Poker Flat, AK

Eagle, AK

—

2.2–3.4

—

4.8–15.3 (8.2)

1.59–13.0 (5.4)

6–67

—

2.4–25.6 (8.5)

3–240 (20)

—

11–77

33–66

9.6

—

2–320 (40)

11.9–32

0–38.9

[VI.12]

[VI.2, VI.5, VI.13]

[VI.14]

[VI.15]

[VI.16]

[VI.9]

[VI.7]

Switzerland

Jungfraujoch —

—

1–12 (6) [VI.17]

Germany

Freiburg —

—

20 [VI.18]

France

Gif-sur-Yvette 6.5

6.77

8.8–10.5

7–9

—

[VI.19]

United Kingdom

Milford Haven 6.77 —

40 [VI.20]

India

Mumbai —

8 —

[VI.21]

Egypt

El-Minia —

4.3–12.86 (9.83) —

[VI.22]

Source: Table 3 of Ref. [V.12].

Note: Averages in parentheses.

—: data not available.

245

AEROSOL RESIDENCE TIMES

both

210

Pb and

210

Bi between sample collection and analysis needs to be allowed

for. In this case, a separate

210

Bi determination can be used to avoid unquantifiable

errors in the ingrowth and decay correction

[VI.8, VI.26].

REFERENCES TO APPENDIX VI

[VI.1] SEINFELD, J.H., PANDIS, S.N., Atmospheric Chemistry and Physics: From Air

Pollution to Climate Change, 3rd edn, John Wiley and Sons, Hoboken, NJ (2006).

[VI.2] POET, S.E., MOORE, H.E., MARTELL, E.A., Lead 210, bismuth 210 and polonium

210 in the atmosphere: Accurate ratio measurement and application to aerosol

residence time determination, J. Geophys. Res.

77 (1972) 6515–6527.

[VI.3] TUREKIAN, K.K., NOZAKI, Y., BENNINGER, L.K., Geochemistry of atmospheric

radon and radon products, Annu. Rev. Earth Planet Sci.

5 (1977) 227–255.

[VI.4] RANGARAJAN, C., EAPEN, C.D., The use of natural radioactive tracers in a study

of atmospheric residence times, Tellus

B 42 (1990) 142–147.

[VI.5] MOORE, H.E., POET, S.E., MARTELL, E.A.,

222

Rn,

210

Pb,

210

Bi and

210

Po profiles

and aerosol residence times versus altitude, J. Geophys. Res.

78 (1973) 7065–7075.

[VI.6] CARVALHO, F.P., Origins and concentrations of

222

Rn,

210

Pb,

210

Bi and

210

Po in the

surface air at Lisbon, Portugal, at the Atlantic edge of the European continental

landmass, Atmos. Environ.

29 (1995) 1809–1819.

[VI.7] BASKARAN, M., SHAW, G.E., Residence time of arctic haze aerosols using the

concentrations and activity ratios of

210

Po,

210

Pb and

Be, J. Aerosol Sci. 32

(2001)

443–452.

[VI.8] DŁUGOSZ-LISIECKA, M., BEM, H., Determination of the mean aerosol residence

times in the atmosphere and additional

210

Po input on the base of simultaneous

determination of

Be,

Na,

210

Pb,

210

Bi and

210

Po in urban air, J. Radioanal. Nucl.

Chem.

293 (2012) 135–140.

[VI.9] GAVINI, M.B., BECK, J.N., KURODA, P.K., Mean residence times of the long-lived

radon daughters in the atmosphere, J. Geophys. Res.

79 (1974) 4447–4452.

[VI.10] MARTIN, P., Uranium and thorium series radionuclides in rainwater over several

tropical storms, J. Environ. Radioact.

65 (2003) 1–18.

[VI.11] McNEARY, D., BASKARAN, M., Residence times and temporal variations of

210

in aerosols and precipitation from southeastern Michigan, United States, J. Geophys.

Res.

112 D4 (2007).

[VI.12] PAPASTEFANOU, C., Residence time of tropospheric aerosols in association with

radioactive nuclides, Appl. Radiat. Isot.

64 (2006) 93–100.

[VI.13] MOORE, H.E., POET, S.E., MARTELL, E.A., “Tropospheric aerosol residence

times indicated by radon and radon-daughter concentrations”, Natural Radiation

Environment

II (Proc. 2nd Int. Symp. Houston, 1972) (ADAMS, J.A.S., LOWDER,

W.M., GESELL, T.F., Eds), United States Department of Commerce, Springfield, VA

(1972)

775–786.

246

APPENDIX VI

[VI.14] MARLEY, N.A., et al., Measurement of

210

Pb,

210

Po and

210

Bi in size-fractionated

atmospheric aerosols: An estimate of fine-aerosol residence times, Aerosol Sci.

Technol.

32 (2000) 569–583.

[VI.15] FRANCIS, C.W., CHESTERS, G., HASKIN, L.A., Determination of

210

Pb mean

residence time in the atmosphere, Environ. Sci. Technol.

4 (1970) 586–589.

[VI.16] FRY, L.M., MENON, K.K., Determination of the tropospheric residence time of

lead-210, Science

137 (1962) 994–995.

[VI.17] GÄGGELER, H.W., JOST, D.T., BALTENSPERGER, U., SCHWIKOWSKI, M.,

Radon and thoron decay product and

210

Pb measurements at Jungfraujoch,

Switzerland, Atmos. Environ.

29 (1995) 607–616.

[VI.18] LEHMANN, L., SITTKUS, A., Bestimmung von Aerosolverweilzeiten aus den RaD

und RaF-Gehalt der atmosphärischen Luft und des Niederschlages,

Naturwissenschaft

46 (1959) 9–10.

[VI.19] LAMBERT, G., et al., Cycle du radon et de ses descendants: Application à 1’étude

des échanges troposphère-stratosphère, Ann. Geophys.

38 (1982) 497–531.

[VI.20] PEIRSON, D.H., CAMBRAY, R.S., SPICER, G.S., Lead-210 and polonium-210 in

the atmosphere, Tellus

18 (1966) 427–433.

[VI.21] RANGARAJAN, C., A study of the mean residence time of the natural radioactive

aerosols in the planetary boundary layer, J. Environ. Radioact.

15 (1992) 193–206.

[VI.22] AHMED, A.A., MOHAMMED, A., ALI, A.E., EL-HUSSEIN, A., BARAKAT, A., A

study on aerosol residence time in El-Minia, Egypt, J. Aerosol Sci.

31 (2000) 470–471.

[VI.23] GAFFNEY, J.S., MARLEY, N.A., CUNNINGHAM, M.M., Natural radionuclides in

fine aerosols in the Pittsburgh area, Atmos. Environ.

38 (2004) 3191–3200.

[VI.24] BASKARAN, M., Po-210 and Pb-210 as atmospheric tracers and global atmospheric

Pb-210 fallout: A review, J. Environ. Radioact.

102 (2011) 500–513.

[VI.25] KIM, C.-K., MARTIN, P., FAJGELJ, A., Quantification of measurement uncertainty

in the sequential determination of

210

Pb and

210

Po by liquid scintillation counting and

alpha-particle spectrometry, Accredit. Qual. Assur.

13 (2008) 691–702.

[VI.26] LOZANO, R.L., SAN MIGUEL, E.G., BOLIVAR, J.P., Assessment of the influence

of in

situ

210

Bi in the calculation of in situ

210

Po in air aerosols: Implications on

residence time calculations using the

210

Po/

210

Pb activity ratios, J. Geophys.

Res.

116 (2011).

247

Appendix VII

USE OF

210

Po AS A RADIOTRACER IN BIOGENIC

PARTICULATE FLUX STUDIES IN THE OCEANS

VII.1. INTRODUCTION

Polonium-210 is present in the world’s oceans, and

210

Po sources and its

distribution and dynamics in the water column are described in more detail in

Chapter 8. The distribution of

210

Po in the water column of the open ocean is

generally similar in various oceanic regions. In the upper layer of the ocean,

210

Po activity concentration in water is lower than that of

210

Pb due to the

overriding importance of atmospheric deposition where

210

Po:

210

Pb is roughly

around 0.1. In the mesopelagic zone, under the euphotic/epipelagic layer,

210

is in excess relative to

210

Pb, probably due to release of

210

Po soluble forms from

sinking faecal pellets and biogenic debris originating from the upper layer. In the

bathypelagic and abyssopelagic zones,

210

Po concentrations are lower than

210

Pb,

and although continuously produced in the water from

226

Ra–

210

Pb radioactive

decay,

210

Po is also continuously removed by sinking particulate matter [VII.1].

The mean residence time of

210

Pb and

210

Po radionuclides can be determined

in the water column in several oceanic regions using the disequilibria between

226

Ra,

210

Pb and

210

Po. In general,

210

Po residence time in the upper layer of the

ocean is short, around 0.6–0.7 a

–1

[VII.1–VII.4]. This short residence time is

due to rapid

210

Po adsorption onto particles such as viruses, bacteria, plankton,

organic debris and inorganic particles that make the gravitational flux of settling

particles. Furthermore,

210

Po is taken up into the cells of phytoplankton and

becomes associated with soluble proteins in the cytosol [VII.5]. Zooplankton,

as phytoplankton grazers, easily absorb

210

Po with cytosolic proteins of ingested

phytoplankton, while metals, including

210

Pb, are segregated and remain mostly

associated with the ejected cell debris.

The grazing activity of zooplankton in the epipelagic zone processes and

removes particulate matter from the ocean surface layer, mostly packing it into

faecal pellets that rapidly sink throughout the water column reach the abyssal

sea-floor in a relatively short time [VII.6, VII.7]. The role of zooplankton in the

removal of

210

Po and other elements from the epipelagic layer of the ocean has

long been known and the associated transport of contaminants (e.g. plutonium

from fallout) in the oceans has been investigated [VII.8–VII.11].

248

APPENDIX VII

Besides phytoplankton and zooplankton, all organisms of marine food

chains in the upper layer of the ocean contribute through similar processes,

although not quantified yet, to concentrate

210

Po, starting from ionic

210

Po in the

soluble phase and gradually incorporating it into large particles and biota tissues.

With the seasons, and in line with the highs and lows of primary productivity in

the epipelagic layer,

210

Po removal rates and flux to the deep-sea with the sinking

particles vary throughout the year [VII.3, VII.12].

VII.2. USE OF

210

Po AS A RADIOTRACER

Studies on the biogenic particle flux to the deep ocean started in the 1960s

with the supply of organic carbon from the euphotic zone to the permanently dark

deep-sea layers and the sustainability of deep-sea food chains. In the 1970s, the

existence of hydrothermal vents and deep-sea food chains based on the synthesis

of organic matter by sulphur reducing bacteria were discovered, which sustain

localized deep-sea food chains that do not depend on sunlight to synthesize

organic compounds. For most of the ocean sea-floor, however, deep-sea life

depends on the organic matter synthesized in the euphotic zone and exported to

the deep sea [VII.13].

Investigations on the carbon cycle have highlighted the importance of the

oceans as a sink for atmospheric carbon. Phytoplankton in the oceans, similar to

forests on land, is a major sink for atmospheric CO

[VII.14, VII.15]. Dissolution

of atmospheric CO

in the ocean is followed by incorporation of CO

into organic

molecules through photosynthesis by phytoplankton, and this incorporation is

followed by organic carbon recycling in, and export from, the euphotic zone with

the sinking particle flux. This association of oceanic organic matter production

and particulate matter export, conceptualized as the ‘biological carbon pump’,

contributes to controlling CO

levels in the atmosphere and buffering the

global climate system [VII.14, VII.16]. Therefore, the downward transfer and

sequestration of carbon in the oceans is a central issue in global climate change.

Studies on particulate organic carbon, particulate inorganic carbon and

biomineral (e.g. calcium and silicon) transfers in the ocean looked into natural

tracers to help quantify transfer fluxes [VII.3, VII.7]. The behaviour of several

radionuclides from the uranium radioactive series in the oceans has been assessed,

especially

210

Pb,

210

Po and

234

Th, and their potential as natural radiotracers for

particle flux [VII.17].

Phytoplankton productivity is connected to enhanced

210

Po absorption from

sea water, and association of this naturally occurring radionuclide with the organic

matter cycle in the surface ocean layer has been established [VII.15, VII.18].

249

BIOGENIC PARTICULATE FLUX STUDIES

Progress has been made in the direct measurement of the biogenic particle

flux in the ocean, especially due to the use of drifting or moored particle

interceptor devices or sediment traps

[VII.6]. Determination of radionuclide

concentrations in samples of particulate material indicate a high association of

several particle reactive radionuclides, such as

210

Pb,

210

Po,

239+240

Pu and

234

Th,

with the total flux of particulate and detrital material

[VII.18–VII.20]. However,

recent work has shown a better correlation between

210

Po and particulate organic

carbon fluxes

[VII.3]. Owing to

210

Po binding to amino acids and incorporation

into proteins in living organisms, the use of

210

Po as a tracer for particulate

organic carbon flux in the ocean water column seems to have a scientific basis

and has recently been explored. Polonium-210 seems to offer a good indirect

means of computing the flux of organic particles and thus particulate organic

carbon in the ocean

[VII.3, VII.18–VII.20].

Although

210

Pb is also a particle reactive radionuclide, as is

234

Th,

210

Pb is adsorbed onto organic and inorganic surfaces but is not significantly

absorbed into marine organisms to the extent that

210

Po is (see Chapter 8). For

suspended particulate matter in coastal environments throughout the year, both

210

Pb and

210

Po total concentrations in water are positively correlated with the

load of suspended particulate matter (p

< 0.01). However,

210

Po concentration

in particulate matter is better correlated with chlorophyll-a (p

< 0.10), while

210

Pb was correlated with the inorganic particle residue (p < 0.01) but not with

chlorophyll-a

[VII.21]. All evidence suggests that

210

Po is bound, and mostly

incorporated in, living matter, while

210

Pb seems to undergo adsorption onto the

mineral and detrital particle fraction. Other naturally occurring radionuclides,

such as

234

Th, have been used in combination with

210

Po to determine the mean

residence time of particles in the surface ocean layer [VII.3]. It seems that

210

is a suitable tracer in oligotrophic conditions (non-bloom of phytoplankton),

while

210

Pb and

234

Th display a better correlation and are thus better tracers of the

inorganic fraction of particulate flux. Different natural radiotracers can provide

different information and, actually, the measurement of

210

Po and

234

Th might

be complementary to better understand the particle flux

[VII.3, VII.7, VII.19,

VII.20,

VII.22].

Further research could be on the relationship of radionuclides and

particulate organic carbon in particles of different size classes, and on the binding

and partitioning (K

) of radiotracers and organic particles. Further studies on

210

Po binding adsorption and absorption by microscopic life forms in the ocean

are still needed, as are studies on the release rates of

210

Po from sinking particles

(carcasses and faecal pellets).

Determination of

210

Po in association with other parameters is a tool to

assess the organic particle flux in various regions of the ocean, and it is expected

250

APPENDIX VII

that by using these natural radiotracers a global picture of the biominerals and

particulate organic carbon fluxes in the world’s oceans will emerge.

REFERENCES TO APPENDIX VII

[VII.1] BACON, M.P., SPENCER, D.W., BREWER, P.G.,

210

Pb/

226

Ra and

210

Po/

210

disequilibria in seawater and suspended particulate matter, Earth Planet. Sci. Lett.

(1976)

277–296.

[VII.2] MASQUÉ, P., SANCHEZ-CABEZA, J.A., BRUACH, J.M., PALACIOS, E.,

CANALS, M., Balance and residence times of

210

Pb and

210

Po in surface waters of

the northwestern Mediterranean Sea, Cont. Shelf Res.

22 (2002) 2127–2146.

[VII.3] LE MOIGNE, F.A.C., et al., Export of organic carbon and biominerals derived from

234

Th and

210

Po at the Porcupine Abyssal Plain, Deep Sea Res. Part I 72

(2013)

88–101.

[VII.4] CARVALHO, F.P., “Uranium series radionuclides in the water column of the

Northeast Atlantic Ocean”, Isotopes in Hydrology, Marine Ecosystems and Climate

Change Studies (Proc. Int. Symp. Monaco, 2011), Vol.

2, IAEA, Vienna

(2013)

427–441.

[VII.5] REINFELDER, J.R., FISHER, N.S., The assimilation of elements ingested by

marine plankton, Science

251 (1991) 794–796.

[VII.6] FOWLER, S.W., KNAUER, G.A., Role of large particles in the transport of elements

and organic compounds through the oceanic water column, Prog. Oceanogr.

(1986)

147–194.

[VII.7] VERDENY, E., et al., POC export from ocean surface waters by means of

234

Th/

238

and

210

Po/

210

Pb disequilibria: A review of the use of two radiotracer pairs, Deep Sea

Res. Part

II 56 (2009) 1502–1518.

[VII.8] CHERRY, R.D., FOWLER, S.W., BEASLEY, T.M., HEYRAUD, M., Polonium-210:

Its vertical oceanic transport by zooplankton metabolic activity, Mar. Chem.

(1975)

105–110.

[VII.9] BEASLEY, T.M., HEYRAUD, M., HIGGO, J.J.W., CHERRY, R.D., FOWLER,

S.W.,

210

Po and

210

Pb in zooplankton fecal pellets, Mar. Biol. 44 (1978) 325–328.

[VII.10] HIGGO, J.J.W., CHERRY, R.D., HEYRAUD, M., FOWLER, S.W., BEASLEY,

T.M., “Vertical oceanic transport of alpha-radioactive nuclides by zooplankton fecal

pellets”, Natural Radiation Environment

III (Proc. Symp. Houston, 1978), Vol. 1

(GESELL, T.F., LOWDER, W.M., Eds), Department of Energy, Oak Ridge, TN

(1980)

502–513.

[VII.11] LINDAHL, P., LEE, S.-H., WORSFOLD, P., KEITH-ROACH, M., Plutonium

isotopes as tracers for ocean processes: A review, Mar. Environ. Res.

(2010)

73–84.

[VII.12] MARTIN, P., et al., Export and mesopelagic particle flux during a North Atlantic

spring diatom bloom, Deep Sea Res. Part

I 58 (2011) 338–349.

251

BIOGENIC PARTICULATE FLUX STUDIES

[VII.13] MILLER, C.B., WHEELER, P.A., Biological Oceanography, Wiley and Blackwell,

New York

(2012).

[VII.14] EPPLEY, R.W., PETERSON, B.J., Particulate organic matter flux and planktonic

new production in the deep ocean, Nature

282 (1979) 677–680.

[VII.15] RILEY, J.S, et al., The relative contribution of fast and slow sinking particles to

ocean carbon export, Global Biogeochem. Cycles

26 (2012).

[VII.16] HENSON, S., et al., A reduced estimate of the strength of the ocean’s biological

carbon pump, Geophys. Res. Lett.

38 (2011).

[VII.17] COCHRAN, J.K., MASQUÉ, P., Short-lived U/Th series radionuclides in the ocean:

Tracers for scavenging rates, export fluxes and particle dynamics, Rev. Mineral.

Geochem.

52 (2003) 461–492.

[VII.18] FRIEDRICH, J., RUTGERS VAN DER LOEFF, M.M., A two-tracer (

210

Po–

234

Th)

approach to distinguish organic carbon and biogenic silica export flux in the

Antarctic Circumpolar Current, Deep Sea Res. Part

I 49 (2002) 101–120.

[VII.19] STEWART, G.M., et al., Comparing POC export from

234

Th/

238

U and

210

Po/

210

disequilibria with estimates from sediment traps in the northwest Mediterranean,

Deep Sea Res. Part

I 54 (2007) 1549–1570.

[VII.20] STEWART, G.M., et al., Exploring the connection between

210

Po and organic matter

in the northwestern Mediterranean, Deep Sea Res. Part I 54 (2007) 415–427.

[VII.21] CARVALHO, F.P., OLIVEIRA, J.M., MALTA, M., “Vegetation fires and release of

radioactivity into the air”, Environmental Health and Biomedicine (BREBBIA,

C.A., et al., Eds), WIT Press, Southhampton (2011)

3–9.

[VII.22] STEWART, G.M., MORAN, S.B., LOMAS, M.W., Seasonal POC fluxes at BATS

estimated from

210

Po deficits, Deep Sea Res. Part I 57 (2010) 113–124.

253

ABBREVIATIONS

APDC ammonium pyrrolidine dithiocarbamate

CR concentration ratio

DCC dose conversion coefficient

DW dry weight

FW fresh weight

ICRP International Commission on Radiological Protection

distribution coefficient

NORM naturally occurring radioactive material

PBL planetary boundary layer

TF transfer factor

255

CONTRIBUTORS TO DRAFTING AND REVIEW

Carvalho, F. University of Lisbon, Portugal

Chambers, D. SENES Consultants Ltd, Canada

Fernandes, S. SENES Consultants Ltd, Canada

Fesenko, S. International Atomic Energy Agency

Holm, E. University of Gothenburg, Sweden

Howard, B. Centre for Ecology and Hydrology, United Kingdom

Martin, P. International Atomic Energy Agency

Phaneuf, M. International Atomic Energy Agency

Porcelli, D. University of Oxford, United Kingdom

Pröhl, G. International Atomic Energy Agency

Twining, J. Consultant

Vandenhove, H. SCK•CEN, Belgian Nuclear Research Centre, Belgium

Consultants Meetings

Vienna, Austria: 22–26 November 2010, 16–20 May 2011, 20–24 February 2012

No. 25

ORDERING LOCALLY

In the following countries, IAEA priced publications may be purchased from the sources listed below or

from major local booksellers.

Orders for unpriced publications should be made directly to the IAEA. The contact details are given at

the end of this list.

CANADA

Renouf Publishing Co. Ltd

22-1010 Polytek Street, Ottawa, ON K1J 9J1, CANADA

Telephone: +1 613 745 2665  Fax: +1 643 745 7660

Email: [email protected]  Web site: www.renoufbooks.com

Bernan / Rowman & Littlefield

15200 NBN Way, Blue Ridge Summit, PA 17214, USA

Tel: +1 800 462 6420 • Fax: +1 800 338 4550

Email: [email protected] Web site: www.rowman.com/bernan

CZECH REPUBLIC

Suweco CZ, s.r.o.

Sestupná 153/11, 162 00 Prague 6, CZECH REPUBLIC

Telephone: +420 242 459 205  Fax: +420 284 821 646

Email: [email protected]  Web site: www.suweco.cz

FRANCE

Form-Edit

5 rue Janssen, PO Box 25, 75921 Paris CEDEX, FRANCE

Telephone: +33 1 42 01 49 49  Fax: +33 1 42 01 90 90

Email: [email protected]  Web site: www.form-edit.com

GERMANY

Goethe Buchhandlung Teubig GmbH

Schweitzer Fachinformationen

Willstätterstrasse 15, 40549 Düsseldorf, GERMANY

Telephone: +49 (0) 211 49 874 015  Fax: +49 (0) 211 49 874 28

Email: [email protected]  Web site: www.goethebuch.de

INDIA

Allied Publishers

1st Floor, Dubash House, 15, J.N. Heredi Marg, Ballard Estate, Mumbai 400001, INDIA

Telephone: +91 22 4212 6930/31/69  Fax: +91 22 2261 7928

Email: [email protected]  Web site: www.alliedpublishers.com

Bookwell

3/79 Nirankari, Delhi 110009, INDIA

Telephone: +91 11 2760 1283/4536

Email: [email protected]  Web site: www.bookwellindia.com

ITALY

Libreria Scientifica “AEIOU”

Via Vincenzo Maria Coronelli 6, 20146 Milan, ITALY

Telephone: +39 02 48 95 45 52  Fax: +39 02 48 95 45 48

Email: [email protected]  Web site: www.libreriaaeiou.eu

JAPAN

Maruzen-Yushodo Co., Ltd

10-10 Yotsuyasakamachi, Shinjuku-ku, Tokyo 160-0002, JAPAN

Telephone: +81 3 4335 9312  Fax: +81 3 4335 9364

Email: [email protected]  Web site: www.maruzen.co.jp

RUSSIAN FEDERATION

Scientific and Engineering Centre for Nuclear and Radiation Safety

107140, Moscow, Malaya Krasnoselskaya st. 2/8, bld. 5, RUSSIAN FEDERATION

Telephone: +7 499 264 00 03  Fax: +7 499 264 28 59

Email: [email protected]  Web site: www.secnrs.ru

UNITED STATES OF AMERICA

Bernan / Rowman & Littlefield

15200 NBN Way, Blue Ridge Summit, PA 17214, USA

Tel: +1 800 462 6420 • Fax: +1 800 338 4550

Email: [email protected]  Web site: www.rowman.com/bernan

Renouf Publishing Co. Ltd

812 Proctor Avenue, Ogdensburg, NY 13669-2205, USA

Telephone: +1 888 551 7470  Fax: +1 888 551 7471

Email: [email protected]  Web site: www.renoufbooks.com

Orders for both priced and unpriced publications may be addressed directly to:

Marketing and Sales Unit

International Atomic Energy Agency

Vienna International Centre, PO Box 100, 1400 Vienna, Austria

Telephone: +43 1 2600 22529 or 22530 • Fax: +43 1 2600 29302 or +43 1 26007 22529

Email: [email protected] • Web site: www.iaea.org/books

16-49041

RELATED PUBLICATIONS

www.iaea.org/books

HANDBOOK OF PARAMETER VALUES FOR THE

PREDICTION OF RADIONUCLIDE TRANSFER IN

TERRESTRIAL AND FRESHWATER ENVIRONMENTS

Technical Reports Series No. 472

STI/DOC/010/472 (194 pp.; 2010)

ISBN 978–92–0–113009–9 Price: €45.00

THE ENVIRONMENTAL BEHAVIOUR OF RADIUM:

REVISED EDITION

Technical Reports Series No. 476

STI/DOC/010/476 (267 pp.; 2014)

ISBN 978–92–0–143310–7 Price: €52.00

HANDBOOK OF PARAMETER VALUES FOR THE

PREDICTION OF RADIONUCLIDE TRANSFER

TO WILDLIFE

Technical Reports Series No. 479

STI/DOC/010/479 (211 pp.; 2014)

ISBN 978–92–0–100714–8 Price: €55.00

Atoms for Peace

INTERNATIONAL ATOMIC ENERGY AGENCY

VIENNA

ISBN 978–92–0–112116–5

ISSN 0074–1914

Polonium-210 is an alpha emitting radionuclide

with no radioactive progeny and produces

only very-low-intensity gamma rays at very low

abundance. This means doses largely arise

from internal exposure. In addition to the

relatively high ingestion dose coefficient of

210

Po, radionuclide transfer in the environment

results in high activity concentrations in certain

foods. This publication focuses on radionuclide

transfers in terrestrial, freshwater and marine

environments, and provides information on key

transfer processes, concepts and models.

The Environmental Behaviour of Poloniumtechnical reportS series no. 484